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    INTERNATIONAL PROGRAMME ON CHEMICAL SAFETY


    ENVIRONMENTAL HEALTH CRITERIA 25





   SELECTED RADIONUCLIDES
                   
   TRITIUM 
   CARBON-14 
   KRYPTON-85
   STRONTIUM-90 
   IODINE 
   CAESIUM-137
   RADON 
   PLUTONIUM





    This report contains the collective views of an international group of
    experts and does not necessarily represent the decisions or the stated
    policy of the United Nations Environment Programme, the International
    Labour Organisation, or the World Health Organization.

    Published under the joint sponsorship of
    the United Nations Environment Programme,
    the International Labour Organisation,
    and the World Health Organization

    World Health Orgnization
    Geneva, 1983


         The International Programme on Chemical Safety (IPCS) is a
    joint venture of the United Nations Environment Programme, the
    International Labour Organisation, and the World Health
    Organization. The main objective of the IPCS is to carry out and
    disseminate evaluations of the effects of chemicals on human health
    and the quality of the environment. Supporting activities include
    the development of epidemiological, experimental laboratory, and
    risk-assessment methods that could produce internationally
    comparable results, and the development of manpower in the field of
    toxicology. Other activities carried out by the IPCS include the
    development of know-how for coping with chemical accidents,
    coordination of laboratory testing and epidemiological studies, and
    promotion of research on the mechanisms of the biological action of
    chemicals.


        ISBN 92 4 154085 0    

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CONTENTS
                                                             Paragraphs

ENVIRONMENTAL HEALTH CRITERIA FOR SELECTED RADIONUCLIDES

      PREFACE  . . . . . . . . . . . . . . . . . . . . . .    1 -   6

I.    INTRODUCTION . . . . . . . . . . . . . . . . . . . .    7 -  22

II.   TRITIUM  . . . . . . . . . . . . . . . . . . . . . .   23 -  77
      A.  INTRODUCTION . . . . . . . . . . . . . . . . . .   23 -  25
      B.  SOURCES  . . . . . . . . . . . . . . . . . . . .   26 -  57
          1. Natural tritium  . . . . . . . . . . . . . .   26 -  29
          2. Nuclear explosions . . . . . . . . . . . . .   30 -  33
          3. Nuclear fuel cycle . . . . . . . . . . . . .   34 -  51
          4. Tritium production plants  . . . . . . . . .   52 -  54
          5. Consumer products  . . . . . . . . . . . . .   55 -  56
          6. Controlled thermonuclear reactors  . . . . .      57
      C.  BEHAVIOUR IN THE ENVIRONMENT . . . . . . . . . .   58 -  61
          1. Natural and fallout tritium  . . . . . . . .   58 -  59
          2. Industrial releases  . . . . . . . . . . . .   60 -  61
      D.  TRANSFER TO MAN  . . . . . . . . . . . . . . . .   62 -  63
      E.  DOSIMETRY  . . . . . . . . . . . . . . . . . . .   64 -  77
          1. Dose per unit intake . . . . . . . . . . . .   64 -  66
          2. Dose per unit release  . . . . . . . . . . .   67 -  77
      F.  REFERENCES

III.  CARBON-14  . . . . . . . . . . . . . . . . . . . . .   78 - 112
      A.  INTRODUCTION . . . . . . . . . . . . . . . . . .   78 -  80
      B.  SOURCES  . . . . . . . . . . . . . . . . . . . .   81 -  98
          1. Natural carbon-14  . . . . . . . . . . . . .      81
          2. Nuclear explosions . . . . . . . . . . . . .   82 -  84
          3. Nuclear fuel cycle . . . . . . . . . . . . .   85 -  98
      C.  BEHAVIOUR IN THE ENVIRONMENT . . . . . . . . . .   99 - 102
      D.  TRANSFER TO MAN  . . . . . . . . . . . . . . . .  103 - 105
      E.  DOSIMETRY  . . . . . . . . . . . . . . . . . . .  106 - 112
          1. Dose per unit intake . . . . . . . . . . . .  106 - 107
          2. Dose per unit release  . . . . . . . . . . .  108 - 112
      F.  REFERENCES

IV.   KRYPTON-85 . . . . . . . . . . . . . . . . . . . . .  113 - 150
      A.  INTRODUCTION . . . . . . . . . . . . . . . . . .  113 - 117
      B.  SOURCES  . . . . . . . . . . . . . . . . . . . .  118 - 128
          1. Natural krypton-85 . . . . . . . . . . . . .     121
          2. Nuclear explosions . . . . . . . . . . . . .  122 - 123
          3. Nuclear fuel cycle . . . . . . . . . . . . .  124 - 128
      C.  BEHAVIOUR IN THE ENVIRONMENT . . . . . . . . . .  129 - 137
          1. Dispersion in the atmosphere . . . . . . . .  130 - 133
          2. Removal from the atmosphere  . . . . . . . .  134 - 137
      D.  TRANSFER TO MAN  . . . . . . . . . . . . . . . .  138 - 141
      E.  DOSIMETRY  . . . . . . . . . . . . . . . . . . .  142 - 150
          1. Dose per unit exposure . . . . . . . . . . .  143 - 144
          2. Dose per unit release  . . . . . . . . . . .  145 - 150
      F.  REFERENCES

V.    STRONTIUM-90 . . . . . . . . . . . . . . . . . . . .  151 - 211
      A.  INTRODUCTION . . . . . . . . . . . . . . . . . .  151 - 154
      B.  SOURCES  . . . . . . . . . . . . . . . . . . . .  155 - 165
          1. Nuclear explosions . . . . . . . . . . . . .  155 - 156
          2. Nuclear fuel cycle . . . . . . . . . . . . .  157 - 165
      C.  BEHAVIOUR IN THE ENVIRONMENT . . . . . . . . . .  166 - 185
          1. Movement in soil . . . . . . . . . . . . . .     166
          2. Transfer to plants . . . . . . . . . . . . .  167 - 171
          3. Transfer to milk . . . . . . . . . . . . . .     172
          4. Transfer to diet . . . . . . . . . . . . . .  173 - 181
          5. Aquatic behaviour  . . . . . . . . . . . . .  182 - 185
      D.  TRANSFER TO MAN  . . . . . . . . . . . . . . . .  189 - 192
      E.  DOSIMETRY  . . . . . . . . . . . . . . . . . . .  193 - 211
          1. Dose per unit intake . . . . . . . . . . . .  193 - 197
          2. Dose per unit release  . . . . . . . . . . .  198 - 211
      F.  REFERENCES

VI.   IODINE . . . . . . . . . . . . . . . . . . . . . . .  212 - 269
      A.  INTRODUCTION . . . . . . . . . . . . . . . . . .  212 - 214
      B.  SOURCES  . . . . . . . . . . . . . . . . . . . .  215 - 234
          1. Natural production . . . . . . . . . . . . .  215 - 216
          2. Nuclear explosions . . . . . . . . . . . . .  217 - 220
          3. Nuclear fuel cycle . . . . . . . . . . . . .  221 - 234
      C.  BEHAVIOUR IN THE ENVIRONMENT . . . . . . . . . .  235 - 255
          1. Nuclear explosions . . . . . . . . . . . . .  235 - 241
          2. Industrial releases  . . . . . . . . . . . .  242 - 255
      D.  TRANSFER TO MAN  . . . . . . . . . . . . . . . .  256 - 259
      E.  DOSIMETRY  . . . . . . . . . . . . . . . . . . .  260 - 270
          1. Dose per unit intake . . . . . . . . . . . .  260 - 261
          2. Dose per unit release  . . . . . . . . . . .  262 - 270
      F.  REFERENCES

VII.  CAESIUM-137  . . . . . . . . . . . . . . . . . . . .  271 - 336
      A.  INTRODUCTION . . . . . . . . . . . . . . . . . .  271 - 274
      B.  SOURCES  . . . . . . . . . . . . . . . . . . . .  275 - 282
          1. Nuclear explosions . . . . . . . . . . . . .  275 - 276
          2. Nuclear fuel cycle . . . . . . . . . . . . .  277 - 282
      C.  BEHAVIOUR IN THE ENVIRONMENT . . . . . . . . . .  283 - 309
          1. Fixation in soil . . . . . . . . . . . . . .  283 - 286
          2. Transfer to plants . . . . . . . . . . . . .  287 - 290
          3. Transfer to milk . . . . . . . . . . . . . .     291
          4. Transfer to meat . . . . . . . . . . . . . .     292
          5. Transfer to diet . . . . . . . . . . . . . .  293 - 301
          6. The lichen-caribou-man foodchain . . . . . .  302 - 303
          7. Aquatic behaviour  . . . . . . . . . . . . .  304 - 309
      D.  TRANSFER TO MAN  . . . . . . . . . . . . . . . .  310 - 319
          1. Absorption and distribution in tissues . . .  310 - 314
          2. Retention half-time  . . . . . . . . . . . .  315 - 317
          3. Transfer factor  . . . . . . . . . . . . . .  318 - 319
      E.  DOSIMETRY  . . . . . . . . . . . . . . . . . . .  320 - 336
          1. Dose per unit intake . . . . . . . . . . . .  320 - 324
          2. Dose per unit release  . . . . . . . . . . .  325 - 336
      F.  REFERENCES

VIII. RADON  . . . . . . . . . . . . . . . . . . . . .  . . 337 - 395
      A.  INTRODUCTION . . . . . . . . . . . . . . . . . .  337 - 340
      B.  SOURCES  . . . . . . . . . . . . . . . . . . . .  341 - 351
          1. Outdoors . . . . . . . . . . . . . . . . . .  341 - 344
          2. Indoors  . . . . . . . . . . . . . . . . . .  345 - 351
      C.  BEHAVIOUR IN THE ENVIRONMENT . . . . . . . . . .  352 - 375
          1. Release from soil  . . . . . . . . . . . . .  352 - 355
          2. Dispersion in air  . . . . . . . . . . . . .  356 - 361
          3. Indoor behaviour . . . . . . . . . . . . . .  362 - 365
          4. Radon daughter concentrations  . . . . . . .  366 - 375
      D.  TRANSFER TO MAN  . . . . . . . . . . . . . . . .  376 - 380
      E.  DOSIMETRY  . . . . . . . . . . . . . . . . . . .  381 - 395
          1. Dose per unit exposure . . . . . . . . . . .  381 - 393
          2. Dose per unit release  . . . . . . . . . . .  394 - 395
      F.  REFERENCES

IX.   PLUTONIUM  . . . . . . . . . . . . . . . . . . . . .  396 - 456
      A.  INTRODUCTION . . . . . . . . . . . . . . . . . .  396 - 401
      B.  SOURCES  . . . . . . . . . . . . . . . . . . . .  402 - 411
          1. Nuclear explosions . . . . . . . . . . . . .  402 - 404
          2. Nuclear fuel cycle . . . . . . . . . . . . .  405 - 406
          3. Other sources  . . . . . . . . . . . . . . .  407 - 411
      C.  BEHAVIOUR IN THE ENVIRONMENT . . . . . . . . . .  412 - 434
          1. Movement in soil . . . . . . . . . . . . . .  412 - 416
          2. Transfer to plants . . . . . . . . . . . . .  417 - 418
          3. Transfer to animals  . . . . . . . . . . . .  419 - 420
          4. Transfer to diet . . . . . . . . . . . . . .  421 - 425
          5. Aquatic behaviour  . . . . . . . . . . . . .  426 - 434
      D.  TRANSFER TO MAN  . . . . . . . . . . . . . . . .  435 - 443
      E.  DOSIMETRY  . . . . . . . . . . . . . . . . . . .  444 - 456
          1. Dose per unit intake . . . . . . . . . . . .  444 - 448
          2. Dose per unit release  . . . . . . . . . . .  449 - 456
      F.  REFERENCES

X.    RADIATION EFFECTS  . . . . . . . . . . . . . . . . .  457 - 476
      A.  SOMATIC EFFECTS  . . . . . . . . . . . . . . . .  459 - 463
          1. Early somatic effects  . . . . . . . . . . .  459 - 461
          2. Late somatic effects . . . . . . . . . . . .  462 - 463
      B.  GENETIC EFFECTS  . . . . . . . . . . . . . . . .  464 - 465
      C.  DOSE-RESPONSE RELATIONSHIPS  . . . . . . . . . .  466 - 469
      D.  RISK ESTIMATES . . . . . . . . . . . . . . . . .  470 - 476

XI.   CONCLUSIONS  . . . . . . . . . . . . . . . . . . . .  477 - 491
      A.  RADIONUCLIDES AND THE ENVIRONMENT  . . . . . . .  477 - 481
      B.  DOSE ASSESSMENTS . . . . . . . . . . . . . . . .  482 - 487
      C.  EFFECTS EVALUATION . . . . . . . . . . . . . . .  488 - 491

XII.  ANNEX
      EXCERPTS FROM "BASIC SAFETY STANDARDS FOR RADIATION 
      PROTECTION 1982 EDITION"                   

NOTE TO READERS OF THE CRITERIA DOCUMENTS

    While every effort has been made to present information in the 
criteria documents as accurately as possible without unduly 
delaying their publication, mistakes might have occurred and are 
likely to occur in the future.  In the interest of all users of the 
environmental health criteria documents, readers are kindly 
requested to communicate any errors found to the Division of 
Environmental Health, World Health Organization, Geneva, 
Switzerland, in order that they may be included in corrigenda which 
will appear in subsequent volumes.

    In addition, experts in any particular field dealt with in the 
criteria documents are kindly requested to make available to the 
WHO Secretariat any important published information that may have 
inadvertently been omitted and which may change the evaluation of 
health risks from exposure to the environmental agent under 
examination, so that the information may be considered in the event 
of updating and re-evaluation of the conclusions contained in the 
criteria documents.

ENVIRONMENTAL HEALTH CRITERIA FOR SELECTED RADIONUCLIDES

    At the request of the United Nations Environment Programme 
(UNEP),  the United Nations Scientific Committee on the Effects of 
Atomic Radiation (UNSCEAR) prepared a paper on the Environmental 
Behaviour and Dosimetry of Radionuclides. In accordance with the 
UNEP proposal, the paper, which was prepared during the 27th - 29th 
sessions of the Committee and was completed and approved at the 
30th session in 1981, is now being published in the WHO/UNEP 
Environmental Health Criteria series.  The EHC document, which is 
entitled "Selected Radionuclides", comprises the integral report 
prepared and edited by UNSCEAR, together with an annex consisting 
of excerpts taken from "Basic Safety Standards for Radiation 
Protection 1982 Edition", Safety Series No 9, a document prepared 
jointly by IAEA/ILO/NEA(OECD)/WHO, and published by the 
International Atomic Energy Agency, to give guidance to the 
appropriate national authorities on the establishment of limits for 
radionuclides.  The selected radionuclides discussed in the 
Environmental Health Criteria document are those of environmental 
importance for the general population and radiation workers. 

    Dr E. Komarov, Environmental Health Division, World Health 
Organization, was responsible for the final layout of the 
Environmental Health Criteria document. 

    The assistance of Dr B.G. Bennett (Monitoring and Assessment 
Research Centre, MARC) in the scientific editing of the 
Environmental Health Criteria document is gratefully acknowledged. 

    The contents of the 1982 UNSCEAR report to the General Assembly 
of the United Nations were taken into account during the 
preparation of the paper on the Environmental Behaviour and 
Dosimetry of Radionuclides, but the report was not quoted as it had 
not been issued at that time. 

ENVIRONMENTAL BEHAVIOUR AND DOSIMETRY OF RADIONUCLIDES

1.  PREFACE

1.  The release of radioactive materials to the environment 
potentially exposes populations to ionizing radiation and increases 
the risk of incurring deleterious health effects. The associations 
of released amounts to effects establish the health criteria for 
radionuclides, i.e., the quantitative relationships that would be 
required to establish release limits governing the management of 
radioactive materials used by man. 

2.  This report has been prepared by the United Nations Scientific 
Committee on the Effects of Atomic Radiation (UNSCEAR) for the 
United Nations Environment Programme (UNEP) to provide background 
information in establishing such health criteria.  In this report 
the more general considerations of environmental behaviour of 
several radionuclides are discussed, including sources, transport 
to man and dosimetry. The radionuclides discussed are those most 
frequently released from natural and man-made sources and the 
greatest contributors to population radiation exposure under normal 
circumstances. 

3.  The compilation of the relevant information is based largely on 
the detailed presentations and evaluations of the sources of 
ionizing radiation by UNSCEAR in its reports to the United Nations 
General Assembly.  The reader is referred to these reports for 
general concepts and for assessments of the dose commitments to man 
from exposures to sources such as natural radioactivity, fallout 
from atmospheric nuclear testing, releases from nuclear power 
production, occupational and medical irradiations. 

4.  Further information to be considered in establishing health 
criteria for radionuclides is that on health effects of 
irradiations.  The relationships between radiation dose and risks 
of health effects in man have recently been re-evaluated based on 
the available data.  This information can be found in the 1977 
report of UNSCEAR.  Only a brief summary of the general aspects of 
radiation effects and of radiation protection considerations is 
presented here. 

5.  The establishment of release limits for radionuclides in 
particular situations cannot be accomplished without rather more 
detailed considerations of the local and regional environment and 
the special pathways of transfer to man.  With this in mind, it is 
recognized that the material given here can only serve as 
background guidance. 

6.  The following scientists have contributed in the preparation of 
this report:  Dr. W.J. Bair, Dr. D. Beninson, Dr. B.G. Bennett, Dr. 
A. Bouville, Dr. P. Patek, Dr. G. Silini and Dr. J.O. Snihs. 

I.  INTRODUCTION

7.  Radionuclides are a special class of environmental substances.  
They are the unstable configurations of chemical elements which 
undergo radioactive decay, emitting radiation in the form of alpha 
or beta particles and x or gamma rays. The interaction of radiation 
with biological materials causes energy to be released to these 
materials which may result in a variety of harmful effects.  
Radiation is thus a potential hazard to man, although it may also 
be used in many beneficial ways, as in medical diagnosis and 
treatment, in industrial and consumer products and in the 
generation of electricity with nuclear reactors. 

8.  The realization of the harmful potential of ionizing radiation, 
which was dramatically brought to the attention of the public by 
the atomic bombing of Hiroshima and Nagasaki in 1945, was the cause 
of considerable attention that has been paid throughout the years 
to the effects of radiation.  As a result of these studies, a great 
deal is now known about radionuclide behaviour in the environment 
and in man and about the somatic and genetic consequences of 
irradiation.  This information surpasses by far that relating to 
any other class of environmental pollutants. 

9.  Considerable experience has been gained in environmental 
radiation measurements, particularly in tracing the movement of 
fallout radionuclides produced in atmospheric testing of nuclear 
weapons.  Much of this information has in turn contributed to the 
general knowledge of atmospheric and oceanic transport processes 
and of bio-geochemical cycles of elements.  Extensive studies of 
radiation effects in animals and numerous epidemiological surveys 
of exposed population groups have by now been conducted.  They have 
considerably enlarged our understanding of the biological effects 
of radiation on man and the environment, although uncertainties 
still remain, particularly regarding the basic mechanisms of action 
and the risk evaluations at low doses and dose rates [U1-U7]. 

10.  A few definitions and general concepts should be introduced 
before the detailed presentation of radionuclide assessments.  The 
basic unit of radioactivity is the becquerel (Bq), corresponding to 
one disintegration per second.  The previously used unit was the 
curie (Ci), one Ci corresponding to 3.7 1010 Bq. 

11.  The basic measure of radiation interaction in irradiated 
materials is the absorbed dose (D).  This quantity is also the 
basis of health risk estimates, under the assumption of a linear 
relationship between dose and risk.  The absorbed dose is defined 
as the mean energy (joules) imparted to the irradiated material per 
unit mass (kg) at the point of interest.  The unit of absorbed dose 
is ca11ed the gray (Gy) which corresponds to 1 J/kg.  The unit of 
absorbed dose previously in use, the rad, is one hundred times 
smaller than the Gy. 

12.  Radiations of different types and energies have different 
effectiveness for producing effects, depending on the amount of 
energy transferred per unit length (LET) along the path of the 
charged particles.  In order to quantify this differing 
effectiveness, use is made of a normalizing quantity called the 
quality factor (Q).  For general purposes of radiological 
protection the assumed values of Q are:  1 for x and gamma rays and 
for electrons;  10 for neutrons and protons;  20 for alpha and 
multiply charged particles. 

13.  The product of the absorbed dose, D, and the quality factor, 
Q, is termed the dose equivalent (H).  The unit of dose equivalent 
is the sievert (Sv).  The previously used unit was the rem (1 rem = 
0.01 Sv).  Use of the dose equivalent allows the summation of doses 
from all types of radiation of different biological effectiveness. 

14.  The exposure of an individual to a source of radiation may be 
expressed in terms of the absorbed dose or dose equivalent during 
the period of exposure.  In the natural radiation environment the 
exposure is continuous and it is sufficient to give the annual 
average dose or dose rate. There are important spatial variations 
to be considered, for example, as a function of the altitude in 
case of exposure to cosmic radiation or as a function of the 
geographical location due to the different radionuclides present in 
soil. 

15.  For specific releases of radioactive materials into the 
environment (atmospheric nuclear tests, operation of nuclear 
reactors) there are also important temporal variations in the 
exposure.  In order to account for the exposures which will occur 
in the future from specific sources, use is made of the dose 
commitment (Dc).  This quantity is the infinite time integral of 
the average individual dose rate.  Dose commitments may not 
represent doses to specific individuals. For example, if the 
radionuclide released has a very long half-life, the dose 
commitment is derived from the doses to successive generations in 
the population. 

16.  The collective harm to a population resulting from the 
exposure of all individuals is related to the collective dose in 
the population, particularly if the linearity of the relationships 
between dose and effects may be assumed for the exposures involved.  
The collective dose (S) in a given population is the summation of 
products of the average individual doses and the number of 
individuals in each range of doses.  The summation may become an 
integral for continuous variations over the entire range of doses.  
The unit of the collective dose is man Gy and the corresponding 
unit of collective dose equivalent is man Sv. 

17.  The measure of the total exposure of a population from a 
specified source or release practice is the collective dose 
commitment (Sc), defined as the infinite time integral of the 
collective dose rate.  The relevant units are man Gy, or man Sv in 
case of the collective dose equivalent commitment. 

18.  In radiation exposure assessments, it is often necessary to 
account for the different sensitivity of individual organs of the 
body with respect to each other or to irradiation of the whole 
body, particularly in the case of internally deposited 
radionuclides.  Weighting factors for the relevant organs may be 
derived for this purpose from relative risk estimates.  These 
factors will be listed in the section on radiation effects with 
some additional discussion. 

19.  The summation of the products of the weighting factors and the 
dose equivalents for individual organs gives a single measure to be 
used as an index of health detriment, called the effective dose 
equivalent (HE).  The concepts of collective and committed doses 
may also be used with this quantity.  Thus a final quantity for 
health assessments may be the collective effective dose equivalent 
commitment, (ScE) which is a collective dose, weighted for the 
effects of doses within the body and dose distributions within the 
population. 

20.  The chain of events leading from the release of radioactive 
materials into the environment to the irradiation of human tissues 
may be expressed schematically as a series of compartments 
connected by transfer pathways.  Such models are necessarily 
simplifications of the actual transfer pathways. The following 
diagram illustrates the transfer stages most usually considered in 
assessments by UNSCEAR. 

FIGURE 1

21.  The basic task in the assessment process is to evaluate the 
transfer factors (Pi,j) which relate the appropriate quantity of 
radioactivity amount or dose in step i of the sequence to the 
appropriate quantity in the subsequent step j. Since the desired 
quantity in the final step is the time integrated dose rate, the 
dose commitment from a specific source, the quantities in the other 
steps are the time integrated activity concentrations.  The 
transfer factor is the quotient of time integrated quantities in 
successive compartments.  The total transfer factor for steps in 
series is the product of the transfer factors involved.  The total 
transfer factor of several parallel pathways is the sum of the 
transfer factors of the individual pathways. 

22.  There are many common features of the behaviour of different 
radionuclides in the environment and their transfer to man.  For 
example, the physical dispersion of radionuclides in the 
environment following release from a source is largely the same for 
broad classes of material, such as particulates and gases.  Several 
models used to describe the transfer of radioactive material within 
an environmental medium or from one medium to the next have general 
applicability.  A review of such general behaviour and modelling 
procedures can be found in the 1982 report of UNSCEAR [U8].  
Therefore, in the presentations which follow only the rather more 
specific aspects of environmental behaviour and dosimetry of the 
radionuclides are considered. 

REFERENCES

U1  United Nations.  Report of the United Nations Scientific
    Committee on the Effects of Atomic Radiation.  Official
    Records of the General Assembly, Thirteenth Session,
    Supplement No. 17 (A/3838).  New York, 1958.

U2  United Nations.  Report of the United Nations Scientific
    Committee on the Effects of Atomic Radiation.  Official
    Records of the General Assembly, Seventeenth Session,
    Supplement No. 16 (A/5216).  New York, 1962.

U3  United Nations.  Report of the United Nations Scientific
    Committee on the Effects of Atomic Radiation.  Official
    Records of the General Assembly, Nineteenth Session,
    Supplement No. 14 (A/5814).  New York, 1964.

U4  United Nations.  Report of the United Nations Scientific
    Committee on the Effects of Atomic Radiation.  Official
    Records of the General Assembly, Twenty-first Session,
    Supplement No. 14 (A/6314).  New York, 1966.

U5  United Nations.  Report of the United Nations Scientific
    Committee on the Effects of Atomic Radiation.  Official
    Records of the General Assembly, Twenty-fourth Session,
    Supplement No. 13 (A/7613).  New York, 1969.

U6  United Nations.  Ionizing Radiation:  Levels and Effects.
    A report of the United Nations Scientific Committee on the
    Effects of Atomic Radiation to the General Assembly, with
    annexes.  United Nations sales publication, No. E.72.IX.17
    and 18.  New York, 1972.

U7  United Nations.  Sources and Effects of Ionizing
    Radiation.  United Nations Scientific Committee on the
    Effects of Atomic Radiation 1977 report to the General
    Assembly, with annexes.  United Nations sales publication
    No. E.77.IX.I.  New York, 1977.

U8  United Nations.  Ionizing Radiation:  Sources and
    Biological Effects.  United Nations Scientific Committee
    on the Effects of Atomic Radiation 1982 report to the
    General Assembly, with annexes.  United Nations sales
    publication No. E.82.IX.8.  New York, 1982.

II.  TRITIUM

A.  INTRODUCTION

23.  Tritium, 3H, is a radioactive isotope of hydrogen which decays 
into the stable nuclide 3He.  Tritium is a pure beta-emitter with a 
half-life of 12.3 a, a maximum energy of 18 keV and an average 
energy of 5.7 keV.  Tritium is produced naturally in the 
atmosphere, where it results from the interaction of cosmic ray 
protons and neutrons with nitrogen, oxygen, and argon.  Man-made 
tritium, in amounts substantially larger than the natural 
inventory, has been injected into the stratosphere by thermonuclear 
explosions.  In addition, tritium is produced during the operation 
of nuclear reactors. 

24.  There are many applications of tritium in industry.  It is 
widely used in consumer products, such as radioluminous timepieces 
and also as a tracer in biomedical research. Environmental tritium 
is mainly found as tritiated water.  As such, it follows the 
hydrological cycle and penetrates into all components of the 
biosphere, including man. 

25.  This document is mainly based on the 1977 UNSCEAR report [U1], 
but makes also extensive use of the contents of recent reviews or 
symposia on tritium [I4, J1, M7, N1, N2]. 

B.  SOURCES

1.  Natural tritium

26.  Natural tritium is produced by nuclear reactions in the 
atmosphere and, to a much smaller extent, in the hydrosphere and in 
the lithosphere.  In addition, some tritium may be created in the 
extra-terrestrial environment and enter the atmosphere along with 
cosmic rays.  Most of the natural tritium is found in the 
environment as tritiated water, generally designated as HTO. 

27.  In the atmosphere, natural tritium is produced by the 
interaction of high energy cosmic rays with atmospheric nitrogen 
and oxygen.  The estimates of the number of atoms of tritium 
produced per unit time and per unit area of the earth's surface 
range from 0.1 to 1.3 cm-2 s-1 [U1].  In the UNSCEAR 1977 report 
[U1], a production rate of 0.25 cm-2 s-1 was adopted;  this 
corresponds to a production rate of 3.6 1016 Bq a-1 in each 
hemisphere and to a global inventory of 1.3 1018 Bq at equilibrium. 

28.  It has been suggested that tritium might be ejected from the 
sun during solar flares [L1] and from stars [F1].  Flamm et al. 
[F2] estimated that the solar flares could account for an 
additional production rate, averaged over the solar cycle, of 0.1 
cm-2 s-1. 

29.  In the lithosphere and in the hydrosphere, tritium is produced 
by interaction of neutrons with 6Li nuclides.  The production rates 
have been assessed at 10-3 cm-2 s-1 in the lithosphere and at 10-6 
cm-2 s-1 in the hydrosphere [F1, K1].

2.  Nuclear explosions

30.  Nuclear tests have been conducted in the atmosphere since 1945 
and have produced tritium in amounts that greatly exceed the global 
natural activity.  The tritium activity arising from atmospheric 
nuclear tests can be estimated from the fission and fusion yields 
or from environmental measurements. 

31.  Bennett [B1] has published an estimate of the total and 
fission yields for each reported atmospheric test from 1945 to 
1978;  according to that compilation, 422 nuclear tests were 
conducted in the atmosphere up to 1979, with cumulative yields of 
217 Mt and 328 Mt for fission and fusion, respectively. The tritium 
activity produced per unit yield depends on the characteristics of 
the device, as well as on those of the explosion site, but is in 
any case much greater for fusion than for fission [N1].  Miskel 
[M1] estimated the yield for fission explosion to be 2.6 1013 Bq 
Mt-1 and that for fusion to be typically 7.4 1017 Bq Mt-1.  The 
total tritium activity produced by atmospheric tests is thus  
assessed at 

    328 Mt (fusion) x 7.4 1017 Bq Mt-1 = 2.4 1020 Bq

Most of this activity was produced during the large yield test 
series which took place during 1954-1958 and 1961-1962;  the 
contribution of the nuclear tests carried out since 1964 is less 
than 5% of the total. 

32.  Almost all the tritium produced by fallout occurs as HTO in 
the atmosphere and the hydrosphere, and thus follows the 
hydrological cycle.  The total activity injected can therefore be 
conceivably derived from measured concentrations in water samples.  
From the study of Schell et al. [S1] on the tritium concentrations 
in precipitation at stations in the IAEA network, it can be 
estimated [U1] that the total production was about 1.7 1020 Bq.  
Other estimates, using vertical profiles of 3H in the oceans as a 
basis, lead to injections of 1.2 1020 Bq (in the oceans only) [O1], 
1.3 1020 Bq [B2, U1], and 2.0 1020 Bq [M2]. 

33.  All the estimates presented above are in fairly good 
agreement, as they lie in the limited range from 1.2 1020 to 2.4 
1020 Bq.  In its 1977 report UNSCEAR adopted a value of 1.7 1020 Bq 
for the total globally dispersed activity of tritium produced in 
atmospheric tests up to 1976 [U1]. 

3.  Nuclear fuel cycle

34.  Tritium occurs in nuclear reactors by ternary fission in the 
fuel and also by neutron activation reactions with lithium and 
boron isotopes dissolved in, or in contact with, the primary 
coolant as well as with naturally-occurring deuterium in the 
primary coolant (Figure II.I). 

FIGURE II.I

35.  Most of the fission product tritium produced in the fuel rods 
is usually retained within the fuel and is not released into the 
environment at the reactor site; it is instead released during fuel 
reprocessing, if that practice is carried out.  The activity 
produced in the coolant is partly or entirely released in the 
effluent streams according to the waste management practices at the 
plant. 

36.  Releases into the environment are mainly in the form of HTO in 
reactors that use water as primary coolant, as well as in fuel 
reprocessing plants. 

(a)   Nuclear reactors

37.  Four types of reactors have been considered (PWR, BWR, HWR, 
GCR), the emphasis being on PWRs and BWRs which currently represent 
the largest share of nuclear capacity.  Estimated generation rates 
and appearance of tritium in effluent streams of reactors are 
summarized in Table II.1. 

38.  The annual production of fission product tritium in the fuel 
rods of a pressurized water reactor (PWR) is in the range of 6 to 9 
1011 Bq per MW(e)a [N1].  A small percentage, 1% or less, is 
expected to be released into the coolant through defects in the 
cladding, currently made of zirconium alloy. In contrast, the use 
of stainless steel cladding in earlier PWRs resulted in the release 
to the coolant of most of the tritium produced in the fuel. 

39.  Tritium generation in the primary coolant (water) of a
PWR is mainly due to reactions with boron (2.6 1010 Bq per
MW(e)a) which is dissolved as boric acid to control
reactivity;  in addition, the maintenance of 2 ppm lithium
hydroxide for pH control [L2] results in the formation of
about 7 108 Bq per MW(e)a.

40.  Environmental tritium discharges from PWRs depend on waste 
management practices as well as on the materials used in the 
reactor.  Average normalized releases of tritium were shown in the 
UNSCEAR 1977 report [U1] to be about 7 1010 Bq per MW(e)a in liquid 
effluents and 7 109 Bq per MW(e)a in airborne effluents for the 
reactors in operation in 1973-1974.  However, large differences 
between PWRs are due to the type of fuel cladding.  For an old 
reactor using stainless steel Kahn et al. [K3] measured 3H releases 
of about 4 1011 Bq per MW(e)a in liquid effluents and 4 1010 Bq per 
MW(e)a in airborne effluents, whereas the combined releases of 9 
PWRs with zirconium alloy clad fuel (current practice) were 
reported by NCRP [N1] to be about 3 1010 and 109 Bq per MW(e)a in 
liquid and airborne effluents, respectively. 

41.  In boiling water reactors (BWRs) tritium is produced by 
ternary fission in the fuel at about the same rate as in PWRs (6 to 
9 1011 Bq per MW(e)a).  The generalized use of zirconium alloy 
cladding limits the tritium release into the coolant to less than 7 
109 Bq per MW(e)a. 

42.  Tritium can be generated by neutron activation in the coolant 
and in the control rods.  Prior to 1971, control rods of boron 
carbide were used in BWRs [S2]; the production of tritium by 
activation of these control rods has been estimated to be about 3 
1011 Bq per MW(e)a.  However, tritium has not been shown to diffuse 
through the boron carbide matrix [T1]. In the coolant itself, 
tritium is generated by activation of deuterium at a rate of about 
4 108 Bq per MW(e)a. 

43.  Tritium activities discharged from BWRs into the environment 
are lower than those of PWRs because less tritium is produced in or 
diffuses into the primary coolant.  UNSCEAR [U1] reported the 
average discharge rates to be 4 109 and 2 109 Bq MW(e)a in liquid 
and airborne effluents, respectively. 

44.  The amount of tritium generated in fuel of heavy water 
reactors (HWR) by ternary fission is approximately the same as in 
light water reactors, but it is largely exceeded by the production 
in the D2O coolant and moderator by neutron activation, which has 
been estimated to be about 2 1013 Bq per MW(e)a [K2]. 


Table II.1  Estimated rates of generation of tritium and of its release in effluent streams of 
different types of reactors (1010 Bq per MW(e)a) [G1, K2, S2, T1, U1]
---------------------------------------------------------------------------------------------------------
                         PWR                    BWR                    HWR                    GCR
Source         ------------------------------------------------------------------------------------------
               Generation  Effluent   Generation  Effluent   Generation  Effluent   Generation  Effluent
                           stream                 stream                 stream                 stream
---------------------------------------------------------------------------------------------------------
 Fission        75          < 0.7      75          < 0.7      55          < 0.6      75          < 0.7

 Activation
 Deuterium     0.004       0.004      0.04        0.04       2000        75a/
 Lithium       0.07        0.07                                                     2           0.4
 Boron         2.6         2.6        30          0


Rounded total  80          3          110         0.5        2000        75         80          1
---------------------------------------------------------------------------------------------------------
a/  Depending on the irradiation time and on the net leakage of heavy water.
45.  Environmental discharges depend upon the D2O leakage which is 
kept as small as possible for economical and radiological reasons, 
and upon the tritium activity in the coolant and moderator, which 
builds up with the irradiation time.  Annual losses of from 0.5% to 
3% are anticipated in HWRs [U1].  For the optimal loss of 0.5% per 
year, the normalized tritium release rate ranges from 1011 Bq per 
MW(e)a in the first year of operation to about 7 1011 Bq per MW(e)a 
in the tenth year.  Based on the latter value as representative of 
the reactor life, the normalized 3H release rates are estimated to 
be 6 1011 and 1.5 1011 Bq per MW(e)a in airborne and liquid 
effluents, respectively [G1].  Reported releases roughly agree with 
these estimates:  they are 6.3 1011 and 2.6 1011 Bq per MW(e)a for 
the Pickering A station in Canada, in airborne and liquid 
effluents, respectively whereas the Atucha reactor in Argentina 
releases about 8 1011 Bq per MW(e)a both in airborne and in liquid 
effluents. 

46.  In gas-cooled reactors (GCR), tritium is produced by ternary 
fission (about 7 1011 Bq per MW(e)a) and by activation of lithium 
in the graphite moderator.  Based on the experience with UK 
reactors (mainly Magnox reactors), the tritium release is about 7 
109 Bq per MW(e)a in liquid effluents and ranges from 109 to 1010 
Bq per MW(e)a in airborne effluents [U1]. 

(b)   Fuel reprocessing plants

47.  At the fuel reprocessing stage of the nuclear fuel cycle (if 
it is undertaken) the elements uranium and plutonium in the 
irradiated nuclear fuel are recovered for reuse in fission 
reactors.  When the fuel elements are reprocessed, the uranium is 
first taken out of its cladding material and then dissolved in 
nitric acid.  Most of the tritium released from fuel during 
dissolution appears in the liquid waste stream while some is 
carried out in the dissolver off-gas stream and a portion is 
immobilized as a solid zirconium compound in the cladding. 

48.  In 1980, the only reprocessing plants operating commercially 
in the world were at Windscale (U.K.) and La Hague and Marcoule 
(France);  their combined capacity was much lower than the amount 
of fuel discharged from reactors worldwide.  Luykx and Fraser [L3] 
have expressed the reported releases from the three reprocessing 
plants during the 1974-1978 time period in terms of activity 
discharged per unit of electricity generated.  The average figures 
for each plant are given in Table II.2. 

Table II.2  Average normalized tritium activities 
discharged into the environment by fuel 
reprocessing plants (1010 Bq per MW(e)a) [L3]
--------------------------------------------------
Plant            Airborne     Liquid        Total
location         effluents    effluents
--------------------------------------------------
Windscale        17           55            72
La Hague         0.4          28.5          29
Marcoule         5.2          41            46
--------------------------------------------------

49.  As there is no retention system for tritium in the currently 
operating reprocessing plants, the activity released corresponds to 
that which is contained in the fuel elements (with the exclusion of 
cladding) at the time of reprocessing. The production rate of 
tritium in reactors being about 75 1010 Bq per MW(e)a (Table II.1), 
approximately half of the theoretical fuel content seems to be 
unaccounted for at the La Hague and Marcoule plants. 

(c)   Summary

50.  In 1980, the installed nuclear capacity was 1.25 105 MW(e) on 
a worldwide scale [I1].  Assuming an average load factor of 0.6, 
the energy produced was 7.5 104 MW(e)a.  Using the average figures 
given previously for production and release in the types of 
reactors considered, the global production and release of tritium 
at the reactor sites in 1980 are estimated to be about 1.5 1017 Bq 
and 4 1015 Bq, respectively.  Table II.3 provides a breakdown of 
the environmental discharges from reactors according to reactor 
type. 

Table II.3  Estimated global discharge of tritium from nuclear 
power stations in 1980
----------------------------------------------------------------
                             Estimated tritium discharges 
Reactor   Number  Capacity   in 1980 (Bq)                       
type              [MW(e)]    Airborne     Liquid       Total
                             effluents    effluents
----------------------------------------------------------------
PWR       96      64239      3.9 1013     1.2 1015     1.2 1015
BWR       62      35170      4.2 1013     8.4 1013     1.3 1014
HWR       14      5963       5.4 1014     2.1 1015     2.6 1015
GCR       36      7086       1.3 1013     3.0 1013     4.3 1013
Other     33      12527      -            -            -
----------------------------------------------------------------
Total     241     124985     6.3 1014     3.4 1015     4.0 1015
----------------------------------------------------------------

51.  In comparison, the tritium releases reported for the three 
currently operating commercial fuel reprocessing plants were about 
2 1015 Bq in 1978.  All together, the current tritium production 
rate in the nuclear fuel cycle is comparable to the natural 
production rate, whereas the release rate is about 20 times less. 

4.  Tritium production plants

52.  Artificial production of tritium on an industrial scale is 
necessary to provide an essential component of thermonuclear 
weapons.  In addition, relatively small amounts of tritium are used 
for other industrial and scientific applications.  The most 
economical way to produce tritium is the irradiation of lithium 
metal, alloys or salts in a nuclear reactor [J1].  The tritium is 
isotopically separated from other hydrogen isotopes and is 
processed in tritium-handling plants [C1]. 

53.  Tritium airborne release rates from Savannah River Plant, 
which is the primary production source of tritium in the U.S.A., 
have ranged from 1.4 1016 Bq a-1 to 9.9 1016 Bq a-1 from 1974 to
1977 with an average of 4.1 1016 Bq a-1 [M4]. Under normal 
operating conditions, the releases are about 20% HT and 80% HTO.  
However, accidental airborne releases, which seem to be essentially 
in the gaseous HT form, have raised the contribution of HT to the 
total activity released to 60% in 1974 and 57% in 1975 [M4].  The 
activity of tritium released in the liquid effluents appears to be 
about 10% of that in the airborne effluents [N1]. 

54.  Data on releases from other tritium production plants have not 
been found in the literature.  However, an indirect estimate of 7 
1016 Bq for the worldwide release of HT in 1977 has been made by 
Mason and Ostlund [M3] on the basis of their measurements of the 
atmospheric HT content. 

5.  Consumer products

55.  Tritium has been used extensively in the dial-painting 
industry for the illumination of timepieces, the radiation emitted 
by 3H being converted into light by a scintillator which is usually 
zinc sulfide containing small amounts of copper or silver.  In 
recent years, this illumination system has been in competition with 
the tritium gas-filled glass tubes, coated internally with 
phosphor, which are used to illuminate some types of LCD (liquid 
crystal display) watches.  Exit signs and electronic tubes are 
other types of consumer products containing tritium [C2, K4, U1, 
W1]. 

56.  In luminous compounds, the fractional release rate of tritium, 
in the form of HTO, HT and short-chain organic radicals of the 
styrene type, is about 5% annually [K4, K5] while it is negligible 
from gas-filled glass tubes.  It has been estimated that about 7 
1016 Bq  was processed in 1978 in the worldwide production of 
timepieces and that the activity released is probably under 1014 Bq 
a-1 for luminous compounds and 2 1012 Bq a-1 for gas-filled glass 
tubes [K5]. Environmental releases due to breakage through accident 
or disposal could be more important [C2, W1]. 

6.  Controlled thermonuclear reactors

57.  Large-scale use of controlled thermonuclear reactors for heat 
or power generation seems quite unlikely in the next 25 years.  
However, if thermonuclear reactors come into use, they will contain 
substantial inventories of tritium and will pose considerable 
tritium management problems [N1].  The production of tritium in a 
nominal 1000 MW(e) controlled thermonuclear reactor is anticipated 
to be about 5 1017 Bq d-1 and the inventory of the order of 1019 Bq 
[C1, C3, H1].  In order to prevent massive releases of tritium into 
the environment, an extraordinary degree of control will be 
required.  However, conceptual designs for fusion power plants show 
that the effluent release rate can be limited to 4 1013 Bq a-1 by 
applying present-day tritium technology [C3]. 

C.  BEHAVIOUR IN THE ENVIRONMENT

1.  Natural and fallout tritium

58.  Natural and fallout tritium are mainly produced in the 
stratosphere where they are essentially found in the HTO form.  
Tritiated water vapour is transferred from the stratosphere to the 
troposphere with a half-time of about one year, then from 
troposphere to the earth's surface through rainfall and molecular 
exchange with a half-time of about ten days.  Tritiated water then 
follows the hydrological cycle. Water deposited on the ocean 
surfaces is diluted in the mixed layer.  Part of it evaporates back 
to the atmosphere, with a much lower concentration, while a smaller 
fraction is transferred to the deep ocean.  Tritiated water 
deposited on land surfaces is partitioned partly to surface run-off 
(leading to a pond, a lake, a stream, or an ocean) and partly to 
infiltration in the soil from where it can be absorbed by plants, 
evaporate, or move with groundwater to a surface stream or to an 
ocean. 

59.  Part of the tritiated water deposited on soils finds its way 
into vegetable and animal products and thus contaminates dietary 
foodstuffs.  Tritium incorporated into those biological materials, 
and in soil and sediments as well, is found to be present in at 
least two separable fractions, one easily exchangeable, that is 
available by freeze-drying (free water tritium fraction) and one 
less easily exchangeable, available by combustion ("organically 
bound" fraction) [B6]. The analysis of soil, water, and various 
components of the diet in the New York area in 1978 [B6] revealed 
that water, soil and diet were in equilibrium with respect to free 
water tritium;  however, the specific activities (activity 
concentration per unit mass of hydrogen) of the "organically bound" 
tritium in various foodstuffs were higher by a factor of 2 to 4 
than those of the water tritium.  It is suggested that tritium was 
incorporated uniformly into biological materials during the period 
of highest deposition rates in the early 1960s and that differences 
in specific activities developed due to longer biological residence 
half-time of the "organically bound" fraction compared to the free 
water tritium fraction [B6]. 

2.  Industrial releases

60.  Industrial releases consist mainly of HTO and HT, and probably 
tritiated methane, CH3T [B7].  The residence times of HT and CH3T 
in the atmosphere are not known with certainty but the estimates 
point to average values of 5 to 10 years [B7]. The main removal 
processes are bacterial action and photochemical oxidation for HT 
and photochemical oxidation alone for CH3T [B7].  In both cases, 
the resulting product is presumably HTO.  As HTO is much more 
biologically active than HT and CH3T, it is this tritium compound 
that is of most concern in the case of industrial releases. 

61.  Industrial releases may be to the atmosphere or to water 
(river or sea).  In addition, releases to ground water have taken 
place but they are of little consequence as the movement of water 

in suitable aquifers is very slow.  The environmental behaviour of 
HTO released by industry is not different from that from natural or 
fallout sources. 

D.  TRANSFER TO MAN

62.  Transfer to man of environmental HTO takes place via 
inhalation, diffusion through skin and ingestion of beverages and 
foodstuffs;  in the case of HT, inhalation is the only meaningful 
pathway to man.  Exposure to an atmosphere contaminated with 
tritiated water vapour results in total absorption of the inhaled 
activity through the lungs and absorption of about 50% of that 
amount through the intact skin [I2].  Ingested tritiated water is 
completely absorbed from the gastro-intestinal tract. 

63.  Absorbed tritiated water is rapidly distributed throughout the 
body via the blood.  Tritiated water in blood equilibrates with 
extracellular fluid in about 12 minutes. However, in poorly 
vascularized tissues, such as bone and fat, equilibrium with plasma 
water may take days to weeks [N2, W2].  The biological half-life of 
tritium in the body following intake of tritiated water has been 
found to range from 2.4 to 18 days among 300 individuals [B3, W3].  
The experience from observations of human cases of accidental 
tritium exposures with intakes large enough to allow relatively 
long-term monitoring shows that the excretion rate can be 
represented as the sum of three exponentials with half-times of 
residence of the order of 10 days, one month, and one year [L4, M5, 
M6, S3].  The first component is believed to reflect the turnover 
of body water while the second and the third components are likely 
to represent the turnover of tritium incorporated into organic 
compounds. 

E.  DOSIMETRY

1.  Dose per unit intake

(a)   Tritiated water

64.  External irradiation from tritium does not need to be 
considered as the range of the electrons emitted by decay (at most 
6 µm in soft tissue) is shorter than the depth of the basal cells 
in the epidermis.  Following a chronic intake of 1 Bq 1-1 of 
tritium (as HTO) in air, water and food the equilibrium dose rate 
in active wet tissue (the totality of soft tissues with the 
exclusion of fat) is 2.6 10-8 Gy a-1. Of that dose, 16% is 
calculated to be due to tritium contained in organic pools of the 
body.  These results were derived by Bennett [B4] based on human 
retention data. 

65.  When all the sources of intake (air, water and food) are 
assumed to be contaminated at the same level, use can be made of 
the specific activity model which consists in assuming that the 
specific activity of tritium (activity concentration per unit mass 
of hydrogen) in the body is the same as that in the intake.  A 
chronic intake of tritium at a concentration of 1 Bq per litre of 

water would thus give rise to an absorbed dose averaged over the 
whole body of 

             10-3             
     1 Bq        1H2O         gH2O       gH
     ----- x ----------- x 18 ---- x 0.1 -----
     1H2O       gH2O           gH        gbody

                 MeV               s               Gy gbody
     x 5.7 10-3 ------ x 3.16 107 --- x 1.6 10-10 ----------
                 Bq s              a                  MeV

     =  2.6 10-8 Gy a-1 per Bq 1-1

This result is numerically equal to that of Bennett [B4].  The 
doses in individual tissues depend on their hydrogen 
concentrations.  According to the values adopted for the Reference 
Man of ICRP [I2], the hydrogen concentration per unit mass is the 
same (10%) in total body and in total soft tissues and is, as a 
first approximation, uniform in the soft tissues.  Hydrogen content 
is lowest in mineral bone (about 4%) and highest in adipose tissue 
(12%).  Since the range of the beta-particles emitted by tritium 
decay is very small, it can be assumed that all the energy emitted 
in a given tissue is absorbed in the same tissue.  The effective 
dose equivalent is therefore numerically equal to the absorbed dose 
averaged over the whole body and is 2.6 10-8 Sv a-1 per Bq 1-1. 
Assuming a rate of intake of 3 litres of water (in beverages and in 
food) per day and a water vapour atmosphere concentration of 8 g m-3, 
the effective dose equivalent per unit intake is found to be 2.2 
10-11 Sv Bq-1 while the effective dose equivalent rate per unit 
atmospheric concentration would be 2.1 10-9 Sv a-1 per Bq m-3. 

(b)   Tritiated hydrogen

66.  The doses from inhalation of HT are much lower than those from 
HTO for a given atmospheric concentration of tritium. The dose rate 
to the lungs per unit concentration of HT in air is about 10-14 Gy 
h-1 per Bq m-3 [I3], while the doses in tissues from the absorbed 
gas are 60 to 150 times smaller [I3].  The corresponding effective 
dose equivalent rate per unit concentration in air is therefore 1.1 
10-11 Sv a-1 per Bq m-3. 

2.  Dose per unit release

(a)   Natural tritium

67.  Doses from natural tritium can be estimated from the few 
tritium measurements in environmental materials that were carried 
out before the contamination with fallout (or that had been 
preserved from contamination).  Activity concentrations of 
continental surface waters were then found to be in the range from 
0.2 to 0.9 Bq 1-1 [K1].  The production rate of natural tritium 
being constant in time and relatively uniform on the global scale, 
the concentrations in all the components of human intake (air, 
water and food) of natural tritium are in steady-state equilibrium 
with the concentrations in continental surface waters.  Using the 

specific activity approach, it is assumed that the specific 
activity of natural tritium is the same in the continental surface 
waters, in all the components of human intake and in the body.  The 
effective dose equivalent rate is thus found to range from 
0.2 Bq 1-1 x 2.6 10-8 Sv a-1  per Bq 1-1  = 5.2 109 Sv a-1 to 
0.9 Bq 1-1 x 2.6 10-8 Sv a-1  per Bq 1-1  = 2.3 10-8  Sv a-1, being 
therefore of the order of 10-8 Sv a-1. The effective dose equivalent 
commitment per unit release would then be 

         10-8 Sv a-1
       ---------------  ca. 1.4 10-25 Sv per Bq
       7.2 1016 Bq a-1

Taking the world's population to be 4 109 people, the global 
collective effective dose equivalent commitment per unit of 
activity produced is about 5 10-16 man Sv per Bq. 

(b)   Nuclear explosions

68.  The doses from fallout tritium can be estimated in the same 
way as those from natural tritium.  On the basis of the variation 
with time of the tritium activity concentration in surface waters 
[B5] and of the latitudinal distribution of the fallout deposition 
[S1], UNSCEAR [U1] estimated the effective dose equivalent 
commitments to the populations of the northern and southern 
hemispheres to be 2 10-5 and 2 10-6 Sv respectively. 

69.  The effective dose equivalent commitment from fallout tritium 
was also estimated indirectly, using the relationship obtained for 
natural tritium between the production rate and the dose rate 

                                W
                   Hc = gammao  -
                                B

where Hc is the effective dose equivalent commitment (Sv) from 
production of fallout tritium in a given hemisphere; gammao is the 
effective dose equivalent rate from natural tritium (gammao = 10-8
Sv a-1);  W is the activity of tritium released by nuclear 
explosions (1.5 1020 Bq in the northern hemisphere and 0.2 1020 Bq 
in the southern hemisphere); and B is the natural rate of 
production (3.6 1016 Bq a-1 in each hemisphere).  The effective 
dose equivalent commitments thus derived are 4.2 10-5 Sv for the 
population of the northern hemisphere and 5.6 10-6 Sv for the 
population of the southern hemisphere.  These results are higher 
than the direct estimates by a factor of 2 to 3.  The global 
collective effective dose equivalent commitments per unit activity 
released are estimated to be 9 10-16 and 4 10-16 man Sv Bq-1 using
the latter and the former method, respectively.  UNSCEAR [U1] used 
an intermediate value of 8 10-16 man Sv per Bq. 

(c)   Nuclear installations

70.  While the production of natural and fallout tritium brings 
about a relatively uniform contamination of the whole biosphere, 
the releases from nuclear installations occur at discrete points on 

the earth's surface giving a highly heterogenous spatial 
distribution of concentrations. 

71.  UNSCEAR'S practice is to divide the collective doses into two 
components:  the local and regional collective doses, which are due 
to the first passage of the plume, over distances of 100 to 1000 km 
from the point of release, and the global collective doses, which 
arise from the mixing of tritium in the whole biosphere.  As the 
doses per unit concentration of tritium in air are much higher for 
HTO than for HT, tritiated water will be the only compound 
considered in the estimate of the local and regional collective 
doses. 

(i)   Local and regional collective dose

72.  A distinction is made between airborne and liquid effluents.  
Tritium present in airborne effluents can contribute to the local 
and regional collective doses through inhalation, absorption 
through skin and ingestion.  As the contribution from the ingestion 
pathway is quite variable from site to site owing to differences in 
local hydrology and water usage, UNSCEAR [U1] has not taken this 
pathway into account in its assessment of the local and regional 
collective doses. Assuming an atmospheric dispersion factor of 5 
10-7 s m-3 at 1 km from the release and a reduction in 
concentration inversely proportional to the 1.5 power of the 
distance expressed in kilometres, the local collective dose per 
unit activity released can be assessed by integration over the 
local area. Integrating from 1 to 100 km for a population density 
of 100 km-2, UNSCEAR [U1] estimated the local collective dose from 
airborne tritium per unit activity released to be about 5 10-17 man 
Sv per Bq. 

73.  The collective dose commitment from the input of 3H to water 
bodies, normalized per unit activity released, can be estimated 
[U1], using the expression 
            
     c      sigmak Nk Ik fk phi
    S  =    -------------------
     1      V(lambda + 1/tau)

where V is the volume of the receiving waters, tau is the turnover 
time of receiving waters, lambda is the decay constant of 3H, Nk is 
the number of individuals exposed by pathway k, Ik is the 
individual consumption rate of pathway item k, fk is the 
concentration factor for the consumed item in pathway k, and phi is 
the collective dose per unit activity ingested collectively by the 
exposed group. 

                          1        
74.  The quantity V(lambda + 1/tau) is the infinite time integral 
of the water concentration per unit of activity released, while the 
quantity multiplied by fk is the infinite time integral of the 
concentration in the consumed item k.  For radionuclide inputs into 
small volumes of water, the concentrations in water and in fish 
will be high, but the population which can be served with drinking 
water or by fish consumption will limited.  For inputs into larger 

volumes of water, the concentrations will be smaller, but the 
populations involved will be correspondingly larger.  It is 
reasonable, therefore, to assume as a first approximation that the 
quantities Nk/V are relatively constant, independent of V.  The 
values for these quantities as well as values for the other 
parameters of the above expression have been extensively discussed 
[U1]. 

75.  A summary of the values used in the assessment, based on 
UNSCEAR [U1], and the evaluation of the collective dose commitments 
for a release of 1 Bq of 3H in liquid effluents are given in Table 
II.4. 

(ii)  Global collective dose

76.  For HTO releases, the global collective effective dose 
equivalent commitment established for fallout tritium (8 10-16 man 
Sv per Bq) can be applied without change.  With respect to HT 
releases, if it is assumed that the conversion to HTO takes place 
on average 5 years after the discharges, the global collective 
effective dose equivalent commitment is estimated to be 

             -0.693 
    8 10-16 e        x 5/12.3 =  6 10-16 man Sv per Bq.

Table II.4  Collective dose factors for 3H in liquid effluents
---------------------------------------------------------------------------
                                        Fresh water      Salt water
---------------------------------------------------------------------------
Activity released, A                    1 Bq             1 Bq
Turnover time of receiving water,       10 a             1.0 a
Sediment removal correction factor, s   1.0              1.0
Time integral of activity in water,
          As
W  = ------------                       6.36 Bq a        0.946 Bq a
     1/tau+lambda

Water utilization, V/N                  3 107 1/man      3 109 1/man
---------------------------------------------------------------------------
FRESHWATER PATHWAYS

1    Drinking water
    Treatment removal factor, f1        1.0
    Consumption, I1                     438 1 a-1
    Collective dose commitment

     c         NI
    S1 = W f1 (--)1D                                     2 10-15 man Sv
               V

Table II.4 (contd.)
---------------------------------------------------------------------------
                                        Fresh water      Salt water
---------------------------------------------------------------------------

2.   Fish
    Concentration factor, f2            1.0
    Consumption, I2                     1 kg a-1
    Collective dose commitment

     c         NI
    S2 = W f2 (--)2D                                     5 10-18 man Sv
               V  

SALT WATER PATHWAYS

3.   Fish
    Concentration factor, f3            1.0
    Consumption, I3                     6 kg a-1
    Collective dose commitment

     c         NI
    S3 = W f3 (--)3D                                     4 10-20 man Sv
               V

4.   Shellfish
    Concentration factor, f4            1.0
    Consumption, I4                     1 kg a-1
    Collective dose commitment

     c         NI
    S4 = W f4 (--)4D                                     7 10-21 man Sv
               V
---------------------------------------------------------------------------

(iii)   Summary of collective dose commitments per unit activity released

77.  Table II.5 summarizes the values obtained above for the 
collective effective dose equivalent commitments per unit of 3H 
activity released.  With respect to the local and regional 
component due to industrial releases, the largest collective 
effective dose equivalent commitment per unit activity released is 
obtained for a river discharge and the smallest for a sea discharge 
while an intermediate value is found for the airborne discharge. 

Table II.5  Summary of collective effective dose 
equivalent commitments per unit tritium activity released 
(man Sv per Bq)
---------------------------------------------------------
Origin                 Local and regional   Global 
                       component            component
---------------------------------------------------------
Natural                                     5 10-16
Nuclear tests                               8 10-16

 Industry

Airborne discharge     5 10-17 (HTO)     )  8 10-16 (HTO)
River discharge        2 10-15           )
Sea charge             5 10-20           )  6 10-16 (HT)
---------------------------------------------------------

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B4  Bennett, B.G.  Environmental tritium and the dose to man. 
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F2  Flamm, E., R.E. Lingenfelter, J. F. MacDonald et al. Tritium 
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I3  International Commission on Radiological Protection. Limits for 
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I4  International Atomic Energy Agency.  Behaviour of Tritium in 
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J1  Jacobs, D.G.  Sources of tritium and its behaviour upon release 
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K1  Kaufamn, S. and W.F. Libby.  The natural distribution of 
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K2  Kouts, H. and J. Long.  Tritium production in nuclear reactors. 
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K3  Kahn, B., R.L. Blanchard, W.L. Brinck et al. Radiological 
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K4  Krejci, K. and A. Zeller, Jr.  Tritium pollution in the Swiss 
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K5  Krejci, K.  Discussion. p. 101  in Behaviour of Tritium in the 
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L1  Lal, D. and B. Peters.  Cosmic ray produced radioactivity
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L2  Locante, J. and D.D. Malinowski.  Tritium in pressurized water 
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L3  Luykx, F. and G. Fraser.  Radioactive effluents from nuclear 
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L4  Lambert, B.E., H.B.A. Sharpe and K.B. Dawson.  Am. Ind. Hyg. 
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M1  Miskell, J.A.  Production of tritium by nuclear weapons. p. 79-
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M2  Michel, L.  Tritium inventories of the world oceans and their 
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M3  Mason, A.S. and H.G. Ostlund.  Atmospheric HT and HTO:  V. 
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    of Tritium in the Environment. IAEA, Vienna, 1979. 

M4  Murphy, C.E. Jr. and M.M. Pendergast.  Environmental transport 
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    p. 361-372  in Behaviour of Tritium in the Environment.  IAEA, 
    Vienna, 1979. 

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M6  Moghissi, A.A., M.W. Carter and E.W. Bretthauer.  Further 
    studies on the long-term evaluation of the biological half-life 
    of tritium.  Health Phys. 23:  805-806 (1972). 

M7  Moghissi, A.A. and M.W. Carter, eds.  Tritium.  Messenger 
    Graphics, Las Vegas, Nevada, 1973. 

N1  National Council on Radiation Protection and Measurements.  
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N2  National Council on Radiation Protection and Measurements.  
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    incorporated in genetic material. NCRP No. 63 (1979). 

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    transport of tritium. p. 303-314  in Behaviour of Tritium in 
    the Environment.  IAEA, Vienna, 1979. 

S1  Schell, W.R., S. Sauzay and B.R. Payne.  World distribution of 
    environmental tritium. p. 374-385  in Physical Behaviour of 
    Radioactive Contaminants in the Atmosphere.  IAEA, Vienna, 
    1974. 

S2  Smith, J.M. and R.S. Gilbert.  Tritium experience in boiling 
    water reactors. p. 57-68  in Tritium (A.A. Moghissi and M.W. 
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S3  Sanders, S.M. Hr. and W. C. Reinig.  Assessment of tritium in 
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    Medica Foundation, Amsterdam, 1968. 

T1  Trevorrow, L.E., B.J. Kullen, R.L. Jarry et al.  Tritium and 
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    United Nations sales publication no. E.77.IX.I.  New York, 
    1977. 

W1  Wehner, G.  Discharges of tritium to the environment from 
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    Environment.  IAEA, Vienna, 1979. 

W2  Woodard, H.Q.  The biological effects of tritium.  United 
    States Atomic Energy Commission.  HASL-229 (1970). 

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    life of tritium.  Health Phys. 9: 911-914 (1963). 

III.  CARBON-14

A.  INTRODUCTION

78.  Carbon-14 has always been present on the earth.  It is 
produced by cosmic ray interactions in the atmosphere.  This 
nuclide is a pure beta-emitter, with a half-life of 5730 years, a 
maximum energy of 185 keV and an average energy of 49.47 keV [N1]. 

79.  Carbon is one of the elements that are essential to all forms 
of life and is involved in most biological and geochemical 
processes on the earth.  Associated with the stable isotopes of 
carbon (12C and about 1.1%13C), there is a very small amount of 14C 
formed in the atmosphere and which has subsequently entered in the 
carbon cycle.  The specific activity of biological carbon, as 
measured in wood samples grown in the nineteenth century, was 0.227 
± 0.001 Bq per gram of carbon [T1], corresponding to an atmospheric 
inventory of 1.4 1017 Bq.  During the present century the specific 
activity of 14C has decreased due to the diluting effect of 
releases into the atmosphere of carbon dioxide from the combustion 
of fossil fuels.  This effect (the Suess effect) accounts for a 
reduction of a few percent. 

80.  In addition to its natural production, carbon-14 is also 
produced by the  detonation of nuclear explosives and by the 
operation of nuclear reactors.  The assessment of the collective 
dose commitments from the releases of man-made carbon-14 is 
facilitated by knowledge of the production rate of natural 
carbon-14. 

B.  SOURCES

1.  Natural carbon-14

81.  Carbon-14 is produced by the action of cosmic ray neutrons on 
nitrogen atoms, both in the stratosphere and in the upper 
troposphere.  UNSCEAR [U3] has estimated the natural production 
rate to be about 1015 Bq per year, a value which has been derived 
from assessments of the natural 14C inventory.  The production rate 
has also been estimated directly from assessments of cosmic ray 
neutrons and the values obtained by different authors range from 1 
to 1.4 1015 Bq per year [U3].  Considering the uncertainties 
involved in determining both the direct production rate and also 
the total 14C inventory of the earth, the estimates are in 
reasonable agreement. 

2.  Nuclear explosions

82.  Carbon-14 is formed in nuclear explosions through the capture 
of excess neutrons by atmospheric nitrogen.  After large 
atmospheric nuclear explosions, most of the 14C is transported into 
the stratosphere, from where it equilibrates with the troposphere 
with a half-time of 1 to 2 years [U3]. 

83.  The inventory of 14C from nuclear explosions has been 
estimated from measurements of excess specific activity in the 
troposphere and in the surface ocean waters, and models 
representing the exchange of 14C between the atmosphere, the 
biosphere and the ocean.  UNSCEAR [U3] has estimated that nuclear 
explosions up to 1972 have injected into the atmosphere 2.1 1017 Bq, 
while subsequent injections have increased this amount by about 1%. 

84.  For the past pattern of atmospheric nuclear explosions, the 
production mentioned above corresponds to about 3.7 1014 Bq per 
megaton.  This value, however, is not representative of any given 
nuclear explosion, because the production of 14C will depend on the 
type of nuclear device exploded and also on whether the explosion  
took place on the surface of the earth or high in the atmosphere.  
A "surface" test will produce approximately 50% of the 14C that 
would be produced by the same device in an "air" test, because 
about one half of the escaping neutrons will be captured in the 
soil or water rather than in the atmosphere. 

3.  Nuclear fuel cycle

85.  Carbon-14 is produced in nuclear reactors and is released to 
the environment at the reactor itself or at reprocessing plants 
where spent fuel is reprocessed.  Only recently has attention been 
given to the production and release of this radionuclide at nuclear 
fuel cycle installations. 

(a)   Nuclear reactors

86.  The production of carbon-14 in nuclear power reactors is due 
to several nuclear reactions in the fuel, core construction 
materials and moderator.  Figure III.I summarizes the relevant 
reactions. 

FIGURE III

87.  Production rates depend upon the neutron flux, the shape of 
the respective neutron spectra and the resulting effective cross 
sections, on the amount of the target elements present in different 
reactor components and on the abundance of the target isotopes in 
the target elements.  The target elements are uranium, nitrogen, 
oxygen, and also carbon in the case of graphite moderated reactors.  

Nitrogen is present as an impurity in the fuel, as dissolved gas in 
the coolant, as nitrogen compounds sometimes used for pH control in 
the coolant, and as an impurity in structural materials.  Oxygen is 
present in water moderators and coolants, in CO2 coolants, and in 
oxide fuels (e.g., UO2). 

88.  The place of origin of 14C within a nuclear reactor has a 
strong influence on the discharge pathway.  One can basically 
distinguish between three locations of 14C generation, namely, 14C 
in the fuel, 14C in structural materials of the core (and solid 
moderator, if applicable) and 14C in the reactor coolant (and 
liquid moderator, if applicable). 

89.  The 14C produced in liquid or gaseous coolants will be present 
in different chemical compounds (CO2, CO, methane), depending on 
the chemical environment.  Under the influence of intensive 
radiation fields several chemical reactions may occur, influencing 
the chemical form of carbon-14.  The compounds in the coolant are 
released mainly together with off-gas and waste water from the 
coolant purification and treatment system.  Part of the carbon-14 
also leaks from the primary coolant circuit into the plant 
ventilation system and is released with ventilation air. 

90.  Significant reactions for the production of 14C in light water 
reactors (LWR) are:  (n) reactions with 17O present in the oxide 
fuel and in the moderator; (n, p) reactions with 14N present in the 
fuel as impurities; and ternary fissions. Ternary fission 
production per unit electrical energy generated is practically 
independent of reactor design, while the normalized production of 
14C by the other reactions depends on the enrichment of the fuel, 
the relative masses of the fuel and moderator, the concentration of 
nitrogen impurities in the fuel and the temperature of the fuel and 
moderator. 

91.  In boiling water reactors (BWR), the gaseous 14C is 
transported with the steam until it arrives at the turbine 
condenser.  There the gases are continuously withdrawn over a 
catalytic recombiner to burn the hydrogen and oxygen produced by 
radiolysis of the primary water.  Measurements have shown that one 
half or more of the total 14C produced in the coolant will be 
discharged in the form of CO2 together with the filtered gases from 
the turbine condenser.  There are other pathways of release of 14C, 
mainly caused by leakage from the primary circuit into the reactor 
building and the turbine hall.  These releases are also mainly in 
the form of CO2.  A part of the 14C remains dissolved in the 
primary water purification and treatment systems, causing smaller 
sources of release, for example in the auxiliary building and 
finally in the waste water system. 

92.  The primary circuit water of a pressurized water reactor (PWR) 
contains hydrogen in excess to recombine the oxygen produced by 
radiolysis.  Under such reducing conditions compounds like methane 
will be formed.  Therefore, contrary to the BWR, a PWR will release 
most of the 14C bound in hydrocarbons.  The main release pathways 
for gaseous compounds of 14C in PWRs are leakages of the primary 
water circuit into the containment air and the degasification of 
the primary water.  The escaping or withdrawn gases may be stored 

in decay tanks prior to release, and the gaseous 14C compounds can 
be oxidized to CO2 or released through charcoal beds.  Leakages may 
also arise in the auxiliary building from the primary water 
purification and treatment systems by way of degasing. Also, a part 
of the 14C compounds stays dissolved in the water and is released 
at the different steps of the waste water treatment. 

93.  The total environmental release of carbon-14 at the reactor, 
expressed as a fraction of the production rate, is on the average 
about 50% in BWRs and 30% in PWRs, but the value is quite variable, 
as has been shown by several recent monitoring programmes [R1, L1].  
UNSCEAR summarized the estimates of production in LWRs from several 
authors, the values being in the range 0.5 to 1.9 109 Bq per 
MW(e)a, and also derived an independent value of about 0.7 109 Bq 
per MW(e)a [U3]. 

94.  Carbon-14 is generated in heavy water reactors  (HWR) through 
reactions similar to those described for LWRs.  Owing mainly to the 
large moderator mass, the production rate of 14C in HWRs is 
expected to be considerably larger than in LWRs [U3].  The 
production rate in pressure vessel reactors is estimated to be 1.7 
1010 Bq per MW(e)a, with 90% generated in the moderator.  The 
production of 14C in CANDU reactors is estimated to be 1.6 1010 Bq 
per MW(e)a, 95% being produced in the moderator. 

95.  In gas-cooled graphite-moderated reactors (GCR), the major 
source of 14C production is the graphite moderator, due to 13C(n, 
gamma)14C reaction and also to the 14N(n, p)14C reaction based 
on the incorporated nitrogen impurity. Production rates have been 
estimated to be about 0.7 1010 Bq per MW(e)a in Magnox reactors and 
1.1 1010 Bq per MW(e)a in advanced gas-cooled reactors (AGR) [U3].  
Production of 14C in the carbon dioxide coolant, mainly from 
activation of nitrogen impurities and from the 17O(n, alpha)14C 
reaction, is a smaller source estimated to be about 108 Bq per 
MW(e)a for Magnox reactors and 4 108 Bq per MW(e)a for AGRs. 

96.  Carbon-14 discharges from Magnox reactors and AGRs result from 
coolant leakage and include 14C released to the coolant from 
corrosion of the moderator.  The fraction released at the reactor 
is about 3% in Magnox reactors and about 6% in AGRs, of the total 
production rate of 14C in these reactors [U3]. 

(b)   Fuel reprocessing plants

97.  While the 14C produced in the reactor coolant and moderator 
has a potential for immediate release at the nuclear reactor, the 
14C produced in the fuel will be released only later during nuclear 
fuel reprocessing.  Depending on reprocessing plant operation 
characteristics the release may be continuous or discontinuous.  
There are few measurements of 14C releases from reprocessing 
installations [S1], but it seems reasonable to assume that almost 
all the inventory of the fuel elements is released during the 
chemical dissolution of the fuel.  In the case of the Purex process 
the 14C is released in the form of CO2. 

(c)   Summary

98.  A very rough estimate can be made of the total production and 
release of 14C from nuclear fuel cycle installations, based on the 
average values given above.  Installed nuclear capacity worldwide 
in 1980 was 1.25 105 GW(e) [I2].  Assuming an average load factor 
for reactor operation of 0.6, the energy produced was 7.5 104 
GW(e)a.  Global production and release of 14C from reactor sites 
are thus estimated to be about 1.4 1014 Bq and 6 1013 Bq, 
respectively.  The estimated discharges by reactor types are given 
in Table III.1.  There are no estimates of production and release 
from other reactor types representing 10% of the total installed 
capacity.  The difference between production and reactor discharge 
estimates will largely represent the release from reprocessing 
plants, to the extent that the fuel is eventually reprocessed. 
Table III.1  Estimated global discharge of carbon-14
from nuclear power stations in 1980
------------------------------------------------------------------------
Reactor  Reactor  Capacity  Production rate   Release    Estimated
type     number   [MW(e)]   [Bq per MW(e)]    fraction   carbon-14
                                              (%)        discharge (Bq)
------------------------------------------------------------------------
PWR      96       64239     7 108             30         8 1012
BWR      62       35170     7 108             50         7 1012
HWR      14       5963      1.6 1010          70         4 1013
GCR      36       7086      9 109             5          2 1012
Other    33       12527     -                 -          -
------------------------------------------------------------------------
Total    241      124985                                 6 1013
------------------------------------------------------------------------

C.  BEHAVIOUR IN THE ENVIRONMENT

99.  Carbon-14 is present in atmospheric carbon dioxide, in the 
biosphere, and in the bicarbonates dissolved in the ocean.  The 
specific activity of natural 14C in the terrestrial biosphere, as 
measured in wood grown in the nineteenth century, was 0.227 ± 0.001
Bq per gram of carbon.  The Suess effect, accounting for a few 
percent decrease of specific activity at present, could reach a 
figure of the order of 20% in the year 2000 [U2], but is of little 
importance in the long range, when fossil fuel resources are 
exhausted. 

100.  Leaving aside the Suess effect, it has been suggested, 
however, that the present-day inventory does not correspond to the 
equilibrium value, but is increasing.  In fact, measurements of 
wood samples of known age show that only cyclic variations of 
atmospheric 14C, amounting to a few percent, have occurred in the 
past 6000 years [U2].  Two types of variations have been 
recognized:  one, with a time scale of the order of 100 years, has 
been explained by the solar wind modulation of the cosmic-ray flux 
density;  the other, with a time constant of more than 1000 years, 
may largely be due to a variation of the geomagnetic shielding of 
the earth. 

101.  Contrary to the case of natural carbon-14, the levels of man-
made carbon-14 are not at steady state in the different 
compartments of the environment.  Due to the very long mean life of 
carbon-14, continuing practices are not expected to last long 
enough to allow the environmental levels to reach the steady state.  
The predictions of the time-evolution of 14C levels in the 
atmosphere, biosphere and ocean after a release into the 
environment require, therefore, the use of compartment models. 

102.  Many models describing the dispersion of released 14C, and 
the subsequent exchange between the different compartments involved 
in the carbon cycle, have been proposed [C1, P1, N2, Y1, N3].  
UNSCEAR [U3] also developed a dynamic model for the assessment of 
doses from 14C released by nuclear explosions. This model includes 
compartments for the atmosphere and short-term biosphere, the 
terrestrial biosphere, the surface ocean and the deep ocean, and 
represents the thermocline layer in the ocean as a thick diffusion 
barrier. 

D.  TRANSFER TO MAN

103.  Carbon-14 released to the environment enters the carbon 
cycle, giving rise eventually to increased levels in man. From 
measurements of fallout carbon-14, it was noted that the specific 
activity in human tissue comes into equilibrium with that of 
atmospheric CO2 with a delay time of about 1.4 years [N5]. 

104.  Intake of carbon by man is primarily from diet. Ingestion 
intake is of the order of 300 g d-1 with nearly complete 
absorption, whereas inhalation intake is about 3 g d-1 with only 1% 
retained in the body [U3].  The total carbon content of the body is 
1.6 104 [I1].  The quotient of this with the intake rate gives an 
estimated mean residence time of carbon in the human body of 53 
days. 

105.  Man comes, therefore, into fairly rapid equilibrium with 
carbon-14 in his immediate environment.  It is generally sufficient 
in carbon-14 dose calculations to adopt a steady-state model which 
assumes that the specific activity of carbon in tissues is in 
equilibrium with that in air and in the diet. 

E.  DOSIMETRY

1.  Dose per unit intake

106.  An intake of carbon-14 at a specific concentration of 0.23 Bq 
per gram of  carbon, corresponding to the present value for natural 
carbon-14, gives rise to the following absorbed dose rate averaged 
over the whole body 

          Bq           Gy g   0.049 MeV/Bq s
     0.23 -- 1.6 10-10 ----   ----------------
          gc           MeV       7 104g

     3.15 107 s/a 1.6 104 gc = 13 µGy a-1

The dose rates in individual tissues depend on their carbon 
concentrations.  The carbon content per unit mass averages 23% for 
the whole body, but ranges from 9% in gonads and 10% in lungs to 
41% in red bone marrow and 67% in adipose tissue [I1].  The annual 
absorbed doses are 5 µGy in gonads, 6 µGy in lungs, 20 µGy in bone-
lining cells and 22 µGy in red bone marrow [U3].  The tissue-
weighted annual effective dose equivalent from natural carbon-14 is 
12 µSv. 

107.  This dose is due almost entirely to ingestion intake of 
carbon-14.  If the carbon intake rate is 300 g d-1 at the specific 
activity of 0.23 Bq g-1, the intake rate of 14C is 69 Bq d-1.  The 
effective dose equivalent per unit ingestion intake of 14C is 

            12 10-6 Sv/a     1 a
            ------------   -------  =  5.2 10-10 Gy Bq-1
               69 Bq/d      365 d

The dose factor for inhalation intake is less by a factor of 10-2, 
since absorption into the body is that much less by this pathway. 

2.  Dose per unit release

108.  The doses given above for natural carbon-14 correspond to
the annual global production of 1015 Bq.  This production is
essentially constant in time and uniform over the world. Therefore, 
equilibrium has become established.  The effective dose equivalent 
commitment per unit release is 

     12 10-6 Sv/a
     ------------  =  1.2 10-20 Sv Bq-1
      1015 Bq/a


The collective dose equivalent rate from natural carbon-14 to the 
world population of 4 109 people is 4.8 104 man Sv a-1. 

109.  The assessment of the dose commitment from a given release of 
man-made carbon-14 is carried out by direct analogy with natural 
carbon-14.  Once the released carbon-14 enters the global carbon 
cycle, the effective dose equivalent commitment per unit release is 
1.2 10-20 Sv Bq-1.

110.  It is difficult to assess with precision the collective dose 
commitment per unit release of carbon-14, because the projected 
increase in the world population is very uncertain. Assuming that 
it will attain an equilibrium value of 1010 persons, in a time 
short compared with the mean effective life of 14C [U3], the 
collective effective dose equivalent commitment per unit released 
is approximately 1.2 10-10 man Sv per Bq. 

111.  In order to calculate the complete collective dose commitment 
[U3] required for assessments of maximum future mean annual doses 
from a continuing but finite practice releasing 14C, it is 
necessary to use dynamic models predicting the time evolution of 
environmental levels. Assuming that power production by nuclear 

fission will last for a few hundred years (for example, 500 years), 
the incomplete collective dose commitment can be calculated using 
the model with diffusion barrier already mentioned.  The incomplete 
collective dose commitment, integrated over 500 years, is about 2.3 
10-11 man Sv per Bq released.  This value is somewhat higher than a 
value of about 1.4 10-11 man Sv per Bq which can be deduced from a 
recent assessment of the environmental significance of 14C [N3], 
but in view of the uncertainties involved, the difference is 
probably insignificant. 

112.  The contribution of local and regional exposures to the 
collective dose commitment is very small, of the order of a 
percent, and can be neglected [N3].  The assessment of individual 
doses at some selected locations, however, is necessary for 
radiation protection purposes.  Its calculations can be carried out 
by the use of specific activity methods. One simple model assumes 
that the specific activity of 14C in air is equal to that in the 
body.  A more sophisticated calculation assumes that the specific 
activity in the vegetation at the location of interest is equal to 
that of air.  The dose can then be assessed from knowledge of the 
relative proportion of contaminated food in the diet.  Both methods 
require the use of micrometeorological models to assess 
quantitatively the dispersion of 14C from the release point to the 
locations of interest.  Some publications [U4, N4, C2], present 
improvements to the classical formulations describing the local 
atmospheric dispersion. 

F.  REFERENCES

C1  Craig, H.  The natural distribution of radiocarbon and the 
    exchange time of carbon dioxide between atmosphere and sea.  
    Tellus 9: 1-17 (1957). 

C2  Clarke, R.  A model for short and medium range dispersion of 
    radionuclides released to the atmosphere.  A first report of a 
    working group on atmospheric dispersion. NRPB-R91 (1979). 

I1  International Commission on Radiological Protection. Report of 
    the task group on reference man.  International Commission on 
    Radiological Protection publication 23 (1975). 

I2  International Atomic Energy Agency.  Power reactors in member 
    states.  IAEA, Vienna, 1980. 

L1  Luykx, F. and G. Fraser.  Radioactive effluents from nuclear 
    power stations and nuclear fuel reprocessing plants in the 
    European community:  discharge data 1962-76.  Radiological 
    aspects.  Commission of the European Communities.  V/4604/78-EN 
    (1978). 

N1  National Council on Radiation Protection and Measurements.  A 
    handbook of radioactivity measurements procedures.  National 
    Council on Radiation Protection report No. 58 (1978). 

N2  Nydal, R.  Further investigation on the transfer of radiocarbon 
    in nature.  J. Geophys. Res. 73: 3617-3635 (1968). 

N3  Nuclear Energy Agency, OECD.  Radiological significance and 
    management of H-3, C-14, Kr-85 and I-129 arising from the 
    nuclear fuel cycle.  Report by an NEA group of experts. 
    OECD/NEA (1980). 

N4  NRPB and CEA.  Methodology for evaluation of radiological 
    consequences of radioactive effluents released in normal 
    operations.  Commission of European Communities. V/3865/79 
    (1979). 

N5  Nydal, R., K. Lovseth and O. Syrstad.  Bomb 14-C in the human 
    population. Nature 232:  418-421 (1971). 

P1  Plesset, M. and A. Latter.  Transient effects in the 
    distribution of carbon-14 in nature.  Proceeding of the 
    National Academy of Sciences 46:  232-241 (1960). 

R1  Riedel, H. and P. Gesewsky.  Zweiter Bericht über Messungen zur 
    Emission von Kohlenstoff-14 mit der Abluft aus Kernkraftwerken 
    mit Leichtwasserreaktor in der Bundesrepublik Deutschland.  
    Bundesgesundheitsamt report STH-13/77 (1978). 

S1  Schuettelkopf, H. and G. Herrman.  14-CO2  Emissionen aus wer 
    Wiederaufarbeitungsanlage Karlsruhe. p.  189  in Report for the 
    Commission of the European Communities. V/2266/78-D (1978). 

T1  Telegadas, K.  The seasonal atmospheric distribution and 
    inventories of excess carbon-14 from March 1955 to July 1969.  
    HASL-243 (1971). 

U2  United Nations.  Report of the United Nations Scientific 
    Committee on the Effects of Atomic Radiaton to the General 
    Assembly, with annexes.  Volume I:  Levels, Volume II: Effects.  
    United Nations sales publication No. E.72.IX.17 and 18.  New 
    York, 1972. 

U3  United Nations.  Sources and Effects of Ionizing Radiation.  
    United Nations Scientific Committee on the Effects of Atomic 
    Radiation 1977 report to the General Assembly, with annexes.  
    United Nations sales publication No. E.77.IX.I.  New York, 
    1977. 

U4  U.S. Nuclear Regulatory Commission.  Regulatory Guide 1.111 
    (1977). 

Y1  Young, J. and A. Fairhall.  Radiocarbon from nuclear weapons 
    test.  J. Geophys.  Res. 73:  1185-1200 (1968). 

IV.  KRYPTON-85

A.  INTRODUCTION

113.  Krypton is element number 36 in the periodic table.  It 
belongs to the group of inert gases together with helium, neon, 
argon, xenon and radon.  It occurs naturally in the atmosphere to 
an estimated extent of 1 to 2 10-6 by volume. 

114.  The naturally occurring stable krypton isotopes and their 
atom percentage abundances are:  78Kr (0.35%), 80Kr (2.27%), 82Kr 
(11.56%), 83Kr (11.55%), 84Kr (56.9%), 86Kr (17.37%) [N1].  The 
radioactive isotopes of krypton include mass numbers of 74-77, 79, 
79m, 81, 81m, 85, 85m, 87-95 and 97.  Some of these occur naturally 
in low trace amounts as a result of cosmic ray induced reactions 
with stable krypton isotopes and by spontaneous fission of natural 
uranium. 

115.  The radioactive isotope 85Kr is produced in nuclear fission.  
With a half-life of 10.7 years, it can become widely dispersed in 
the atmosphere following release.  The average fission yields 
differ by about a factor of 2 for 239Pu and 235U, being about 0.6 
and 1.3 atoms per 100 fissions, respectively (Table IV.1). 

Table IV.1  Fission yields of 
krypton-85 [C2]
-----------------------------
            Fission yield (%)
Nuclide     thermal    fast
-----------------------------
232Th                  4.14
233U        2.28       2.12
235U        1.32       1.33
238U                   0.74
239Pu       0.558      0.62
-----------------------------

116.  The decay scheme of 85Kr is presented in Figure IV.I. Two 
beta particles and a single gamma photon are emitted, along with 
several low-energy conversion electrons and x-rays. 

117.  Being chemically inert, krypton and other inert gases are not 
usually involved in biological processes.  They are, however, 
dissolved in body fluids and tissues following inhalation.  Krypton 
is characterized by low blood solubility, high lipid solubility and 
rapid diffusion in tissue [K1].  The biological involvement of 
inert gases has been noted in numerous studies [K1]. 

FIGURE IV


B.  SOURCES

118.  Krypton-85 is produced by cosmic ray interactions in the 
atmosphere, in nuclear power reactors, and nuclear explosions.  The 
main release source is the dissolution step in the reprocessing of 
nuclear fuel. 

119.  Concentrations of 85Kr in the atmosphere increased sharply 
after 1955 due to the production and testing of nuclear weapons and 
the development of the nuclear power industry.  More recently the 
input rates of 85Kr into air have decreased [H2].  There have been 
reductions in plutonium production for military purposes and in 
nuclear fuel reprocessing. 

120.  A review of 85Kr measurement data for 1950-77 has been 
prepared by Rozanski [R1].  The most recent data indicate that 
concentrations in air have stablized at about 0.6 Bq/m3 in the 
northern hemisphere and 0.4 Bq/m3 in the southern hemisphere [R1].  
The major sources are in the northern hemisphere, accounting for 
the higher levels in that hemisphere. 

1.  Natural krypton-85

121.  Krypton-85 is present in small amounts in the environment as 
a result of spontaneous fission of natural uranium and interactions 
of cosmic ray neutrons with atmospheric 84Kr.  The steady state 
environmental inventories of 85Kr from these sources have been 
calculated:  7.4 1010 Bq in the upper 3 m of the total land and 
water surface due to spontaneous fission of natural uranium, 3.7 
1011 Bq in the atmosphere from cosmic ray production and 3.7 105 Bq 
in the oceans from the atmospheric source [D1].  These estimates, 

in comparison with the estimates of man-made sources of 85Kr to 
follow, are negligible in contributing to the world's total 85Kr 
inventory.

2.  Nuclear explosions

122.  Since 85Kr is produced during fission, it has been generated 
by nuclear weapon tests.  The total amount of 85Kr produced in 
nuclear testing can be calculated from the ratio of 85Kr/90Sr 
fission yield of 0.08, giving an activity ratio of 0.22 [C2].  
Measurements of 90Sr activity have been reported and discussed in 
the reports of UNSCEAR [U1-U7]. There have been 6 1017 Bq of 90Sr 
produced in weapon testing through 1976 [U7], corresponding to 
about 1.3 1017 Bq of 85Kr. 

123.  Another source of 85Kr associated with nuclear weapons is in 
the production of plutonium in military reactors.  The amount of 
85Kr released from this source is estimated to be two times higher 
than that from the weapon tests [D1].  Naval propulsion reactors 
also contribute to the 85Kr inventory with an annual production in 
the region of 1.1 to 1.9 1016 Bq [B1].  Including all sources, the 
total amount of 85Kr produced in operations for military purposes 
is still rather small in comparison to the prospective generation 
of 85Kr by the nuclear power industry. 

3.  Nuclear fuel cycle

124.  Krypton-85 is produced by fission in the fuel of nuclear 
reactors and in very low trace amounts in the moderator or coolant, 
due to contamination with fissile material.  The rates of 85Kr 
production are related to the type of fuel and degree of burn-up.  
Production and emission rates may be conveniently normalized to 
unit electrical energy generated (for power reactors) or to the 
electrical energy generated by the reactors serviced (for fuel 
reprocessing plants). 

125.  The amounts of 85Kr produced vary according to reactor type.  
For thermal reactors, the range of estimated production is about 
1.1 to 1.5 1013 Bq/MW(e)a.  For FBRs the values are about 25% 
smaller [E1, M1], for HTGRs 50% higher [B3].  A production rate of 
1.4 1013 Bq/MW(e)a has been correlated with some measurements from 
reprocessing plants [U7] and this value can be taken for general 
evaluations. 

126.  An estimate of 85Kr annual generation from reactor operation 
can be obtained from the installed capacity of nuclear reactors of 
1.25 105 MW(e) worldwide in 1980 [I1], with the assumptions of 60% 
utilization and average 85Kr generation rate of 1.4 1013 Bq/MW(e)a: 

   1.25 105 MW(e)a x 0.6 x 1.4 1013 Bq/MW(e)a = 1 1018 Bq/a

The actual release rate is less, since delays occur before 
reprocessing and not all fuel is reprocessed. 

127.  Reported releases of 85Kr and other fission noble gases were 
listed in the 1977 report of UNSCEAR [U7].  There are large 
differences in the release values of the various plants.  Although 

the relevant data are not very extensive, there are indications of 
improved retention of 85Kr at reactors in recent years due to the 
installation of additional hold-up tanks or adsorption columns. 

128.  In the reprocessing plant the spent fuel elements are 
dismantled and the nuclear material dissolved.  Procedures to 
separate 85Kr from gaseous effluents and to provide long-term 
retention are under study, but current practice is to allow 
controlled release to the atmosphere. 


C.  BEHAVIOUR IN THE ENVIRONMENT

129.  Krypton-85 discharged to the environment disperses in the 
atmosphere and largely remains there until decay.  It can become 
washed out by rain and diffuse into surface layers of soil and 
oceans, but these processes account for very little transfer of 
85Kr from the atmosphere. 

1.  Dispersion in the atmosphere

130.  Materials released to the atmosphere are transported downwind 
and dispersed according to atmospheric mixing processes.  The 
estimation of this dispersion is commonly approached by using a 
diffusion-transport equation.  Several models have been developed 
for this purpose, using a variety of boundary conditions and 
simplifying assumptions.  Most of them are based on the Gaussian 
plume diffusion model [S1, I2], which has been shown to be adequate 
in many practical situations.  The krypton concentrations in air at 
various distances for a release from a 30 m high stack are shown in 
Table IV.2 [C5]. 

Table IV.2  Krypton-85 
concentration in air for a 
release of 1 Bq/s (stack height 
30 m, Pasquill category D) 
[C5]
-------------------------------
Distance        Concentration
(km)            (Bq/m3)
-------------------------------
1               4.8 10-7
10              1.3 10-8
100             4.4 10-10
1000            3.2 10-11
-------------------------------

131.  For estimation of dispersion at greater distances, some 
shortcomings in the Gaussian model are evident in the assumptions 
that the meteorologic conditions and the direction of the wind 
remain constant throughout the transit of the plume.  To overcome 
these difficulties, long-range models have been developed [A1, D3, 
M2], which follow the trajectories of masses of air passing over 
the release point and take into account the changing meteorologic 
conditions with time.  A survey of several diffusion models and of 
their applications is given in [C5]. 

132.  The global circulation of 85Kr can be approximated by a 
simple compartment model, consisting of single compartments 
representing the atmosphere in the northern and in the southern 
hemispheres.  Following a single release, equilibrium 
concentrations in the atmosphere are achieved after about two 
years.  Further decrease in concentrations is due to radioactive 
decay.  In applying this model, Kelly et al. [K3] determined that 
the integral concentration in air would be 5.3 10-18 Bq a m-3 per 
Bq released.  The atmospheric mass was assumed to be 3.8 1021 g, 
equivalent to 3.1 1018 m3 at STP. 

133.  The dispersion calculations of Machta et al. [M2] are based 
on detailed meteorological considerations and allow population-
weighted exposures to be determined.  Table IV.3 lists the average 
surface air concentrations of 85Kr in latitude bands following 
release of 1 Bq in the 30-50° N latitude band.  Uniform 
concentrations are achieved after two years, after which the 
integral concentration until complete decay is 

                   10.73 a
    22 10-20 Bq/m3 -------  =  3.4 10-18 Bq a/m3
                    1n 2

Adding the contributions from the first two years gives
3.9 10-18 Bq a/m3 for the population weighted integral
concentration of 85Kr in air from a release of 1 Bq.

2.  Removal from the atmosphere

134.  There is very little removal of 85Kr from the atmosphere, 
except by radio-active decay.  The low solubility of krypton in 
water limits the accumulation of 85Kr in rainwater.  Adsorption of 
85Kr on particulate matter in air and subsequent deposition of the 
particles provides a removal means of very low efficiency [N1]. 

135.  The transfer of 85Kr to soil can occur by diffusion 
processes;  however, estimates of this transfer can account for 
only about 0.05% of the total krypton in the atmosphere [N1]. 
Therefore, soil in general is not an important removal sink for 
85Kr. 

136.  The efficiency of the oceans as a sink for 85Kr can be 
determined from the natural krypton content of the atmosphere and 
of the mixed layer of the ocean.  From estimates of the krypton 
concentration in air, the atmospheric volume and the density 
krypton (STP), a total mass of about 1.64 1016 g of krypton in the 
atmosphere is calculated [N1].  Assuming that the mixed layer of 
the ocean extends to 100 m depth and an area of 3.6 1018 cm2, and 
using the measured average krypton concentration in this layer of 
seawater of 5 10-8 by volume [B4], a total mass of 6.7 1012 g of 
krypton in the mixed layer of the ocean is obtained.  This 
corresponds approximately to 0.04% of the atmospheric mass of 
krypton. 

Table IV.3  Average surface air concentration of krypton-85
(1 Bq emitted uniformly over one year in 30-50° N latitude band)
[M2]
---------------------------------------------------------------
                      Krypton-85 concentration   Population 
Latitude band                (10-20 Bq/m3)       distribution % 
                      Year 1    Year 2  Year 3
---------------------------------------------------------------
70 - 90° N            23        32      22        -
50 - 70° N            25        31      22        12.6
30 - 50° N            23        30      22        32.0
10 - 30° N            19        27      22        39.0
10° N - 10° S         11        22      22        11.5
10 - 30° S            6.3       22      22        3.4
30 - 50° S            5.1       20      22        1.5
50 - 70° S            4.3       19      22        0.05
70 - 90° S            3.8       19      22        -

Population weighted
integral 
concentration
(10-20Bq a/m3)        19.5      27.6    22.0
---------------------------------------------------------------

137.  An estimate of the total mass of krypton in the oceans as a 
whole is obtained using an average concentration by volume of 
krypton in the oceans of 9 10-8 [B4], a total ocean volume of 1.4 
1024 cm3, and a krypton density of 3.73 10-3 g/cm3 at STP.  This 
calculation results in a total ocean inventory of about 4.7 1014 g
of krypton, or approximately 3% of the total atmospheric krypton 
[N1].  These figures clearly indicate that the oceans can serve 
only as a minor sink for 85Kr discharged into the atmosphere. 

D.  TRANSFER TO MAN

138.  Following release to the atmosphere 85Kr becomes widely 
dispersed.  Exposure of man occurs by external irradiation from the 
passing cloud or the dispersed gas and by internal irradiation 
following inhalation of 85Kr and absorption in tissues. 

139.  After intake, 85Kr is distributed in the body by blood and 
lymph fluids and is absorbed in the various tissues.  A person 
immersed in an atmosphere of 85Kr at low concentration would rather 
quickly come into equilibrium with it.  The concentrations in body 
tissues are determined by multiplying the concentration in air by a 
partitioning factor, called the Ostwald's coefficient.  The 
relevant values reflect the rate at which tissues are perfused with 
blood, the solubility of the gas in the several tissues and the 
velocity of diffusion of krypton across anatomical boundaries.  The 
concentration of 85Kr in the body is not uniform, the concentration 
in the adipose tissue being nearly 50 times higher than that in 
other parts of the body. 

140.  As a first approximation, one may only account for a 
difference in the absorption behaviour of krypton in fat and non-
fat tissues, with values of the Ostwald coefficient of 0.45 for fat 

and 0.07 for non-fat tissue [N1].  Other more elaborate models use 
weight-related coefficients, where the density of the absorbing 
tissue is taken into account [S2]. 

141.  The total body retention of 85Kr has been subjected to 
exponential analysis.  Several clearance rates have been 
recognized.  Recent work has suggested a model for krypton in the 
body consisting of five compartments [C6].  The fastest component 
probably represents the clearance from circulating blood, 
particularly blood plasma (T´ = 21.5 ± 5.7 s).  The second 
component (4.74 ± 2 min) appears to be representative of 
haemoglobin clearance.  The next slower component (19.8 ± 6.6 min) 
is most likely related to clearance of krypton from muscle.  The 
two components with the slowest clearance rates can be related to 
body fat compartments.  A half-time of about 2.4 h is attributed to 
a fat compartment not located in adipose tissue.  The retention 
half-time of krypton in adipose tissue is the slowest component and 
is correlated significantly with the total body fat content.  The 
relationship is T´(h) = 0.17 (percentage fat) + 0.75 [C6]. 

E.  DOSIMETRY

142.  Krypton-85 released to the environment causes a radiation 
dose to man through external irradiation from amounts in air and 
through internal irradiation from amounts within the body.  Tissues 
are irradiated both from the activity in the organ itself and from 
the activity present in the surrounding organs. 

1.  Dose per unit exposure

143.  The equilibrium absorbed dose rates to body organs per unit 
concentration of krypton-85 in air are summarized in Table IV.4 
[N1].  For comparison, the recently published values of the ICRP 
are also listed [I3].  The ICRP values represent minor adjustments, 
except for the lungs, for which the beta dose due to 85Kr in the 
airways of the lungs has been disregarded. 

144.  The dose equivalent rates in various organs are listed in 
Table IV.5.  These are the ICRP values [I3].  The quality factor 
for 85Kr radiation is one.  Therefore the dose equivalent rates are 
numerically equal to the absorbed dose values.  When combined with 
the tissue weighting factors suggested by the ICRP to account for 
varying incidence of health effects, the effective dose equivalent 
rate is obtained, which for 85Kr is 8 10-9 Sv/a per Bq/m3. 


Table IV.4  Equilibrium absorbed dose rate to body organs per unit air concentration from 
immersion in a semi-infinite cloud of krypton-85 (10-9  Gy/a per Bq/m3) [N1]
------------------------------------------------------------------------------------------
Source                                                 Organ                              
                            Skin     Adipose  Lungs   Red bone  Skeleton  Ovaries  Testes
                                     tissue           marrow
------------------------------------------------------------------------------------------
 Krypton-85 in air
Photons in air              4.1      3.2      3.0     3.8       4.1       1.3      3.5
Betas in air                490.0    -        -       -         -         -        -
Bremsstrahlung in air       0.6      0.57     0.51    0.97      1.1       0.30     0.68
Bremsstrahlung in skin      0.015    0.0018   0.0011  0.0010    0.0030    0.0006   0.0027

 Krypton-85 in the body
Photons in the body         0.0006   0.0006   0.0006  0.0007    0.0006    0.0008   0.0008
Betas in the body           0.10     0.30     0.10    0.21      0.10      0.10     0.10
Bremsstrahlung in the body  0.0001   0.0002   0.0002  0.0003    0.0003    0.0002   0.0003
Betas in airways of lung    -        -        4.9     -         -         -        -
------------------------------------------------------------------------------------------
Total                       490      4.1      8.5     5.0       5.3       1.7      4.3
------------------------------------------------------------------------------------------
Total [I3]                  410               3.8     5.0       5.4       4.6        
------------------------------------------------------------------------------------------

Table IV.5  Dose equivalent rates from submersion in 
semi-infinite cloud of 85Kr (10-9 Sv/a per Bq/m3) [I3]
------------------------------------------------------------
                        Dose         Weighting   Effective 
                        equivalent   factor      dose
                        rate                     equivalent 
                                                 rate
------------------------------------------------------------
Gonads                  4.6          0.25        1.1
Breast                  3.9          0.15        0.6
Red bone marrow         5.0          0.12        0.6
Lungs                   3.8          0.12        0.46
Bone surface            5.4          0.03        0.16
Spleen                  4.0          0.06        0.24
Small intestinal wall   3.8          0.06        0.23
Kidneys                 3.5          0.06        0.21
Adrenal glands          3.5          0.06        0.21
Liver                   3.3          0.06        0.20
Skin                    410.0        0.01        4.1
------------------------------------------------------------
Total                                            8.1
------------------------------------------------------------

2.  Dose per unit release

145.  The collective effective dose equivalent commitment per unit 
release of 85Kr into the atmosphere, Sc1, can be determined, based 
on the procedures previously used by UNSCEAR [U7].  In the local

region of the release, the following formula can be applied 

     c    x                  100 km   r    -1.5
    S1 = (Q)1 km deltaN phi /       (1 km)    2 pi r dr
                             1

A dispersion factor, x/Q, of 5 10-7 s/m3 is assumed, which agrees 
with the data of Table IV.2.  The population density, deltaN is 
assumed to be 100 man/km2 in the local region extending to 100 km 
distance.  The dose factor, phi is 8 10-9 Sv/a per Bq/m3 as given 
above.   The concentration of 85Kr in air is assumed to decrease as 
a function of the distance, r, from the release point.  The 
integral of the distance dependence over the local region times the 
dispersion factor, the population density, and the dose factor 
gives the collective effective dose equivalent commitment per unit 
85Kr activity released. The result is 

     c
    S1 (local) = 1 10-18 man Sv/Bq

146.  The collective dose equivalent commitment per unit release of 
85Kr to the global population can be determined from the following 
formula [U7] 

     c    infinite .---  
    S1 = /         D(t)  N(t)dt
          0
      .---
where D(t) is the per caput dose rate per unit activity released 

and N(t) is the population size. 

147.  After uniform mixing in the global atmosphere, the per caput 
dose rate from 85Kr is equal to the product of the concentration in 
air, Co, which decreases due to radioactive decay of 85Kr, and the 
dose factor, phi.  The population is assumed to increase at a rate, 
nu, of 2% per year.  The collective effective dose equivalent 
commitment is, thus, 
                               
     c    infinite   -lambda t         nu t     C0 phi N0
    S1 = /        C0e          phi  N0e    dt = ---------
          0                                     lambda-nu

148.  The population-weighted surface air concentrations of Table 
IV.3 can be used to estimate the global collective dose.  An 
initial world population of 4 109 is assumed.  The following 
contributions are obtained in the first two years following the 
release of 85Kr to the atmosphere: 

 c                                    Bq a/m3         Sv/a
S1 (1st year) = 4 109 man 19.5 10-20 -------- 8 10-9  ------
                                        Bq            Bq/m3
              = 6 10-18 man Sv/Bq

 c                                       Bq a/m3         Sv/a
S1 (2nd year) = 4.08 109 man 27.6 10-20 -------- 8 10-9  ------
                                           Bq            Bq/m3
              = 9 10-18 man Sv/Bq

Thereafter, the formula in the preceding paragraph can be applied. 

                                        Bq a/m3         Sv/a
                 4.16 109 man 22 10-20 -------- 8 10-9  -----
 c                                        Bq            Bq/m3
S1 (> 2 years) = -----------------------------------------------
                                1n 2
                               -------   - 0.02 1/a
                               10.73 a

               = 1.6 10-16 man Sv/Bq

The total global collective dose equivalent commitment per unit 
85Kr activity released is 

     c
    S1 (global) = 1.8 10-16 man Sv/Bq

149.  A more approximate estimation, namely assuming instant mixing 
of the 85Kr in the global atmosphere of 4 1018 m3, gives the same 
result.  The initial concentration is then 25 10-20 Bq/m3.  Using 
the formula in paragraph 147, with an initial population of 4 109, 
gives the value of 1.8 10-16 man Sv/Bq for the global collective 
dose equivalent commitment per unit 85Kr activity released. 

150.  It is seen by comparison that the local contribution to the 
collective dose equivalent commitment per unit release of krypton-
85 is negligible.  Therefore, the total dose estimate is 
independent of the location of the release. 

E.  REFERENCES

A1  Apsimon, H.M. and A.J.H. Goddard.  Modelling the atmospheric 
    dispersal of radioactive pollutants beyond the first few hours 
    of travel. p. 124-135  in The seventh International Technical 
    Meeting on Air pollution Modelling and its Application.  
    Proceedings of a symposium.  Airlie, Va., U.S.A., 1976. 

B1  Bernhardt, D.C., A.A. Moghissi and J.W. Cochran. Atmospheric 
    concentrations of fission product noble gases. p. 97-118  in 
    The Noble Gases (A.A. Moghissi and J.W. Cochran, eds.).  U.S. 
    Government Printing Office, Washington, 1975. 

B3  Bonka, H., R. Schulten, K. Brüssermann et al.  Zukünftige 
    radioaktive Umweltbelastung in der BRD durch radionuklide aus 
    kerntechnischen Anlagen im Normalbetrieb.  Berichte der 
    Kernforschungsanlage Jülich, Jül-1220 (1975). 

B4  Bieri, R.H., M. Koide and E.D. Goldberg.  The noble gas 
    contents of Pacific seawaters.  J. Geophys. Res. 71: 5243-5247 
    (1966). 

C2  Crouch, E.A.C.  Fission product yields from neutron induced 
    fission, atomic data and nuclear data tables, Vol. 19, No. 5.  
    Academic Press Inc., New York, 1977. 

C5  Commission of the European Communities.  Methodology for 
    evaluation the radiological consequences of radioactive 
    effluents released in normal operations.  V/3865/79 - EN, FR 
    Directorate of Health Protection (1979). 

C6  Cohn, S.H., K.J. Ellis and H. Susskind.  Evaluation of the 
    health hazard from inhaled krypton-85.  International Symposium 
    on biological implications of radionuclides released from 
    nuclear industries.  IAEA-SM-237/46 (1979). 

D1  Diethorn, W.S. and W.L. Stockho.  The dose to man from 
    atmospheric krypton-85.  Health Phys. 23:  653-662 (1972). 

D3  Depres, A. and J. LeGrand.  Une méthode d'évaluation des 
    transferts atmosphériques ŕ longue distance. p. 237-249  in La 
    dispersion en milieu physique naturel.  Proceedings of a 
    seminar.  Cadarache, France, 1978. 

E1  Erdman, C.A. and A.B. Reynolds.  Radionuclide behaviour during 
    normal operations of liquid metal cooled fast breeder reactors, 
    part 1:  Production.  Nucl. Saf. 16: 43-52 (1975). 

H2  Heller, D., W. Roedel, K.O. Münnich et al.  Decreasing release 
    of krypton-85 into the atmosphere. Naturwissenschaften 64(7):  
    383 (1977). 

I1  International Atomic Energy Agency.  Power reactors in member 
    states IAEA,  Vienna, 1980. 

I2  Islitzer, N.F. and  D.H. Slade.  Diffusion and transport 
    experiments in meteorology and atomic energy.  U.S. Atomic 
    Energy Commission report USAEC-TID-24190 (1968). 

I3  International Commission on Radiological Protection. Limits for 
    intakes of radionuclides by workers. International Commission 
    on Radiological Protection report 30, supplement to part I.  
    Pergamon Press, Oxford, 1979. 

K1  Kirk, W.P.  Krypton-85:  A review of the literature and an 
    analysis of radiation hazards, USEPA report EPA-NP-19251 
    (1972). 

K3  Kelly, G.N., J.A. Jones, P.M. Bryant et al.  The predicted 
    radiation exposure of the population of the European community 
    resulting from discharge of krypton-85, tritium, carbon-14 and 
    iodine-129 from the nuclear power industry to the year 2000.  
    Commission of the European Communities, Directorate of Health 
    Protection.  V/2676/75 (1975). 

M1  Martin, A. and M. Apsimon.  The forecasting of radioactive 
    wastes arising from nuclear fuel reprocessing. p. 55-60  in 
    Management of Radioactive Wastes from Fuel Reprocessing. 
    Proceedings of a symposium jointly organized by OECD/NEA and 
    International Atomic Energy Agency.  Paris, 1972. 

M2  Machta, L., G.J. Ferber and J.L. Hefter.  Regional and global 
    scale dispersion of krypton-85 for population dose 
    calculations.  p. 411-424  in Physical Behaviour of Radioactive 
    Contaminants in the Atmosphere.  International Atomic Energy 
    Agency publication STI/PUB/355.  Vienna, 1974. 

N1  National Council on Radiation Protection and Measurements.  
    Krypton-85 in the atmosphere - accumulation, biological 
    significance and control technology. NCRP No. 44 (1975). 

R1  Rozanski, K.  Krypton-85 in the atmosphere 1950-1977:  A data 
    review.  Envir. International 2:  139-143 (1979). 

S1  Sutton, O.G.  The theory of Eddy diffusion in the atmosphere. 
    Proc. R. Soc. (London), Ser. A 135:  143-152 (1932). 

S2  Snyder, W.S., L.T. Dillman, M.R. Ford et al.  Dosimetry for a 
    man immersed in an infinite cloud of 85-K5, p. 119-122  in The 
    Noble Gases (A.A. Moghissi and R.E. Stanley eds.).  U.S. 
    Government Printing Office, Washington, 1975. 

U1  United Nations.  Report of the United Nations Scientific 
    Committee on the Effects of Atomic Radiation.  Official Records 
    of the General Assembly, Thirteenth Session, Supplement No. 17 
    (A/3838).  New York, 1958. 

U2  United Nations.  Report of the United Nations Scientific 
    Committee on the Effects of Atomic Radiation.  Official Records 
    of the General Assembly, Seventeenth Session, Supplement No. 16 
    (A/5216).  New York, 1962. 

U3  United Nations.  Report of the United Nations Scientific 
    Committee on the Effects of Atomic Radiation.  Official Records 
    of the General Assembly, Nineteenth Session, Supplement No. 14 
    (A/5814).  New York, 1964. 

U4  United Nations.  Report of the United Nations Scientific 
    Committee on the Effects of Atomic Radiation.  Official Records 
    of the General Assembly, Twenty-first Session, Supplement No. 
    14 (A/6314).  New York, 1966. 

U5  United Nations.  Report of the United Nations Scientific 
    Committee on the Effects of Atomic Radiation.  Official Records 
    of the General Assembly, Twenty-fourth Session, Supplement No. 
    13 (A/7613).  New York, 1969. 

U6  United Nations.  Ionizing Radiation:  Levels and Effects. 
    Report of the United Nations Scientific Committee on the 
    Effects of Atomic Radiation to the General Assembly, with 
    annexes.  United Nations sales publication No. E.72.IX.17 and 
    18. New York, 1972. 

U7  United Nations.  Sources and Effects of Ionizing Radiation.  
    Report of the United Nations Scientific Committee on the 
    Effects of Atomic Radiation 1977 report to the General 
    Assembly, with annexes.  United Nations sales publication No. 
    E.77.IX.1.  New York, 1977. 

V.  STRONTIUM-90

A.  INTRODUCTION

151.  Strontium is element number 38 in the periodic table. It is 
an alkaline earth element and is therefore similar to calcium, 
barium and radium.  It follows calcium through the food chains from 
environment to man, but some degree of discrimination exists 
against strontium.  Both strontium and calcium are retained in the 
body largely in bone. 

152.  Since the early days of atmospheric nuclear testing the 
importance of 90Sr as a contributor to the radiation exposure of 
man has been recognized.  Strontium-90 is a radionuclide formed in 
the process of nuclear fission.  It has a radioactive half-life of 
29.1 years and decays by beta emission.  Its daughter, 90Y, is also 
radioactive with a half-life of 64.0 hours, and decays by beta 
emission to the stable isotope 90Zr.  The decay scheme for 90Sr and 
90Y is given in Figure V.I.  A summary of fission yields for 90Sr 
is given in Table V.1. 

FIGURE V

Table V.1  Fission yields of strontium-90 [C3]
-------------------------------------
                   Fission yield (%) 
Nuclide            Thermal     Fast
-------------------------------------
235U               5.84        5.21
239Pu              2.12        2.05
238U                           3.20
232Th                          7.66
Average for
nuclear tests a/        3.50
-------------------------------------
a/ From reference [H1].

153.  Large amounts of 90Sr were released in nuclear tests and 
dispersed throughout the world.  Strontium-90 is also produced in 
the nuclear fuel cycle, but only small amounts are released to the 
environment.  Strontium-90 in the environment is efficiently 
transferred to human diet.  The absorption of 90Sr by the body is 
relatively high and it has a long biological retention time. 

154.  Because of the correspondence in behaviour of strontium and 
calcium in the environment and in man, it has been the practice to 
express measurement results in diet and bone as quotients of 90Sr 
to Ca concentrations.  Discrimination is reflected as ratios of 
strontium to calcium quotients in samples to those in precursor 
samples in the transfer chain. Expressing results in terms of the 
strontium to calcium quotients has the practical advantage that for 
many environmental transfer processes, such as absorption into the 
body, secretion into milk and deposition in bone, the ratios remain 
relatively constant and predictable.  However, since the average 
levels of calcium in diet and man are nearly constant, assessments 
of 90Sr can also be made on the basis of amounts per unit mass or 
volume of material. 

B.  SOURCES

1.  Nuclear explosions

155.  Strontium-90 is produced in nuclear explosions in the amount 
of approximately 3.7 1015 Bq per Mt of fission energy. Measurements 
of the fission debris from large nuclear tests gave a 90Sr fission 
yield estimate of 3.5% [H1].  Assuming 1 kt of fission energy 
corresponds to 1.45 1023 fissions [H1] and using the current best 
estimate of the 90Sr half-life (29.12 ± 0.24 a) [N3], the 90Sr 
production yield is estimated to be 3.8 ± 0.1 1015 Bq per Mt 
fission energy.  Large deviations are possible for individual 
tests. 

156.  Approximate fission energy yields of nuclear weapons tests 
conducted in the atmosphere have been published.  The total for 
tests through 1962 was 194 Mt of fission energy [F1, U3].  This 
would correspond to a production of about 74 1016 Bq of 90Sr.  An 
additional 9 1016 Bq have been produced in atmospheric tests to the 
end of 1980.  A portion of the total amount of 90Sr was local 
fallout, deposited in the immediate vicinity of testing regions.  

Local fallout is important especially for lower yield tests 
detonated on the land or water surface.  The best estimates of the 
activity of globally distributed 90Sr come from measurements of 
90Sr deposition.  This amounts to 6 1017 Bq for all tests 
conducted through 1980. 

2.  Nuclear fuel cycle

(a)   Nuclear reactors

157.  Strontium-90 is produced by fission in the fuel of nuclear 
reactors.  The amounts produced vary depending on the fuel 
composition, reactor type, and degree of fuel burn-up achieved.  
The yields for various fission processes were given in Table V.1.  
In fairly high burn-up fuel (33000 MW(t)d t-1) of a PWR, the 90Sr 
production is estimated to be 2.83 1015 Bq per tonne of fuel, 
corresponding to 9.5 1013 Bq per MW(e)a of electricity 
generated [O1]. 

158.  Small amounts of 90Sr produced in the fuel in nuclear 
reactors may reach the coolant through defects in the fuel 
cladding.  In coolant purification or following coolant leakage, 
90Sr may reach the gaseous and liquid effluent streams.  In 
controlled amounts, some of the effluents are released to the 
environment. 

159.  The activity releases of 90Sr were listed in the 1977 report 
of UNSCEAR [U6].  Average discharges were of the order of 0.01 to 4 
106 Bq per MW(e)a in PWRs and BWRs, respectively, with most of the 
release in liquid effluents.  Somewhat larger releases of 90Sr in 
liquid effluents from GCRs arise primarily from spent fuel storage 
pools.  There are considerable variations in the release amounts 
from individual reactors per unit electricity generated. 

160.  Assuming for each reactor type that the limited data of the 
90Sr activity released per unit of electrical energy generated are 
representative, it is possible to obtain a very rough estimate of 
the total amount of 90Sr released from reactors worldwide.  Using 
the installed capacities of the various reactor types in 1980 [I1] 
and assuming a reactor utilization of 60%, the estimated annual 
release from all reactors is about 2 1012 Bq (Table V.2).  In this 
calculation, it is assumed for the reactor types for which no data 
are available, that the releases are similar to those from BWRs. 

Table V.2   Estimated global discharges of strontium-90
from nuclear power stations in 1980
------------------------------------------------------------
Reactor   Reactor   Capacity   Production    Estimated
type      number    [MW(e)a]   rate [Bq      strontium-90
                               per MW(e)a]   discharge (Bq)
------------------------------------------------------------
PWR       96        64239      9 104         0.3 1010
BWR       62        35170      4 106         8 1010
GCR       36        7086       4 108         170 1010
Other     47        18490      4 106         4 1010
------------------------------------------------------------
Total     241       124985                   2 1012
------------------------------------------------------------

161.  There have not been many reports on 90Sr measurements in the 
environment surrounding nuclear reactors that can be attributed to 
reactor operation.  In fact, 90Sr would not be a likely 
radionuclide to be investigated, and furthermore results of 
measurements could hardly be made independent of fallout 90Sr.  The 
much more readily detectable 137Cs, for example, is released in 
activity amounts up to several hundred times that of 90Sr [U6].  
Caesium-134 is released from reactors in amounts about 60% less 
than 137Cs and, in addition, it would not be confused with weapons 
fallout in the environment.  For these reasons, estimates of 90Sr 
levels in the environment from reactor release must generally be 
based on measured releases and environmental dispersion 
calculations. 

(b)    Fuel reprocessing plants

162.  In fuel reprocessing plants the fuel is dissolved to recover 
uranium and plutonium for reuse.  All the 90Sr and other fission 
products as well go to waste streams.  The radionuclide activities 
in airborne and liquid effluents from fuel reprocessing plants have 
been recorded by UNSCEAR [U6]. The data for 90Sr are summarized in 
Table V.3. 

Table V.3  Average normalized discharges of strontium-80
to the environment from fuel reprocessing plants
(Bq per MW(e)a) [U6]
--------------------------------------------------------------
Plant                            Airborne    Liquid    Total
                                 effluents   effluents
--------------------------------------------------------------
Windscale (United Kingdom)       5 106       2 1011    2 1011

Nuclear Fuel Services (U.S.A.)   2 106       5 108     5 108
--------------------------------------------------------------

163.  Strontium-90 is released from fuel reprocessing plants 
primarily in liquid effluents.  Relative to the total amounts of 
90Sr in spent fuel, the release amounts are not large.  The 
fractional releases in liquid effluents are approximately 2 10-3 
from the Windscale plant in the United Kingdom and were about 
5 10-6 from the small Nuclear Fuel Services plant in the U.S.A., 
which is no longer in operation. 

164.  The fractional release of 90Sr in airborne effluents is about 
3 10-8, compared to the amounts present in spent fuel. The 
experiences at Windscale and at Nuclear Fuel Services are similar 
in this regard. 

165.  The releases of radioactivity into the Irish Sea from the 
Windscale plant and the corresponding levels in the environment 
have been closely studied [N1].  Estimates of inventories of 
activity in the sea and sediments and of movements of the activity 
in its passage into the North Sea are being made [M1].  Some 
results of environmental surveys around the reprocessing plant at 
La Hague have also been published [S1].  Most of the attention, 
however, is focused on the greater quantities of other fission 
products released. 

C.  BEHAVIOUR IN THE ENVIRONMENT

1.  Movement in soil

166.  The downward penentration of 90Sr in soil is slow, although 
it is more rapid than for 137Cs or 249Pu.  Even after several years 
90Sr remains in the upper few centimetres in undisturbed soil.  The 
rate of movement varies with soil type; a low content of clay and 
humus, a high content of electrolytes and a rapid movement of water 
increase penetration [U2].  The mechanism of movement is thought to 
involve both leaching and diffusion. 

2.  Transfer to plants

167.  Plants acquire 90Sr by direct deposition onto foliage and by 
root uptake of 90Sr in the soil.  Absorption into the leaves is 
relatively slow and superficial material is readily lost by 
weathering.  The translocation of 90Sr from plant leaf or grain 
surfaces to other parts of the plant is small. 

168.  Capture of 90Sr on inflorescences of plants is of importance 
for entry to grain.  Concentrations in husked grain will be higher 
than in the milled product during periods of deposition. 

169.  Uptake from soil is normally the primary mode of 90Sr entry 
into plants.  The quantity of absorbable calcium in soil is an 
important factor in determining the extent of 90Sr absorption by 
plants.  Uptake is greatest from soils of low calcium content.  
Uptake is thus reduced by the addition of lime, but usually not by 
a factor exceeding 3 [U2].  When soils contain adequate calcium for 
growth and the exchange capacity is largely saturated with calcium, 
the addition of lime has little or no effect. 

170.  Other factors which affect root uptake of 90Sr include the 
clay and humus content of the soil, pH, the concentration of 
electrolytes other than calcium and the moisture content. The 
addition of organic matter and fertilizers to soil may have varying 
effects on plant uptake, which are, however, usually not large when 
these materials are applied at normal agricultural levels. 

171.  Plant-base absorption of 90Sr has been noted to be quite 
efficient.  The 90Sr trapped in the surface root mat is relatively 
undiluted with the calcium in the soil and is in a particularly 
favourable position for absorption.  This would help to explain the 
higher concentrations of 90Sr in grain in periods shortly after 
deposition and would be an important process in permanent pastures. 

3.  Transfer to milk

172.  The total quantity of ingested 90Sr secreted into the milk of 
cows is variable, depending on the milk yield.  Values range from 
0.5 to 2% of a single oral administration [U2]. With continuous 
ingestion under normal conditions of feeding, several 
investigations have shown that about 0.08% of the amount given 
daily is secreted per litre of milk.  The transfer to goats milk 
may be more than ten times greater, corresponding to a higher 
proportion of dietary calcium secreted into the milk as well. 

4.  Transfer to diet

173.  The transfer of fallout 90Sr to diet has been extensively 
studied.  Assessments by UNSCEAR have been based on application of 
generalized transfer models [U6].  The transfer from deposition to 
diet is quantitatively described by means of the transfer factor 
P23 defined as the time-integrated 90Sr/Ca quotient in the diet 
divided by the 90Sr integrated deposition density.  The integrals 
may be replaced by summations if the relevant quantities are 
assessed over discrete intervals of time.  In fact, annual values 
are the most generally available information.  The transfer factor 
P23 is usually expressed in mBq a/gCa per Bq m-2. 

174.  The transfer to diet from a specific deposition occurs over 
an extended period as long as 90Sr remains in soil available for 
root uptake.  The model used by UNSCEAR to describe the transfer of 
90Sr from deposition to diet is 

                                    infinite       -µm
    C(n) = b1 f(n) + b2 f(n-1) + b3 sigma   f(n-m)e
                                    m=1

where C(n) is the 90Sr/Ca quotient in total diet, in a food group, 
or in an individual food item, in the year n;  f(n) is the annual 
deposition density in the year n, and b1, b2, b3, and µ are factors 
which can be derived from reported data by regression analysis [U5, 
U6].  The first term in the equation represents the contribution to 
dietary 90Sr per unit deposition density in the current year, while 
the second term expresses separately the contribution from 
deposition in the previous year, reflecting also the use in the 
current year of stored food produced in the previous year.  The 
third term expresses the contribution to dietary 90Sr from the 
deposition density in all previous years, resulting from root 
uptake and taking into account decay and loss of availability due 
to downward movement in soil or to other physical or chemical 
changes which may occur.  The inverse of µ is the mean life of 
available 90Sr in soil, which varies for individual foods and soil 
conditions. 

175.  The transfer factor P23 describes the cumulative transfer of 
90Sr to diet per unit deposition density.  Using the model 
described above, the expression of the transfer factor is 

                       infinite  -µm                e
    P23 = b1 + b2 + b3 sigma    e    = b1 + b2 + b3 -----
                       m=1                          1-e

The transfer factor P23 can therefore be estimated from the 
parameters b1, b2, b3 and µ obtained by regression analysis from 
reported data. 

176.  The values of the parameters obtained from the regression 
fits to 90Sr deposition density and diet data from the fallout 
measurement programmes in New York and Denmark for dietary 
components and for the total diet have been reported by UNSCEAR 

[U6].  For total diet, the values of the parameters (b1 ~ 1.0, 
b2 ~ 0.9, b3 ~ 0.3, µ ~ 0.1) give estimates of the initial 
transfer (b1 + b2) of 1.9 mBq a (gCa)-1 per Bq m-2 and of the 
long-term transfer (last term of equation above) of 2.9 mBq a 
(gCa)-1 per Bq m-2.  The total transfer, the value of P23, is 
4.8 mBq a (gCa)-1 per Bq m-2. 

177.  Similar data from Argentina have also been evaluated, with 
reasonable agreement found for all three areas for individual foods 
and for total diet.  Some differences are noted which are due to 
the different definition of the food groups or to different soil 
conditions and agricultural practices in the three countries. 

178.  The long-term transfer of 90Sr to diet from a single input to 
the environment can be illustrated from the values of the transfer 
parameters obtained from component groups of diet.  For example, it 
is determined that 90% of the cumulative transfer from a single 
release of 90Sr is completed within 9 years for meat, fish and 
eggs, 12 years for grain products, 14 years for milk, 32 years for 
vegetables and 77 years for fruit.  More rapid transfer indicates 
that direct deposition processes are more important.  Slow transfer 
represents primarily uptake from the slowly decaying deposit of 
90Sr in soil.  Since the exponential decreases of the transfer for 
the various groups differ, a direct fit to total diet data with a 
single exponential transfer function is expected to be less 
accurate than the summation of fits for the individual components. 

179.  The transfer of 90Sr from deposition density to some 
individual foods, particularly to milk, has been studied for a 
number of different areas of the world.  These results are listed 
in Table V.4.  The transfer factors range from 2.1 to 7.6 mBq a/gCa 
per Bq m-2. 

Table V.4  Parameters of the transfer function between deposition density and 90 Sr/Ca in milk
-------------------------------------------------------------------------------------------------
Parameter  Northern    San        New    United   Denmark  Argentina  Norway  Australia  Faroe
a/         hemisphere  Francisco  York   Kingdom                                         Islands
-------------------------------------------------------------------------------------------------
b1         0.84        0.61       0.69   0.89     0.99     1.39       0.70    2.07       2.70
b2         0.54        0.61       0.23   0.47     0.46     1.24       0.44    1.27       1.38
b3         0.22        0.19       0.19   0.15     0.23     0.12       1.02    0.30       0.81
µ          0.12        0.19       0.13   0.13     0.13     0.12       0.33    0.08       0.21
P23        3.1         2.1        2.3    2.4      3.1      3.5        3.7     7.2        7.6
-------------------------------------------------------------------------------------------------
a/  The unit for parameters b1, b2, b3 is mBq a(gCa)-1 per Bq m-2
    The unit for parameter µ is a-1
    The unit for the transfer factor P23 is mBq a(gCa)-1 per Bq m-2.

180.  Milk has been used as an indicator of the levels of 90Sr in 
total diet in areas where foods other than milk were not analysed.  
This must be done, however, with some caution.  It is usually the 
case that following a period of 90Sr deposition, the 90Sr/Ca 
quotient in milk declines somewhat more rapidly than the 90Sr/Ca 
quotient in total diet.  Where data are available, it is seen that 

the exponential factor is greater in milk than in total diet.  This 
means that the diet-milk ratio of 90Sr/Ca values changes with time 
during and after the period in which the 90Sr deposition occurs. 

181.  The diet-milk ratios at particular times also show 
considerable variation from one country to another [U6].  In most 
countries where milk is an important component of the diet, the 
diet-milk ratio has averaged about 1.4 for most of the fallout 
years.  A trend to increasing ratios in the future is expected if 
the deposition density rate remains at a very low level.  In 
countries where milk is not an important component of diet, the 
diet-milk ratio will have a much higher value. 

5.  Aquatic behaviour

182.  Strontium, like calcium, appears mainly in ionic form in 
water and is not strongly sorbed by suspended particulate 
materials.  The fraction of strontium found in the particulate 
phase in several freshwater systems ranged from 1 to 10% [V1]. 

183.  It has been of interest to determine the behaviour of 90Sr in 
aquatic environments in the vicinity of nuclear installations.  A 
number of determinations of the concentration factors (ratios of 
integrated concentrations or of equilibrium concentrations in 
organisms and in the water) have been performed in recent years for 
various marine biota by measurements of stable and radioactive 
strontium under laboratory and field conditions [C1, F2, N2, U4, 
U1].  Typical values are 100 for algae, 2 to 10 for crabs and 
lobster, about 1 for the flesh of ocean fish and 5 for fresh water 
fish. 

184.  The primary uptake of strontium and also calcium by fish 
occurs directly from the water.  Therefore, accumulations in 
organisms are little dependent on trophic level [V1].  The 
concentration factor for fish depends inversely on the 
concentration in the water.  Vanderploeg et al. [V1] have suggested 
a quantitative relationship.  The concentration factors for fish 
bone are about two orders of magnitude greater than for flesh. 

185.  The UNSCEAR aquatic model can be used to estimate the 
transfer of 90Sr from the aquatic environment to diet for 
generalized discharge situations [U6].  Further discussion of this 
is presented in the section on dosimetry. 

D.  TRANSFER TO MAN

186.  Strontium-90 is acquired by man primarily through ingestion 
of 90Sr contaminated food.  Terrestrial pathways are generally more 
important than aquatic pathways in transferring 90Sr to man.  From 
fallout experience it is noted that 90Sr in drinking water always 
contributes less than 5% of the total ingestion intake and 90Sr in 
fish is a minor contributor even in countries where consumption of 
fish is high.  For example, it is estimated that only about 3% of 
the fallout 90Sr intake by man in Japan between 1966 and 197l came 
from fish [U1]. 

187.  Correlations of fallout 90Sr in diet with measurements of 
90Sr in bone have provided a relationship which is used to evaluate 

the transfer to man.  The transfer function used by UNSCEAR [U5, 
U6] is: 
                        infinite        -µm
    Cb(n) = c Cd(n) + g sigma   Cd(n-m)e
                        m=o

where Cb(n) is the 90Sr/Ca quotient in bone in the year n, Cd(n) is 
the dietary 90Sr/Ca quotient in the year n, and c, g, and µ are 
constants determined from regression analysis of the bone and diet 
data.  The parameter can be associated with that portion of 90Sr 
intake which is retained for a short term on bone surfaces and is 
readily exchanged with plasma, and the parameter g with that 
portion more tightly retained in bone. The exponential term 
describes the effective removal rate of 90Sr from bone due to 
radioactive decay and bone remodelling. 

188.  The transfer factor P34 linking diet and bone, is defined as 
the ratio of the time integrated 90Sr/Ca quotient in bone to that 
in diet.  It may also be thought of as the cumulative transfer and 
retention 90Sr in bone per unit intake in diet.  The expression for 
the transfer factor using the above function is 

                infinite -µm         g
    P34 = c + g sigma   e    = c + ------
                m=o                1-e

Data are available for regression analysis applying this model only 
for 90Sr in vertebrae.  It is recognized that there is greater 
initial retention in cancellous bone such as vertebrae than in 
compact bone, but there is also more rapid turnover. Therefore, the 
time integrated results are expected to be representative of the 
skeleton as a whole. 

189.  The results of regression fits of fallout 90Sr diet and adult 
vertebrae data for various localities have been given in UNSCEAR 
reports [U5, U6].  Representative values of the parameters are 0.02 
Bq a (gCa)-1 in bone per Bq a (gCa)-1 in diet for both c and g and 
about 0.2 a-1 for µ.  The values determined for P34 range from 0.12 
to 0.16 Bq a (gCa)-1 in bone per Bq a (gCa)-1 in diet. 

190.  From the values of the exponential parameter µ it is possible 
to infer the  mean residence time of 90Sr in bone. This ranges from 
3.4 to 6.7 years, corresponding to bone turnover rates of 12 to 23% 
per year.  There is no reason to expect that the metabolic 
behaviour should differ in the various areas of the world, and as 
far as is known, dietary composition of usual foods does not affect 
90Sr availability. Therefore, the differences in the estimate are 
attributed to variations in the survey measurements. 

191.  To account for the bone 90Sr/Ca quotients in children, 
Bennett [B2] has used age-dependent parameters in the above 
transfer function.  For children under 9 years the parameter c was 
found to be zero, consistent with a single exponential transfer 
model as originally proposed by Rivera [R1].  The best fit was 
obtained with a turnover rate of 90Sr varying from about 100% per 
year down to about 40% per year in the pre-teenage years and then 
falling with age to about 20% per year for adults.  Beninson [B1] 
had reported a similar variation in turnover rate with age in 
Argentina.  The fractional retention of strontium, the fraction of 
dietary intake incorporated into the skeleton, was also found by 
Bennett [B2] to vary with age being five to seven times higher for 
infants than for adults.  Additional considerations of 90Sr 
metabolism as a function of age have been presented by Papworth and 
Vennart [P1]. 

192.  The initial 90Sr concentration in the newborn must be 
determined from an empirical relationship with the mother's diet.  
The 90Sr/Ca quotient in bone of newborn varies from 0.1 to 0.2 
times the 90Sr/Ca quotient in diet of the mother during the year 
prior to the birth.  An average of about 0.15 is obtained from the 
survey data [B1, B2]. 

E.  DOSIMETRY

1.  Dose per unit intake

193.  It is useful for radiological assessments to have expressions 
for the dose per unit ingestion intake and per unit inhaled amount.  
The absorbed doses in bone marrow and in bone-lining cells per unit 
integrated activity of 90Sr in bone have been evaluated based on 
the work of Spiers [S2].  These are the transfer factors P45 
relating activity in bone to the doses [U5, U6].  The values are 
P45 (bone marrow) = 0.38 mGy per Bq a (gCa)-1;  P45 (bone-lining 
cells) = 0.53 mGy per Bq a (gCa)-1.  If it may be assumed that the 
dietary calcium intake rate is 1 g daily (365 g/a), the transfer 
factor to bone per unit intake, P34, is 0.14/365 Bq a (gCa)-1 per 
Bq.  This multiplied by the factor P45 gives the absorbed doses in 
bone marrow and bone-lining cells.  The results are given in Table V.5. 

Table V.5  Absorbed dose per unit intake of 
strontium-90 (Gy/Bq)
--------------------------------------------------
             Lung       Bone marrow   Bone-lining
                                      cells
--------------------------------------------------
Ingestion    -          1.5 10-7      2.0 10-7

Inhalation   5.8 10-9   4.9 10-7      6.9 10-7
--------------------------------------------------

194.  The ICRP Task Group lung model gives guidance regarding the 
disposition of inhaled radioactivity [I3, I4] in the respiratory 
tract.  The respiratory system is divided into the nasopharyngeal 
region (NP), the tracheobronchial region (TB), and the pulmonary 
region (P).  For the present dosimetric assessments it will be 

assumed that the 90Sr is associated with typical ambient aerosols 
of average diameter 0.5 µm.  The 90Sr compounds are grouped in 
Class D, retention in the lung being in the order of days, 
specifically 0.5 d for the portion deposited in the pulmonary 
region. 

195.  Fractional deposition of 0.5 µm particles in the NP, TB, and 
P regions are 0.14, 0.08, and 0.30 and subsequent fractional 
transfers to blood for the Class D compounds are 0.5, 0.95, and 
1.0, respectively.  Total transfer to blood is thus 0.446 of the 
inhaled amount.  Fractional transfer from blood to bone is 0.3 
[I2]. 

196.  The lung dose as a function of the inhaled activity can be 
calculated from the expression 
                _
            k A E f  1.44 TB
    Dlung = ----------------
                 M
                                                             _
where k is a dosimetric constant, A is the activity inhaled, E is 
the average energy per disintegration, f is the fraction of the 
activity retained in the lung, TB is the retention half-time in the 
lung, and M is the mass of the lungs.  The lung dose per unit 
activity of 90Sr inhaled is therefore: 

                      Gy/d        MeV      1.44 0.5d
    Dlung = 13.8 10-6 ------ 1.13 --- 0.52 ---------
                      Bq MeV      dis      1000 g
                      g  dis

          =  5.8 10-9 Gy/Bq

197.  Assuming that the mean residence time of 90Sr in bone is 10 
years, applicable to the skeleton as a whole, the integrated 
concentration of 90Sr in bone per unit inhaled amount is 0.446 x 
0.3 x 10 years ‰ 100 gCa = 1.3 10-3 Bq a/gCa per Bq inhaled.  The
absorbed doses in bone marrow and bone lining cells are obtained by 
multiplying the values of the transfer factor P45 given above.  The 
results are listed in Table V.5. 

2.  Dose per unit release

(a)   Nuclear explosions

198.  The dose commitment from 90Sr released by nuclear explosions 
can be assessed using the environmental compartment model outlined 
in the introduction.  The transfer factor of a sequence of steps in 
series is the product of the transfer factors of each step.  The 
dose commitment, Dc, is related to the integrated deposition 
density of 90Sr, F, by the following expression: 

          Dc = P23 P34 P45F

where P23, P34 and P45 are the transfer factors discussed 
previously.  Average values for these transfer factors, as assessed 
in the 1977 UNSCEAR report [U6], are P23 = 5 mBq a/gCa per Bq m-2, 
P34 = 0.14, and P45 as given previously. 

199.  The dose commitment from 90Sr ingestion per unit of 
widespread deposition density of 90Sr such as from nuclear 
explosions is thus 0.3 µGy per Bq m-2 for bone marrow and 0.4 µGy 
per Bq m-2 for bone lining cells. 

200.  The dose commitment from 90Sr via the inhalation pathway can 
be estimated from the time integrated concentration of 90Sr in air.  
Multiplying the dose commitment per unit inhalation intake (Table 
V.5) by the inhalation intake rate of 22 m3 d-1 gives the value of 
1.3 10-7 Gy per Bq d m-3 for the dose commitment to lungs per unit 
integrated concentration of 90Sr in air. 

201.  The dose commitment to lungs can also be referred to measured 
values of the integrated deposition density. Dividing the 
integrated deposition density (Bq m-2 by the average deposition 
velocity (m s-1) gives the time integrated concentration in air.  
From long-term measurements of fallout 90Sr, the average deposition 
velocity is 2.2 cm/s in New York, 2.0 cm/s in Argentina, and 1.5 
cm/s in Denmark [A1, C2, E1]. The values determined from annual 
measurements at all three sites range from 1.2 to 2.9 cm/s.  The 
value of 2 cm s-1 can be taken as representative.  One Bq m-2 
integrated deposition density thus corresponds to 5.8 10-4 Bq d 
m-3) in air.  The dose commitment to lung per unit integrated 
deposition density of 90Sr is thus 7.5 10-11 Gy per Bq m-2. 

202.  The total amount of 90Sr released to the environment by 
nuclear tests, 6 1017 Bq, has given a population-weighted 
integrated deposition density of 1940 Bq m-2 in the world as a 
whole [U6].  The world population is 4 109.  With these values the 
collective dose commitments per unit activity of 90Sr released may 
be estimated.  The results, which are summarized later, apply to 
the geographic pattern of past nuclear tests. 

(b)   Nuclear installations

203.  The activity of 90Sr in airborne effluents is dispersed by 
turbulent air movement and is eventually deposited on the ground.  
It then enters the ingestion pathway.  The average situation over 
several years of routine discharges will result in a nearly 
completed deposition within a region with a radius R.  The average 
integrated deposition density, for a discharged activity A is F = 
A/pi R2.  The number of individuals exposed to that mean integrated 
deposition density, on the other hand, is N = deltaN pi R2, 
provided the population density deltaN can be assumed to be 
constant over the relevant area. 

204.  It follows that the collective dose commitment per unit 
activity released, Sc1, can be assessed by the expression 

    Sc = P23 P34 P45 deltaN
     1
Using the values for the transfer factors given previously, and 
assuming a population density of 25 man km-2, the collective dose 
commitments per unit activity released are estimated to be about 7 
10-12 man Gy per Bq for bone marrow and about 9 10-12 man Gy per Bq 
for bone-lining cells.  These values apply provided that food is 
locally produced and that the production suffices for the 

population density under consideration.  The estimates would 
probably be lower in actual circumstances. 

205.  The contribution of inhalation to the collective dose 
commitment, for effluent releases over many years, can be estimated 
by integration of the functions describing the atmospheric 
dispersion, assuming complete depletion by deposition within 100 
km.  The collective dose commitment contribution per unit activity 
released would be given by the expression 

     c    X                    100 km   r    -1.5
    S1 = (Q)1 km I deltan phi /       (1 km)      2 pi r dr
                               1 km   

       X
where (Q)1 km is the dispersion factor at 1 km from the release 
point, I is the individual intake rate of air, deltaN is the 
population density, phi is the dose per unit activity inhaled of 
90Sr and r is the distance from the release point [U6]. 

206.  The collective dose commitment may also be estimated using 
the deposition velocity, thus avoiding the need to specify the 
deposition area.  That is, the integrated deposition density from a 
discharged activity A is A/pi R2. Dividing by the deposition 
velocity gives the integrated concentration of 90Sr in air.  
Multiplying by the population, deltaN pi R2, removes the areal 
dependence. Using a population density of 25 man km-2, an air 
intake rate of 22 m3 d-1, a deposition velocity of 0.5 cm s-1 
appropriate for particulates from near surface releases, and the 
dosimetric factors given in Table V.5, estimates of the collective 
r dose commitments per unit release are obtained. The results are 
listed in the summary Table V.6. 

207.  The collective dose commitment from the input of 90Sr in 
water bodies, normalized per unit activity released, can be 
estimated [U5] using the expression 

     c   sigmak  Nk  Ik  fk  phi
    S1 = ------------------------
           V(lambda + 1/tau)

Here V is the volume of receiving waters, tau is the turnover time 
of receiving waters, lambda is the decay constant of 90Sr, Nk is 
the number of individuals exposed by pathway k, Ik is the 
individual consumption rate of pathway item k, fk is the 
concentration factor for the consumed item in pathway k, and phi is 
the collective dose per unit activity ingested collectively by the 
exposed group. 

                           1       
208. The quantity V(lambda + 1/tau) is the infinite-time integral 
of the water concentration per unit of activity released, while the 
quantity multiplied by fk is the infinite-time integral of the 
concentration in the consumed item k.  For radionuclide inputs into 
small volumes of water, the concentration in water and in fish will 

be high, but the population which can be served with drinking water 
or by fish consumption will be limited.  For inputs into larger 
volume of water the concentrations will be smaller, but the 
population involved will be correspondingly larger.  It seems 
reasonable to assume, as a first approximation, that the quantities 
Nk/V are relatively constant and independent of V. 

209.  A summary of the values used in the assessments presented by 
UNSCEAR [U6] is given in the following listing: 

    Parameter                    fresh water   sea water
1.  tau, turnover time of        10 a          1 a
    receiving water
2.  Correction factor for        0.78          1.0
    sediment removal
3.  V, water utilization factor  3 107 1/man   3 109 1/man
    N
4.  fk, concentration factor
    for item k
        drinking water           0.5
        fish                     5             1
        shellfish                              1
5.  Ik, consumption rate for
    item k
        drinking water           440 1/a
        fish                     1 kg/a        6 kg/a
        shellfish                              1 kg/a

Using these values and the dosimetric factors given in Table V.5, 
the collective dose commitments per unit activity released in 
liquid effluents are estimated to be about 9 10-12 man Gy/Bq for 
bone marrow and 1 10-11 man Gy/Bq for bone lining cells for 
discharges into fresh waters, and a rounded value of 4 10-16 man 
Gy/Bq for both tissues for discharges into the sea.  The values for 
sea discharges underestimate the collective dose commitment because 
they neglect the contribution from large-scale mixing with a longer 
residence time.  However, due to the relatively short life of 90Sr, 
this contribution can only have a small effect on the estimates. 

3.  Summary

210.  Table V.6 summarizes the values obtained above for the 
collective dose commitments, normalized per unit activity released, 
for releases by atmospheric nuclear explosions and in effluents of 
the nuclear power industry.  In both cases they are the result of 
generalized assessments and substantial variations should be 
expected in site-specific cases. 

211.  For 90Sr and 90Y radiations, the quality factor is one, and 
the weighting factors are 0.12 for lung and for bone marrow and 
0.03 for bone lining cells.  The effective dose equivalent 
commitment per unit intake of 90Sr is thus 2.4 10-8 Sv/Bq 
(ingestion) and 8.0 10-8 Sv/Bq (inhalation).  The collective 
effective dose equivalent commitments per unit release from nuclear 
installations and from nuclear tests are included in Table V.6. 

Table V.6  Summary of collective dose commitments per unit
strontium-90 activity released (10-14 man Gy per Bq)
----------------------------------------------------------------
                          Lung   Bone    Bone lining  Effective
                                 marrow  cells        a/
----------------------------------------------------------------
 Nuclear explosions
  Ingestion                      400     500          60
  Inhalation              0.1    8       10           1

 Nuclear installations
  Release to air b/
    Ingestion                    700     900          100
    Inhalation            0.7    60      90           10
  Release to fresh water
    Drinking water               900     1000         100
    Fish                         20      30           3
  Release to salt water
    Fish                         0.03    0.04         0.005
    Shellfish                    0.005   0.007        0.0008
----------------------------------------------------------------
a/  Collective effective dose equivalent commitment 
    (10-14 man Sv/Bq).
b/  Assumes population density of 25 man/km2 which is fully 
    sustained by local food production.

F.  REFERENCES

A.  Aarkrog, A. and J. Lippert.  Environmental radioactivity in 
    Denmark in 1971, 1972, 1973, 1974 and 1975.  Danish Atomic 
    Energy Commission.  Riso reports 265 (1972), 291 (1973), 305 
    (1974), 323 (1975) and 363 (1976). 

B1  Beninson, D., A. Migliori de Beninson, C. Menossi et al. 
    Radioestroncio en el hombre en funcion de la edad. Trabajo 
    presentado en el Quinto Congreso International de la "Société 
    francaise de radioprotection".  Grenoble, 1971. 

B2  Bennett, B.G. Strontium-90 in human bone - 1976 results for New 
    York and  San Francisco. p. I-69-I-84  in Health and Safety 
    Laboratory environmental quarterly report HASL-328.  New 
    York,1977 

C1  Cancio, D., J.A. Llauró, N.R. Ciallella et al. Incorporación de 
    radioescio por organismos marinos. p. 347-356  in Radioactive 
    Contamination of the Marine Environment. Proceedings of a 
    symposium, Seattle, 1972. IAEA publication STI/PUB/313.  
    Vienna, 1973. 

C2  Comisión Nacional de Energía Atomica.  Argentina. Information 
    submitted 1977. 

C3  Crouch, E.A.C.  Fission product yields from neutron induced 
    fission.  Atomic Data and Nuclear Data Tables, Vol. 19, No. 5.  
    Academic Press, New York, 1977. 

E1  Environmental Measurements Laboratory, U.S. Department of 
    Energy.  Environmental quarterly and appendix. Environmental 
    Measurements Laboratory report EML-334. New York, 1978. 

F1  Federal Radiation Council.  Estimates and evaluation of fallout 
    in the United States from nuclear weapons testing conducted 
    through 1962 - Report No. 4 (1963). 

F2  Foyn, L.  Some marine radioecological problems at the nuclear 
    power station establishment at Oslofjorden. Fisheries 
    Directorate Sea Research Institute report series B, No. 10.  
    Bergen, Norway, 1973. 

H1  Hallden, N.A., I.M. Fisenne, D.Y. Ong et al.  Radioactive decay 
    of weapons debris p. 194-199  in Health and Safety Laboratory 
    fallout program quarterly summary report HASL-117.  New York, 
    1961. 

I1  International Atomic Energy Agency.  Power reactors in member 
    states. IAEA, Vienna, 1980.12 

I2  International Radiological Protection.  Report of Committee II 
    on Permissible Dose for Internal Radiation. ICRP publication 2, 
    Pergamon Press, 1959. 

I3  International Commission on Radiological Protection.  Task 
    group on lung dynamics.  Deposition and retention models for 
    internal dosimetry of the human respiratory tract. Health Phys. 
    12: 173-226 (1966). 

I4  International Commission on Radiological Protection.  Task 
    group of Committee 2.  The metabolism of compounds of plutonium 
    and other  actinides.  ICRP publication 19, Pergamon Press, 
    1972. 

M1  Mitchell, N.T.  Radioactivity in surface and coastal waters of 
    the British Isles, 1972-1973.  U.K. Ministry of Agriculture  
    Fisheries and Food.  Fisheries Radiobiological Laboratory 
    report FRL-7, FRL-8, FRL-9, FRL-10 (1971, 1973, 1975). 

N1  National Radiological Protection Board.  The data submitted by 
    the United Kingdom to the United Nations Scientific Committee 
    on the Effects of Atomic Radiation for the 1977 Report to the 
    General Assembly.  NRPB report R47, Harwell (1976). 

N2  Norwegian Institute for Water Research.  Release of radioactive 
    materials from nuclear power stations.  Report No. 2.  
    Dispersal mechanisms, pathways and concentration factors for 
    radionuclides in the cooling waters.  Report 0-177/70  (1974). 

N3  National Council on Radiation Protection and Measurements.  A 
    handbook of radioactivity measurements procedures.  NCRP report 
    No. 58, Washington D.C. (1978). 

O1  Oak Ridge National Laboratory.  Siting of fuel reprocessing 
    plants and waste management facilities.  Oak Ridge National 
    Laboratory report ORNL-4451 (1970). 

P1  Papworth, D.G. and J. Vennart.  Retention of 90-Sr in human 
    bone at different ages and the resulting radiation doses.  
    Phys. Med. Biol. 18:  169-186 (1973). 

R1  Rivera, J. and J.H. Harley.  The HASL bone program 1961-1964.  
    U.S. Atomic Energy Commission report HASL-163.  New York, 1965. 

S1  Scheidhauer, J., R. Ausset, J. Planet et al.  Programme de 
    surveillance de l'environment marin du centre de La Hague. 
    p. 347-365  in Population Dose Evaluation and Standards for 
    Man and His Environment.  IAEA publication STI/PUB/375. Vienna, 
    1974. 

S2  Spiers, F.W., G.D. Zanelli, P.J. Darley et al. Beta-particle 
    dose rates in human and animal bone. p. 130-148  in Biomedical 
    Implications of Radiostrontium Exposure.  U.S. Atomic Energy 
    Commission Symposium Series 25 (1972). 

U1  Ueda, T., Y. Suzuki and R. Nakamuru.  Transfer of caesium-137 
    and strontium-90 from the environment to the Japanese 
    population via the marine environment  in Population Dose 
    Evaluation and Standards for Man and His Environment. IAEA 
    publication STI/PUB/375.  Vienna, 1974. 

U2  United Nations.  Report of the United Nations Scientific 
    Committee on the Effects of Atomic Radiation.  Official Records 
    of the General Assembly, Seventeenth Session, Supplement No. 16 
    (A/5216).  New York, 1962. 

U3  United Nations.  Report of the United Nations Scientific 
    Committee on the  Effects of Atomic Radiation.  Official 
    Records of the General Assembly, Nineteenth Session, Supplement 
    No. 14 (A/5814).  New York, 1964. 

U4  United Nations.  Report of the United Nations Scientific 
    Committee on the effects of Atomic Radiation.  Official Records 
    of the General Assembly, Twenty-First Session, Supplement No. 
    14 (A/6314).  New York. 1966. 

U5  United Nations.  Ionizing Radiation:  Levels and Effects. 
    Report of the United Nations Scientific Committee on the 
    Effects of Atomic Radiation to the General Assembly, with 
    annexes.  United Nations sales publication, No. E.72.IX.17 and 
    18.  New York, 1977. 

U6  United Nations.  Sources and Effects of Ionizing Radiation.  
    United Nations Scientific Committee on the Effects of Atomic 
    Radiation 1977 report to the General Assembly, with annexes.  
    United Nations sales publication No. E.77.IX.I.  New York, 
    1977. 

V1  Vanderploeg, H.W., D.C. Porzyck, W.H. Wilcox et al. 
    Bioaccummulation factors for radionuclides in freshwater biota.  
    Oak Ridge National Laboratory report ORNL-5002 (1975). 

VI.  IODINE

A.  INTRODUCTION

212.  Iodine is a volatile element which is very mobile in the 
environment.  It is non-uniformly distributed in nature, its 
abundance in the lithosphere being about 5 times higher than in the 
ocean waters.  The iodine in the sea apparently originates from 
erosion of the land masses.  Recycling to the terrestrial biosphere 
occurs through evaporation of sea water and decomposition of 
substances of marine origin. 

213.  There are at least 25 iodine isotopes with mass numbers 
ranging from 117 to 141.  All except 127I are radioactive.  Omitting 
the very short-lived 140I and 141I, thirteen isotopes are produced 
by fission: 
127I (stable), 128I (25 min), 129I (1.57 107a), 130I (12.4 h),
131I (8.06 d), 132I (2.3 h), 133I (21 h), 134I (52.8 min),
135I (6.7 h), 136I (83 s), 137I (23 s), 138I (5.9 s) and 139I
(2 s).  From the point of view of environmental contamination and 
resulting doses to man, the most important isotopes of iodine are 
131I and 129I.  They are the only radioactive isotopes of iodine 
produced by fission with half-lives longer than one day.  Iodine-
131 is a beta-emitter with a half-life of 8.06 days and a maximum 
beta energy of 0.81 MeV emitting also gamma rays of 0.36 and 0.64 
MeV and other  energies. Iodine-129 has a very long half-life (1.57 
107 a);  it is a beta-emitter (maximum energy:  0.15 MeV) with an 
accompanying gamma ray of 0.09 MeV in 8% of the disintegrations 
[D1].  The two isotopes are mainly found in the environment as a 
result of nuclear explosions and releases from nuclear reactors and 
fuel reprocessing plants.  Only these two isotopes of iodine are 
considered in this report. 

214.  Iodine enters the metabolism of living organisms and is 
selectively taken up and concentrated in the thryoid gland; it 
plays a major role in the synthesis of the thyroid hormone and is 
secreted in milk.  Owing to its decay properties, 131I has been  
extensively used in the medical field for diagnosis and treatment 
of thyroid abnormalities. 

B.  SOURCES

1.  Natural production

215.  Like any other fission product, 129I and 131I are present in 
the environment as a result of spontaneous fission of natural 
uranium.  In view of its very long half-life, 129I has accumulated 
in the earth's crust and also in the ocean waters from where it is 
available to disperse in the whole biosphere. 

216.  In 1962, Edwards [E1] predicted the natural 129I/127I atom 
ratio in sea water to be about 3 10-14 from spontaneous fission of 
uranium-238 and estimated that contributions of the same order of 
magnitude would arise from spontaneous fission of 235U and from 
production in the upper atmosphere by interaction of energetic 

protons, neutrons and photons on isotopes of xenon.  Experiments 
later confirmed the validity of Edwards' estimate, the 129I/127I 
atom ratios derived from measurements in iodine-rich minerals 
ranging from 2 10-15 to 10-13 [M1, S1].  Taking the concentrations 
of 127I in sea water to be 0.064 µg g-1 and the global volume of 
sea water to be 6 1017 m3, a natural activity of 129I of 7 109 Bq 
in the oceans is obtained from an 129I/127I atom ratio of 3 10-14. 

2.  Nuclear explosions

217.  The activity of 131I (or of 129I) generated in nuclear 
explosions can be derived from the measurements of 90Sr deposition 
[U1] and from the ratios of the fission yields of 131I (or 129I) 
and of 90Sr for nuclear weapons tests [H2]. 

218.  The total activity of 90Sr produced in nuclear explosions 
through 1976, which has been globally distributed, is estimated to 
be 6 1017 Bq [U1].  This does not include the local fallout, which 
is deposited in the vicinity of the test area and which can be 
important for the lower yield tests detonated near or on the land 
surface. 

219.  On the basis of measured fission product yields of individual 
nuclides obtained by analysis of debris from megaton weapons, the 
yields for 90Sr, 129I and 131I are estimated to be 0.035, 0.0126 
and 0.029, respectively [H2]. 

220.  The activities of 129I and 131I produced in nuclear tests 
through 1976, which gave rise to globally distributed fallout, are 
thus found to be 4 1011 Bq and 6 1020 Bq.  Owing to the half-lives 
of the two nuclides, practically all of the activity of 129I is 
still present in the environment where it will remain for millions 
of years whereas practically all of the activity of 131I has 
decayed. 

3.  Nuclear fuel cycle

(a)   Nuclear power plants

221.  Iodine-131 and iodine-129 are produced by fission in the fuel 
of nuclear reactors.  The equilibrium activity of 131I per unit of 
electrical power is established after a few weeks of irradiation in 
uranium fuel at about 3 1015 Bq per MW(e) [U1] and increases 
slightly from the beginning of the fuel irradiation to the end, as 
the result of the larger fission yield of plutonium which 
contributes increasingly to power production as the burn-up 
proceeds.  The corresponding activity of 131I produced per unit 
energy generated is about 9 1016 Bq per MW(e)a. 

222.  Small amounts of 131I produced in the fuel may reach the 
coolant of the nuclear reactor through defects in the fuel 
cladding.  In coolant purification or following coolant leakage, 
131I may reach the gaseous and liquid effluent streams.  The 
reported 131I releases in airborne effluents, as summarized by 
UNSCEAR [U1], show an extremely wide variability due mainly to 

different waste treatment systems.  The overall activity discharged 
per unit energy generated is of the order of 107 Bq per MW(e)a from 
PWRs and 108 Bq per MW(e)a from BWRs [I1, U1] while the limited 
data from GCRs and HWRs indicate that the average airborne releases 
are comparable to those from PWRs [U1]. 

223. Airborne iodine can occur in various chemical forms in the 
airborne effluents.  Elemental iodine (I and I2), organic iodine 
with methyl iodide (CH3I) as the simplest organic compound and 
hypoiodous acid (HOI) may be present in significant proportions, 
iodine is also to some extent bound to particulates. 

224. Few data are available on the proportions of organic and 
inorganic forms of the iodine released to the atmosphere. Analyses 
at power stations in the Federal Republic of Germany show that only 
a very small fraction (usually less than 1%) of the iodine released 
in airborne effluents is in particulate form [W1].  Recent 
measurements in the U.S.A. indicate that, in airborne effluents 
from PWRs, on the average 31% of 131I is organic, 40% HOI, 27% 
elemental and 2% in particulate form [P1], whereas in BWRs, the 
proportions were found to be somewhat different, namely 40% 
organic, 20% HOI, 28% elemental and 12% particulate [P2].  These 
proportions, however, are expected to vary significantly according 
to the type of waste treatment in use. 

225. In liquid effluents from LWRs, 131I is found in amounts 
comparable to the airborne releases from BWRs (108 Bq per MW(e)a) 
[U1].  It is not reported in the liquid effluents from other types 
of reactors. 

226. In 1980, the installed nuclear capacity was 1.25 105 MW(e) on 
a worldwide scale [I2].  Assuming an average load factor of 0.6, 
the energy produced was 7.5 104 MW(e)a.  The global production and 
release of 131I at the reactor sites in 1980 are estimated to be 
about 7 1021 Bq and 2 1013 Bq, respectively, using the figures 
given previously for production and release, and assuming that the 
releases are similar to those from BWRs for the reactor types for 
which no data are available.  The estimate of the global release is 
very crude, but nevertheless shows that only a tiny fraction of the 
order of 10-9 of the 131I produced in the reactors is discharged 
into the environment.  Table VI.1 provides a break-down of the 
releases from reactors according to reactor type. 

227.  Being a volatile element, iodine is readily released to the 
atmosphere in the case of a reactor accident.  The two reported 
reactor accidents which have caused measurable irradiation of the 
public occurred at Windscale (U.K.) in October 1957 and at Three 
Mile Island in March 1979.  The release of 131I to the atmosphere 
was about 7 1014 Bq in the Windscale accident [L1] and 6 1011 Bq in 
the Three Mile Island accident [H3]. 

228.  The activity of 129I produced in a nuclear reactor is much 
lower than that of 131I.  The production of 129I per unit energy 
generated is approximately 5 107 Bq per MW(e)a [U1] corresponding 
to an inventory of 1.5 108 Bq per MW(e) after three years of fuel 

irradiation.  Iodine-129 has not been identified in routine 
discharges of nuclear reactors.  A rough estimate of the activity 
of 129I discharged per unit energy generated can be calculated 
assuming that the ratio of the release rate to the inventory in 
reactor fuel is the same for 129I and 131I.  As the activity of 
129I in fuel is at least 2 107 times lower than that of 131I, the 
release rate of 129I per unit energy generated is at most 0.1 Bq 
per MW(e)a; activity of 129I discharged from nuclear reactors in 
1980 would thus be of the order of 104 Bq. 

(b)   Fuel reprocessing plants

229.  At the fuel reprocessing stage of the nuclear fuel cycle (if 
it is undertaken), the elements uranium and plutonium in the 
irradiated nuclear fuel are recovered for reuse in fission 
reactors.  Before reprocessing, the spent fuel elements are stored 
under water until 131I has decayed to insignificant amounts.  
Storage times of six months and one year result in the reduction of 
the activities of 131I originally present in the fuel by factors of 
6.5 106 and 4.3 1013, respectively. 

230. In 1980, the only reprocessing plants operating commercially 
were at Windscale (U.K.) and La Hague and Marcoule (France):  in 
addition there were several small experimental reprocessing 
facilities, such as the one at Karlsruhe (Federal Republic of 
Germany).  The combined capacity of the reprocessing plants was 
much lower than the amount of fuel discharged from reactors 
worldwide.
Table VI.1  Estimated global discharges of 131I from nuclear power 
stations in 1980
-------------------------------------------------------------------------
                           Estimated release    Estimated discharges
                           rates                in 1980
Reactor  Number  Capacity  (Bq per MW(e)a)      (Bq)
type             (MW(e)a)  -----------------    -------------------------
                           Airborne  Liquid     Airborne  Liquid  Total
-------------------------------------------------------------------------
PWR      96      64239     107       108        6 1011    6 1012  7 1012
BWR      62      35170     108       108        4 1012    4 1012  8 1012
HWR      14      5963      107       108        6 1010    6 1011  7 1011
GCR      36      7086      107       108        7 1010    7 1011  8 1011
Other    33      12527     108       108        1 1012    1 1012  2 1012
-------------------------------------------------------------------------
Total    241     124985    -         -          6 1012    1 1013  2 1013
-------------------------------------------------------------------------

231. The activity of 131I released into the environment from 
reprocessing plants depends critically on the storage time: in 
practice with the growing backlog of fuel for reprocessing a 
storage time of one year is common but even a relatively small 
quantity of fuel with a short storage time will dominate the total 
131I releases.  Luykx and Fraser [L2] have expressed the reported 
releases of 131I from  Windscale, La Hague, Marcoule and Karlsruhe 

during 1974-1978 in terms of activity discharged per unit of 
electricity generated.  The results range from less than 1.5 105 to 
7 107 Bq per MW(e)a and are presented in Table VI.2.  The 
activities of 131I discharged in liquid effluents have not been 
reported. 

232.  The discharges of 129I depend upon the specific waste 
treatment at the reprocessing plant.  With regard to airborne 
effluents, the reported activites released per unit of electricity 
generated were, on average during the 1975-1978 time period, 2.7 
106 Bq per MW(e)a at Windscale and 4.8 105 Bq per MW(e)a at 
Karlsruhe, representing about 4% and 1%, respectively, of the fuel 
content [L2].  In a series of measurements from November 1975 to 
August 1977 the average values for the components of 129I 
discharges from Karlsruhe were reported at 74% inorganic, 23% 
organic and 2% aerosol [B1]. 
Table VI.2  Average normalized activities of 129I and 131I
discharged into the environment by fuel reprocessing plants
(Bq per MW(e)a)
---------------------------------------------------------------------------
                   Iodine-129                        Iodine-131
Plant      ----------------------------      ------------------------------
Location   Airborne   Liquid     Total      Airborne   Liquid     Total
           effluents  effluents             effluents  effluents
---------------------------------------------------------------------------
Windscale  2.7 106    5.6 107    5.9 107    2.6 106    N.A. a/   >2.6 106

La Hague   N.A.       N.A.       N.A.       1.1 107    N.A.      >1.1 107

Marcoule   N.A.       N.A.       N.A.       7.4 107    N.A.      >7.4 107

Karlsruhe  4.8 105    N.A.       N.A.       >1.5 105   N.A.      -
---------------------------------------------------------------------------
a/  N.A. = Data not available.

233.  In recent years, the 129I released in liquid effluents was 
only measured at Windscale.  They average at 5.6 107 Bq per MW(e)a 
which corresponds fairly well with the theoretical fuel content 
[L2]. 

234.  The information on 131I and 129I activities discharged per 
unit electricity generated is summarized in Table VI.2. If it is 
assumed for the four reprocessing plants, that all the 129I 
contained in the fuel is discharged into the environment, the total 
129I released in 1978 was about 3 1011 Bq, which is 7 orders of 
magnitude higher than the total activity estimated to be released 
from reactors.  With regard to 131I, it is much more difficult to 
assess the total activity released from fuel reprocessing plants, 
as the activities discharged into liquid effluents have not been 
reported.  However, using the pessimistic assumptions that the 
activity contained in the airborne effluents represents 1% of the 
activity present in the fuel at the time of reprocessing and that 
the rest of the activity is discharged into liquid effluents, it is 

found that the total annual 131I discharges from fuel reprocessing 
plants are about 5 1011 Bq, which is much less than the global 
discharges from reactors. 

C.  BEHAVIOUR IN THE ENVIRONMENT

1.  Nuclear explosions

235.   The behaviour in the environment of 131I produced in the 
nuclear explosions has been extensively studied, especially the 
air-vegetation-milk pathway which is generally the most significant 
route by which humans are exposed.  Much of the literature has been 
referenced in UNSCEAR reports [U1-U5].  Environmental 
concentrations of 131I following large nuclear explosions are 
significant; they are easily measured and this allows the transfer 
factors to be derived from observations.  In contrast, the 
environmental concentrations (and the resulting dose rates) of 129I 
are extremely low and have only been measured in a few studies.  
Although some of the aspects of environmental behaviour of 131I 
following nuclear explosions apply also to 129I, the discussion in 
this section will be limited to 131I. 

236.  Radioactive fallout is observed to circle the earth in 20-30 
days on average [R1] which is approximately the mean residence time 
of an aerosol in the troposphere and is longer than the mean life 
of 131I.  It is thus during its first pass around the earth that a 
given atom of 131I formed in a nuclear explosion will either decay 
in the atmosphere or deposit on the earth's surface.  It is 
unlikely that during such a short period the clouds of debris 
become well mixed.  The ground-level air concentrations of 131I at 
a particular station fluctuate according to meteorological 
conditions and are not necessarily representative of a larger 
region nor of a latitude band [P3]. 

237.  Information on the physical and chemical nature of fallout 
131I is very limited.  In the U.K., late 1961, an average 75% of 
the activity was in particulate form, the rest being in the gaseous 
state [E2] but in the U.S.A. in 1962 the particulate fraction was 
found to vary from 10 to 90% [P4]. These large variations may be 
partly explained by the physical and chemical transformations 
undergone by 131I following its formation: Voillequé [V1] observed 
that the fraction of the total airborne 131I associated with 
particulates is about 0.5 to 0.7 in the first few days following a 
nuclear explosion but that it later decreases to be approximately 
0.3 after two months.  In the gaseous fraction, the proportion of 
organic compounds was found to increase in the same two months time 
period from one fourth to about three fourths of the gaseous iodine 
[V1]. 

238.  Iodine-131 deposition on the ground and on vegetation occurs 
by dry and wet deposition.  The rate of dry deposition can be 
characterized by the deposition velocity on the vegetation which 
was derived to be about 2 10-2 m s-1 [C2, H4] from measurements 
performed during the series of tests of late 1961.  When 
precipitation occurs, fallout 131I is deposited at a much faster 

rate than in dry weather, essentially by rain-out, i.e., in-cloud 
mechanisms, rather than by wash-out, i.e., below the cloud 
processes [U5].  On the other hand, rain washes the surface of the 
leaves and thus removes some of the radio-iodine.  Chamberlain and 
Chadwick in the U.K. [C2] and Hull in the U.S.A. [H4] calculated on 
the basis of their measurements that in late 1961 about 50% of the 
131I falling out in rain was retained on herbage. 

239.  Even though the observed ground-level air activity 
concentrations and deposited activities of 131I vary widely from 
one area to another according to meterological conditions, it is 
possible to obtain a rough estimate of the total activity density 
deposited, weighted over the world's population, from the average 
ratio of the 131I/140Ba deposited activity densities.  Data from 
Argentina covering the years 1966 to 1973 [B2, B3, C3] reveal that 
the 131I/140Ba ratio of the annual deposited activity densities 
varied from 0.4 to 1.3 with a median value of 0.6.  Relevant 
information is also provided by the air concentrations of 131I and 
140Ba which are measured in the stations of the global network of 
the U.K. Atomic Energy Agency [C4].  The annual average of the 
integrated air activity concentration 131I/140Ba ratios of nine 
stations scattered over the world ranged from 0.19 to 3.1 with a 
median value of 0.46.  However, the corresponding ratios of the 
deposited densities were higher as only the particulate fraction of 
131I was measured in the air. Assuming that the particulate 
fraction of 131I represents half of the total activity of that 
nuclide in the air, the median value would be approximately 0.9, 
which is comparable to the Argentinian value of 0.6.  An 
intermediate value of 0.7 will be adopted in this document. 

240.  As the average ratio of the 140Ba/95Zr deposited activity 
density is estimated to be 0.62 [U2] and the population-weighted 
global average deposition density of 95Zr from all tests is 
approximately 2.4 104 Bq m-2 [U2], the population-weighted global 
average deposition density of 131I is thus found to be on the order 
of 104 Bq m-2. 

241.  Fresh milk is usually the main source of 131I in food because 
of the concentration achieved by the grazing animal and the short 
storage period of milk.  Besides, milk plays an important worldwide 
role in the diet of infants.  The relationship between the 
integrated cow's milk concentration and the deposition density has 
been derived from measurements in  Argentina [B2] to be 6.3 10-4 Bq 
a 1-1 per Bq m-2 and to show little variation from year to year.  
The integrated milk concentrations observed throughout the world by 
the fallout network stations have been reported by UNSCEAR [U1-U5]. 

2. Industrial releases

242. The environmental behaviour of radio-iodines released from 
nuclear facilities differs in some aspects from that of fallout as 
the chemical forms are not in the same proportion and as the 
releases occur at discrete points on the surface of the earth both 
in the atmosphere and in the aquatic environment.  As the 
authorities are concerned with the total impact resulting from the 

releases of radionuclides, all the important pathways leading to 
man have been investigated, mainly through laboratory and field 
experiments, and occasionally following unplanned releases.  In 
comparison to 131I and to fallout, much more 129I is released from 
industrial operations and it will be considered in this section, 
together with 131I in the discussion of the local and regional 
aspects, and on its own in the discussion of the global aspects. 

(a)   Local and regional aspects

(i)  Atmospheric releases

243. The behaviour in the atmosphere of the radio-iodines released 
from nuclear facilities is complicated by the various forms that 
iodine may take (particulate, elemental, organic, or as hypoiodous 
acid).  Elemental iodine readily deposits on forage and enters the 
cow-milk-man pathway.  Organic iodine is retained much less 
efficiently by vegetation and its deposition velocity is 200 to 
1000 times smaller than that of the elemental form [A1, H5].  
Particulate associated iodine and hypoiodous acid will be deposited 
at rates intermediate between those for the elemental and organic 
forms [V2]. Physico-chemical transformations occurring during 
atmospheric transport may also affect the distribution of the 
various forms of iodine, since some of them are not stable in 
sunlight.  On the basis of photochemical considerations, the 
atmospheric residence times for I2 are estimated to be less than a 
minute during the day, much shorter than the residence times for 
CH3I and other organic iodides in plant effluents (about 60 hours 
in sunlight) [V1].  The elemental form would be expected to become 
rapidly associated with airborne aerosols, so that deposition at 
distances beyond the immediate vicinity of the release would be 
largely governed by the particulate behaviour.  It must be pointed 
out, however, that the atmospheric residence times of CH3I derived 
from environmental measurements is much longer than that obtained 
from photochemical considerations (about 100 days to be compared 
with 60 hours in sunlight) [V1]. 

244. Numerous field and laboratory experiments have been conducted 
to determine the deposition velocity on vegetation of various forms 
of iodine [B4, H6, V2].  Most of the experiments dealt with 
elemental iodine, for which the deposition velocity was found to 
vary with the temperature, the relative humidity of the air, the 
wind speed and the vegetation density.  The best fit of the 
experimental data is obtained assuming that the deposition velocity 
on vegetation of elemental iodine is proportional to the wind speed 
and to the vegetation density and is an exponential function of the 
temperature and of the relative humidity of the air [A2]. Typical 
values of the deposition velocity on vegetation are 2 10-2 m s-1 
for elemental iodine and 5 10-5 m s-1 for organic iodide.  In the 
case of particulates, a value of 10-3 m s-1 was found representative 
for grass, about 2 10-3 m s-1 for clover and about 3 10-4 m s-1 for 
vegetation-free soil [H7].  Since the deposition velocity varies 
with the vegetation density, less variability is encountered by 
normalizing the values by the mass of dry vegetation per unit area.  
UNSCEAR [U1] adopted a normalized value of 5 10-3 m3 kg-1 s-1 for 

the deposition velocity on grass of radio-iodine in effluents from 
nuclear installations. 

245. Regarding atmospheric releases of radio-iodine, the main 
pathways to man are inhalation and consumption of fresh milk and 
fresh leafy vegetables;  consumption of beef is also taken into 
consideration in the case of 129I. 

246. The assessment of the transfer of iodine to milk requires the 
knowledge of the value of the following parameters in addition to 
the deposition velocity;  the residence half-time of iodine on 
vegetation, the average mass of grass consumed per cow and per day 
under average agricultural conditions during the grazing season, 
the fractional pasture grazing time and the fractional transfer of 
the daily ingested activity by the cow per unit volume of produced 
milk. 

247. The residence half-time of 131I on grass is 3-6 days, most 
estimates lying around 5 days [B4, B5].  This figure seems to be 
valid irrespective of the iodine production source (fallout, 
nuclear plants, experiments) and of the climatic characteristics of 
the region [B4].  The corresponding residence half-time of stable 
iodine (or of 129I) on grass is about l4 days.  The depletion 
mechanisms involved are: transfer to the roots;  volatilization; 
leaching by atmospheric precipitation; mechanical removal by wind, 
rain or other agents;  death or decomposition of the leaves or of 
their surface layer [B4, C2].  The opinions do not agree on the 
relative importance of those mechanisms [B4, C2]. 

248.  The daily grass requirement of a lactating cow is estimated 
at 10 kg dry matter [B4, H6].  The pasture grazing time varies 
according to the climatic conditions and to the cattle management 
practices.  An average grazing time of six months per year was 
assumed by UNSCEAR [U1].  During the winter months, the cows are 
held in the stable and consume dry fodder in which the activity 
concentration of 131I will have decayed to insignificant levels and 
that of 129I can be assumed to be the same as that in herbage. 

249.  The transfer of iodine is usually expressed as the fraction 
present in milk of the ingested activity under equilibrium 
conditions.  This quotient was found to be around 5 10-3 d 1-1 [B6, 
U1].  The value adopted by UNSCEAR in its 1977 report is an upper 
estimate of 10-2 d 1-1 [U1].  In the case of a simple 
administration of 131I to the cow, Lengemann and Comar [L3] 
observed that the maximum concentration is reached within one day 
and that it is followed by a rapid decrease (half-time of about 1.5 
days) in the first 3-4 days and a slower decrease (half-time of 
about 3 days) afterwards. Among the different factors which can 
have an influence on the grass to milk transfer of iodine, the two 
most important may be the milk productivity by animals and the 
season, a higher iodine secretion in milk occurring with a 
productivity increase and in the warm season [B4, G1]. 

250.  Taking account of all the parameters given above, an 
integrated air concentration of 1 Bq a m-3 of 131I or 129I would 

result in an integrated milk concentration for 131I of 160 Bq a 1-1 
and for 129I of 870 Bq a 1-1. Transfer via wet deposition is 
usually insignificant over the course of a year [U1]. 

251.  The transfer of iodine from air to fresh leafy vegetables has 
been assessed by UNSCEAR [U1] on the basis of the values given 
above for the deposition velocity and the residence time on the 
vegetation, and of a dry-to-wet vegetation weight ratio of 0.5.  In 
addition, a fractional removal by washing of 0.4 [U6] and an 
average 7 day marketing delay (resulting in a decay factor of 0.55 
for 131I and 1.0 for 129I) were taken into account.  Time 
integrated conditions of 340 Bq a kg-1 (fresh weight) for 131I and 
1740 Bq a kg-1 (fresh weight) for 129I are calculated for the case 
of a time-integrated air concentration of 1 Bq a m-3. 

252. An estimate of the transfer of 129I has also been carried out 
by UNSCEAR [U1].  Using the values given above for the deposition 
velocity on grass, residence half-time on grass and grass 
consumption rate, the resulting transfer factor is 260 Bq a kg-1 
per Bq a m-3 of 129I in air if the fractional transfer of daily 
ingested activity per unit mass of meat is taken to be 3 10-3 
d kg-1 [P5]. 

253.  In the assessment of the 129I  concentration in grass 
following deposition of that nuclide, the root uptake from soil has 
not been taken into account.  It has been estimated [B7] that this 
pathway could contribute at equilibrium only about 20% of the 
concentration in grass arising from direct deposition. 

(ii)  Aquatic releases

254.  Information on the behaviour of iodine in the aquatic 
environment is rather limited.  The UNSCEAR aquatic model [U1] can 
be used to estimate the transfer of 131I and 129I from the aquatic 
environment to diet for generalized discharge situations.  In that 
model, it is assumed that iodine is not removed to sediments but 
that during treatment of drinking water, 20% of the activity 
contained in raw water is removed. For the two isotopes of iodine 
considered, the concentration factors are taken to be 15 1 kg-1 for 
fresh water fish, 20 1 kg-1 for marine fish and 100 1 kg-1 for 
shellfish.  Further discussion is presented in the section on 
dosimetry. 

(b)   Global aspect

255.  Because of its very long half-life (1.57 107 a) and of the 
mobility of iodine in the environment, 129I may become widely 
distributed on the global scale.  Whether released into the 
atmosphere or into the aquatic environment, 129I will eventually 
reach the oceans in a time period very short in comparison with its 
half-life.  Iodine-129 will then be recycled to the atmosphere and 
the terrestrial biosphere, mainly by evaporation of seawater.  
Atmospheric water is exchanged within about 10 days and it follows 
therefore that there is a rapid exchange of iodine.  The specific 
activity approach has been the usual method to assess the dose 

commitments and the long-term environmental concentrations from 
129I discharges.  According to that approach, the specific activity 
of 129I per unit mass of stable iodine of any environmental 
material of the terrestrial biosphere (including air and food-stuffs) 
will be in the long term equal to that of seawater.  Assuming 6 
1023 g to be the mass of ocean waters with an iodine concentration 
of 0.064 µg per gram of water [T1], the specific activity of 129I 
per unit mass of stable iodine obtained in sea water after a 
release of 1 Bq of that radionuclide into the environment is  found 
to be 2.6 10-17 Bq g-1.  If it is assumed that there is no 
environmental sink for iodine, 129I will be recycled throughout its 
mean life of 2.3 107a. 

D.  TRANSFER TO MAN

256.  Iodine is an element of fundamental importance for the human 
organism since it is an essential component of the thyroid hormone, 
which is necessary for the growth and metabolism of the body.  The 
metbolic cycle of iodine in man, especially in the adult, is 
sufficiently well known in its fundamental behaviour, as a 
consequence of the large number of clinical studies carried out in 
the last years with radioactive isotopes of iodine [B4]. 

257.  The absorption by the blood from the gastro-intestinal tract 
is complete and very rapid.  It is absorbed at the rate of about 5% 
per minute and it can be considered to be complete after two hours 
[B4].  When inhaled in the form of inorganic iodide or as 
methyliodide, a fraction of about 70% is absorbed [M2] whereas more 
than 90% of it is absorbed when inhaled in the form of elemental 
vapour [M3]. 

258.  The uptake by the thyroid of the iodine contained in blood as 
well as the size of the thyroid gland are both very dependent upon 
the daily intake of stable iodine [D2].  The model adopted by ICRP 
[I3] for the metabolism of iodine applicable to adults is based on 
the three-compartment model of Riggs [R2].  ICRP [I3] assumes that 
30% of the iodine entering the blood is translocated to the thyroid 
while the remainder goes directly to excretion.  Iodine in the 
thyroid is assumed to be retained with a biological half-life of 
120 days and to be lost from the gland in the form of organic 
iodine.  Organic iodine is assumed to be uniformly distributed 
among all organs and tissues of the body other than the thyroid and 
to be retained there with a biological half-life of 12 days.  One-
tenth of this organic iodine is assumed to go directly to faecal 
excretion and the rest is assumed to be returned to blood as 
inorganic iodine. 

259.  A quantitative assessment of the transfer of iodine to man 
must take into consideration the variation with age of the various 
parameters involved, which are essentially the mass of the thyroid 
gland, the fractional uptake by the thyroid, the effective 
residence half-time in the thyroid, the breathing rate and the 
consumption rate of foodstuffs.  The values adopted for 129I and 
131I by UNSCEAR in its 1977 report [U1] are presented in Table 
VI.3.  It is to be noted that the metabolic parameters for adults 

are not in complete agreement with those adopted by ICRP [I3]; the 
resulting differences in the thyroid absorbed doses vary according 
to the pathway and the isotope considered but they are in all cases 
less than 50%. 

Table VI.3  Variation with age of the parameters used in the 
assessment of the transfer to man of 129I and 131I [U1]
-----------------------------------------------------------------
                                               Age              
                                   6       4         14    Adult
                                   months  years     years
-----------------------------------------------------------------
Mass of the thyroid   (g)          2       4         14    20
Effective half-time
 in the thyroid (d)   131I         6.0     6.3       6.9   7.6
                      129I         23      28        48    136

 Inhalation pathway
 Fractional uptake by the thyroid  0.30    0.26      0.26  0.26
 Breathing rate       (m3 a-1)     1150    3530      6440  8030

 Ingestion pathway
 Fractional uptake by the thyroid  0.40    0.35      0.35  0.35
 Consumption rate:
   Milk               (1  a-1)     330     180       150   90
   Leafy vegetables   (kg a-1)     0       13        20    30
   Beef               (kg a-1)     0       8         15    27
   Drinking water     (1  a-1)     438     438       438   438
   River fish         (kg a-1)     1       1         1     1
   Ocean fish         (kg a-1)     6       6         6     6
   Shellfish          (kg a-1)     1       1         1     1
-----------------------------------------------------------------


E.  DOSIMETRY

1.  Dose per unit intake

260.  As iodine is selectively taken up in the thyroid gland, its 
concentration in that organ is considerably higher than in the 
other organs and tissues in the body.  The stable iodine 
concentration in the thyroid tissue of an adult is of the order of 
500 µg g-1 while in the rest of the body it is much less than 1 µg 
g-1 [S2].  Since the major contribution of the absorbed doses from 
129I and 131I is due to the emission of the beta-particles, which 
have a short range in the human tissues, the absorbed doses in the 
thyroid are about 1000 times higher than those in the other organs 
and tissues, which will not be considered in this document. 

261.  The significant variation with age of the metabolic 
parameters is reflected in the thyroid absorbed doses per unit 
intake.  Table VI.4 presents the values of the age-dependent 
thyroid absorbed doses per unit intake adopted by UNSCEAR in its 
1977 report [U1].  These values are derived from the metabolic 
parameters presented in Table VI.3 and from the following figures 

for the energy absorbed in the thyroid (MeV) for disintegration:  
0.18, 0.18, 0.19 and 0.19 MeV for 131I and 0.060, 0.061, 0.063 and 
0.064 MeV for 129I for the ages of 0.5, 4, 14 years and adult, 
respectively. 

Table VI.4  Age-dependent absorbed doses in the thyroid gland
per unit intake of 131I and 129I (Gy Bq-1)
------------------------------------------------------------
Thyroid aborbed dose                Age                     
per unit intake       6 months  4 years   14 years  Adult
------------------------------------------------------------
Inhalation: 131I      3.2 10-6  1.5 10-6  4.9 10-7  3.8 10-7
            129I      4.1 10-6  2.2 10-6  1.1 10-6  2.3 10-6

Ingestion:  131I      4.3 10-6  2.0 10-6  6.5 10-7  5.1 10-7
            129I      5.4 10-6  3.0 10-6  1.5 10-6  3.0 10-6
------------------------------------------------------------

2.  Dose per unit release

(a)   Nuclear explosions

262.  The thyroid dose commitment via the ingestion (milk) pathway 
from 131I released by nuclear explosions can be assessed from the 
sequential product of transfer factors 

           Dc = P23 P35 F

where F is the integrated deposition density, P23 is the deposition 
to milk transfer factor and P35 is the milk to thyroid dose 
transfer factor.  The value of F weighted for the world's 
population and that of P23 for fallout deposition were given above 
as 104 Bq m-2 and 6.3 10-4 Bq a 1-1 per Bq m-2, respectively.  The 
value of P35 can be derived from the age-dependent consumption 
rates of milk given in Table VI.3, the age-dependent thyroid doses 
per unit ingested activity presented in Table IV.4, and from the 
assumption that the three groups of children are representative of 
the age groups 0-1, 1-9 and 10-19 years, respectively, and that 
these groups contain respectively 2, 16, and 20% of the population 
[U1]. The value of P35 (milk) is thus found to be 1.3 10-4 Gy per 
Bq a 1-1. The thyroid dose commitment for the world's population 
arising from 131I from global fallout of past nuclear explosions is 
therefore estimated to be about 8 10-4 Gy.  Most of the dose 
commitment was in fact delivered in the early 1960s. Taking the 
world's population at that time to be 3 109 persons, the collective 
dose commitment would be about 2 106 man Gy.  Since 6 1020 Bq of 
131I were estimated to give rise to global fallout, the individual 
and collective thyroid dose commitments per unit activity released 
are found to be 1.3 10-24 Gy Bq-1 and 3 10-15 man Gy Bq-1, 
respectively. 

(b)   Nuclear industry

(i)  Local and regional contribution

263.  Atmospheric releases.  The contribution of the inhalation 
pathway to the collective dose commitments from effluent releases 
can be estimated from the integrated concentrations of 131I and 
129I in ground-level air.  Assuming that all the activity released 
in the atmosphere will eventually deposit on the ground, the 
integrated concentration in ground-level air is the total amount of 
131I or 129I released per unit area of the deposition region 
divided by the deposition velocity, vd.  The population affected is 
the population density deltaN, times the area of the deposition 
region.  The collective dose commitment per unit activity released 
is given by the expression Sc = deltaN phi/vd where phi is the age-
weighted thyroid dose per unit integrated air concentration. 

264.  On the basis of a figure of 5 10-3 m3 kg-1 s-1 for the 
deposition velocity on grass per unit area density of vegetation 
(see paragraph 244), using a value of 0.1 kg (dry) m-2 for the mass 
of grass per unit area, and assuming that the activity deposited on 
grass represents one fourth of the total activity deposited on 
ground, vd would be 2 10-3 m s-1 or 6 104 m a-1.  From the data 
contained in Tables VI.3 and VI.4 and the age distribution given 
above, the age-weighted thyroid dose per unit integrated air 
concentration would be 3.4 10-3 and 1.4 10-2 Gy per Bq a m-3 for 
131I and 129I respectively.  As the deposition area is expected to 
be very large, the population density is assumed to be about 25 
persons km-2, that is 2.5 10-5 persons m-2.  The collective thyroid 
dose commitments are thus estimated to be 1.4 10-12 and 5.8 10-12
man Gy Bq-1 for 131I and 129I, respectively. 

265.  The contribution of the ingestion pathway (consumption of 
milk, leafy vegetables and meat) to the collective dose commitments 
per unit activity released can be assessed by the expression Sc = 
P13 P35 deltaN  where P13 is the air to dietary product to age-
weighted thyroid dose transfer factor, and deltaN the population 
density.  Using the values and assumptions given previously, the 
collective thyroid dose commitments are estimated to be, for 131I, 
8.7 10-12 and 2.3 10-12 man Gy Bq-1 for consumption of milk and 
fresh leafy vegetables, respectively, while the values for 129I 
would be 1.2 10-10, 5.0 10-11 and 6.4 10-12 man Gy Bq-1 for 
consumption of milk, fresh leafy vegetables and beef, respectively. 

266.  Aquatic releases.  The collective thyroid dose commitment per 
unit activity of 131I and 129I discharged into the aquatic 
environment can be estimated, as in the 1977 UNSCEAR report [U1], 
using the expression 

    Sc =     N I f phi    
         V(lambda + 1/tau)

where V is the volume of the receiving waters, tau the turn-over 
time of receiving waters, lambda the decay constant of the 
radionuclide considered, N the number of individuals exposed, I the 

individual consumption rate of the foodstuff considered, f the 
concentration factor of the radionuclide in that foodstuff, and phi 
the thyroid dose per unit activity ingested. 

                           1       
267. The quantity V(lambda + 1/tau) is the infinite time integral 
of the water concentration per unit activity released, while that 
quantity multiplied by f is the infinite time integral of the 
concentration in the consumed item (fish, for example).  For inputs 
into small volumes of water, the concentrations in water and in 
fish will be high, but the populations which can be served with 
drinking water or by fish consumption will be limited.  For inputs 
into large volumes of water, the concentrations will be smaller, 
but the populations involved will be larger.  It is reasonable, 
therefore, to assume as a first approximation that the quantities 
V/N are relatively constant;  they are taken to be 3 107 and 3 109 
litre per man, for fresh water and sea water, respectively.  Using 
the values given previously in the text and in the tables, the 
collective thyroid dose commitments per unit activity of 131I and 
129I can be estimated.  The results are presented in Table VI.5. 

Table VI.5  Collective thyroid dose commitments 
per unit activity of 131I and 129I released 
in the aquatic environment (man Gy Bq-1) 
------------------------------------------------
Type of release         131I          129I 
and pathway 
------------------------------------------------
 Release to fresh water 
  -  Drinking water     3.2 10-13     3.2 10-10
  -  Fish               1.4 10-14     1.4 10-11

 Release to sea water
  -  Fish               1.1 10-15     1.1 10-13
  -  Shellfish          8.8 10-16     1.1 10-14
------------------------------------------------

(ii)  Global contribution

268.  Using the specific activity approach described previously, 
the activity concentration of 129I per unit mass of 127I is the 
same in the sea water and in the human thyroid.  Assuming that the 
concentration of stable iodine per unit mass of thyroid is 80, 180, 
300 and 600 µg g-1 at ages 6 months, 4 years and 14 years and for 
adults, respectively and using the age distribution given 
previously, a specific activity of 1 Bq per gram of stable iodine 
in the thyroid would lead to an age-weighted annual thyroid dose of 
1.5 10-7 Gy.  Since a release of 1 Bq 129I results in a long-term 
concentration of 2.6 10-17 Bq g-1 stable iodine per Bq (see 
paragraph 255), the collective thyroid dose commitment arising from 
discharges of 129I would be about 9 10-7 man Gy Bq-1, assuming a 
world population of 1010 and no sink for iodine in the environment. 

269.  Most of the collective dose commitment to the thyroid is 
delivered in the far future, as the mean life of 129I is 2.3 107 a.  

The average dose rate per unit activity released would be extremely 
low (4 10-24 Gy a-1 per Bq released).  The estimate of the 
collective thyroid dose commitment could be significantly in error 
if there exists a mechanism that efficiently removed iodine from 
the biosphere at a rate significantly greater than the radioactive 
decay constant of 129I of 4 10-8 a-1.  Such a mechanism could be 
the retention by ocean sediments. 

(c)   Summary

270. Table VI.6 summarizes the results given above for the thyroid 
collective dose commitments per unit activity released, and 
provides also the collective effective dose equivalent commitments 
per unit activity released.  The latter quantities are obtained by 
multiplying the former by 0.03. 

Table VI.6  Summary of collective thyroid dose commitments 
per unit activity released and of collective effective dose 
equivalent commitments per unit activity released
---------------------------------------------------------------
                        131I        129Ia    131I       129Ia
                        ------------------   ------------------
                          (man Gy Bq-1)        (man Sv Bq-1)
---------------------------------------------------------------
 Weapon tests            3 10-15b             9 10-17b

 Industrial releases
a)  Atmosphere
   Inhalation           10-12       6 10-12  3 10-14    2 10-13
   Ingestion
   - Milk               1 10-11     1 10-10  3 10-13    3 10-12
   - Leafy vegetables   2 10-11     5 10-11  6 10-14    2 10-12
   - Beef               -           6 10-12  -          2 10-13

b)  Rivers
   Ingestion
   - Water              3 10-13     3 10-10  9 10-15    9 10-12
   - Fish               1 10-14     1 10-11  3 10-16    3 10-13

c)  Oceans
   Ingestion
   - Fish               1 10-15     1 10-13  3 10-17    3 10-15
   - Shellfish          9 10-16     9 10-14  3 10-17    3 10-15
---------------------------------------------------------------
a   First passage estimates.  The long-term, global estimates 
    for 129I from all sources and by all pathways are 9 10-7 man 
    Gy Bq-1 and 3 10-8 man Sv Bq-1. 
b   Milk consumption.

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VII.  CAESIUM-137

A.  INTRODUCTION

271.  Caesium is element number 55 in the periodic table.  It is an 
alkali metal like potassium, and it resembles potassium 
metabolically.  Whereas potassium is an essential element for man, 
there is no evidence that caesium is also an essential trace 
element.  In fact, stable caesium, 133Cs, is fairly rare in the 
biosphere and in geological occurrence.  Average occurrence in the 
earth's crust is 3 µg g-1.  In specific rock types the estimated 
average concentration is 1 µg g-1 in basalts and 5 µg g-1 in 
granite.  The K/Cs ratio in basalts is 7500 [T2].  Stable caesium 
occurrence in fresh water, lakes and rivers ranges between 0.01 and 
1.2 ng g-1 and is 0.5 ng g-1 in the ocean [K2].  Stable potassium 
is more abundant, with usual concentrations of 0.2 to 10 µg g-1 in 
fresh waters and 380 µg g-1 in the ocean [V1]. 

272.  The radioactive isotope 137Cs is produced in nuclear fission 
and is one of the more significant fission products. The fission 
yield is relatively high, about 6 atoms per 100 fissions, 
independent of the type of fission in uranium or plutonium (Table 
VII.1).  It has a radioactive half-life of 30.17 a and its beta 
decay is accompanied by a gamma ray of moderate energy.  Figure 
VII.I shows the decay scheme and lists the primary transition 
energies. 

Table VII.1  Fission yields of 
caesium-137 [C3]
-----------------------------
            Fission yield (%)
Nuclide     -----------------
            Thermal   Fast
-----------------------------
235U        6.21      6.12
239Pu       6.64      6.50
238U                  5.93
232Th                 6.73
-----------------------------

FIGURE VII

273.  The chemical similarity of caesium and potassium and the 
opportunity to make simultaneous measurements by gamma spectrometry 
of 137Cs and naturally-occurring 40K has encouraged the expression 
of 137Cs concentrations relative to the potassium concentration in 
a manner analogous to that used for strontium and calcium.  
However, caesium and potassium are not interdependent and do not 
behave in such a regular manner in biological systems as do 
strontium and calcium.  As the levels of potassium in diet and man 
remain roughly constant (1.4 g per litre of milk and 2 g per 
kilogram of body weight), the 137Cs/K quotients can be converted 
easily to 137Cs concentrations.  Dietary intake of 137Cs increases 
in proportion to the amount of food consumed, however the 137Cs/K 
quotient in diet is relatively constant for adult and children 
diets for widespread contamination situations [G4].  An additional 
advantage of expressing 137Cs levels in the body in terms of the 
137Cs/K quotient is that age and sex differences are minimized and 
the values correlate more closely to 137Cs concentrations per unit 
of lean body mass, which seems to be a more important parameter for 
dosimetric purposes than the whole body mass.  It is expected, 
however, that assessments in the future will be presented 
independently for 137Cs, without so much reliance on the stable 
congener element. 

274.  A great deal of information has accumulated on 137Cs in the 
environment, particularly the measurements of fallout 137Cs in air, 
deposition, diet and man.  Much of the literature has been 
referenced by UNSCEAR over the years. Recent reviews of 137Cs data 
have been published by Moiseev and Ramzaev [M5] and the United 
States National Council on Radiation Protection and Measurements 
[N1].  This document is not extensive in terms of references cited.  
The representative references for most statements can be taken as 
the starting points for the more extensive literature available. 

B.  SOURCES

1.  Nuclear explosions

275.  Atmospheric testing of nuclear weapons has resulted in 
widespread distribution in the environment of radioactive fission 
and activation products.  Extensive measurements of fallout 
radioactivity have been conducted.  The data have been reported and 
discussed in each report of UNSCEAR [U2, U3, U4, U5, U6, U7, U8].  
The stratospheric inventory of 90Sr has been measured in a long-
term programme [L1].  Global networks to measure fallout deposition 
have reported results for 90Sr [F1] and 137Cs [C1].  The activity 
ratio of 137Cs to 90Sr in long-term deposition has been found to be 
relatively constant at about 1.6 [U7], although variations for 
individual samplings are encountered [S1]. 

276.  The total amount of 90Sr produced in weapons testing through 
1980, which has been globally dispersed, is estimated to be 6.0 
1017 Bq [F1].  Less than 1% of this amount remained in the 
stratosphere [L1].  The remainder has been deposited on the earth's 
surface.  This corresponds to 9.6 1017 Bq of 137Cs produced in 
nuclear testing.  Radioactive decay has reduced the cumulative 
deposit of 137Cs to 6.9 1017 Bq [C1], 76% of which is in the 
northern hemisphere and 24% in the southern hemisphere. 

2.  Nuclear fuel cycle

(a)   Nuclear reactors

277.  Caesium-137 is produced by fission in the fuel of nuclear 
reactors.  The amounts produced depend on the degree of fuel burn-
up achieved and to some extent on the type of fuel and the neutron 
spectrum in the reactor.  In fairly high burn-up fuel (33000 MW[t]d 
t-1) of a pressurized water reactor, the 137Cs production is 
estimated to be 3.9 1015 Bq per tonne of fuel, corresponding to 1.3 
1014 Bq per MW(e)a of electricity generated [O1]. 

278.  Small amounts of fission products produced in the fuel in 
nuclear reactors may reach the coolant through defects in the fuel 
cladding.  In coolant purification or following coolant leakage, 
these fission products may reach gaseous and liquid effluent 
streams.  In controlled amounts, some of the effluents are released 
to the environment. 

279.  Reported amounts of 137Cs released to the environment from 
reactors have been summarized by UNSCEAR [U8].  The averaged 
release rates for some reactor types are included in Table VII.2. 

Table VII.2  Estimated global discharges of 137Cs
from nuclear power stations in 1980
--------------------------------------------------------------
Reactor  Reactor  Capacity   Release rate      Estimated
type     number   [MW(e)a]   [Bq per MW(e)a]   discharge (Bq)
--------------------------------------------------------------
PWR      96       64239      6 107             2 1012
BWR      62       35170      9 108             19 1012
GCR      36       7086       2 109             9 1012
--------------------------------------------------------------
Other    47       18490      9 108             10 1012
--------------------------------------------------------------
Total    241      124985                       4 1013
--------------------------------------------------------------

These are not necessarily the typical situation for a particular 
reactor.  During a specific year it is normally the case that a 
somewhat larger release occurs from a single reactor with 
insignificant releases reported from all other sites.  This means 
that there is a large range in release distribution, covering three 
or four orders of magnitude.  For all reactor types the release of 
137Cs is primarily in liquid effluents;  negligible amounts, in 
comparison, occur in airborne effluents. 

280.  Assuming the averaged normalized release rates to be 
representative for the reactor groups, it is possible to obtain a 
very rough estimate of the total amount of 137Cs released from 
reactors worldwide.  Using the installed capacities of the various 
reactor types as of 1980 [I4] and assuming a reactor utilization of 
60%, the estimated annual release from all reactors is about 4 1013 
Bq.  In this calculation, it is assumed for the reactor types for 
which no data are available, that the releases are similar to those 
from BWRs. 

(b)   Fuel reprocessing plants

281.  In fuel reprocessing plants the fuel is dissolved to recover 
uranium and plutonium for re-use.  All of the 137Cs and other 
fission products as well go to the waste streams. The radionuclide 
activities in airborne and liquid effluents from fueld reprocessing 
plants have been recorded by UNSCEAR [U8]. 

282.  Caesium-137 is released from fuel reprocessing plants 
primarily in liquid effluents.  Averaged release rates during the 
years 1971-1972 were 0.6, 90 and 520 109 Bq/MW(e)a from the Nuclear 
Fuel  Services plant (U.S.A.) (no longer in operation), La Hague 
(France) and Windscale (U.K.), respectively [U8].  The release of 
caesium-137 in liquid effluents is small relative to the amount in 
spent fuel.  The fractional liquid releases from the Windscale 
plant were approximately 4 10-3.  The only data for 137Cs in 
airborne effluents is for the Nuclear Fuel Services plant, which 
corresponded to 4 104 Bq/MW(e)a [U8], several orders of magnitude 
less than in liquid effluents. 


C.  BEHAVIOUR IN THE ENVIRONMENT

1.  Fixation in soil

283.  Caesium is generally rather strongly fixed in soil. Downward 
migration and availability to plants is thereby reduced.  In 
mineral soils the movement of 137Cs is appreciably less than that 
of 90Sr.  Three to four years after deposition on the soil surface, 
the median depth to which it has penetrated is usually less than 2 
cm [F5].  Its mobility may be somewhat greater in organic soils.  
Much smaller amounts of 137Cs than of 90Sr are leached out of the 
soil to enter rivers and lakes. 

284.  There are exceptional areas, however, where caesium fixation 
in soil is much less, allowing enhanced transfer of caesium to 
plants.  Marei et al. [M2] identified regions in the USSR where the 
soil is wet, peaty and podzolic, from which transfer of 137Cs into 
the food chain is 10 times higher than for other areas.  Other 
regions of the world where the soils give rise to high 137Cs 
transfer into diet have also been identified, for example, in the 
Faroe Islands, New Zealand and Sweden [U8]. 

285.  It has been reported that in clay minerals the important 
factor in fixation of caesium is the ability of certain layered 
silicates such as micas, vermiculites and illites to adsorb or fix 
trace quantities of caesium [T1].  Caesium ions are trapped in the 
interlayer regions of vermiculite or at the frayed edges of illites 
and micas.  Caesium is thus more strongly retained in soils 
containing predominantly micaceous minerals.  Soils which do not 
contain large quantities of micaceous minerals, such as tropical 
soils, peat soils, and podzolic soils, exhibit less retention and 
allow greater uptake of caesium by plants. 

286.  Fixation of caesium by sediments in aquatic environments 
occurs in a similar fashion to fixation in soil.  The preferential 
adsorption of 137Cs to the micaceous component of sediments has 
been demonstrated under environmental conditions. 

2.  Transfer to plants

287.  Caesium-137 may be transferred to plants by direct deposition 
onto plant surfaces or by root uptake from accumulated deposits in 
soil.  In general, direct foliar absorption is the predominant mode 
of plant contamination when the deposition rate is relatively high.  
Root uptake is low except in those cases mentioned above, when soil 
conditions allow low fixation of caesium. 

288.  Caesium depositing on plant surfaces is retained to the same 
extent as other particulate debris.  A removal half-time of l4 days 
due to weathering is generally assumed.  Once absorbed by the 
plant, caesium is readily redistributed through the plant.  
Relationships between air concentrations and subsequent 
concentrations of 137Cs in pasture plants have been discussed by 
Hawthorne et al. [H4] and Pelletier and Voilleque [P2]. 

289.  Caesium may enter plants by plant base absorption before 
becoming fixed in soil. Thus, retention of 137Cs in the root mat of 
pastures may allow 137Cs to be relatively more available to plants 
for a period of a year or more [U3].  A high level of organic 
matter in soil can enhance the absorption of caesium by plants 
[B1].  Sorption of organic molecules on clay surfaces prevents the 
retention of caesium and also of potassium by these minerals.  
Thus, for permanent pastures in temperate regions, the frequent 
high organic matter content of the upper soil layer allows shallow 
rooted grass to absorb 137Cs relatively more freely for a somewhat 
more extended period following deposition.  Mushrooms have been 
shown to concentrate very effectively 137Cs from soil [G2, M1], 
which may be associated with the highly organic areas of growing. 

290.  Uptake to plants of 137Cs from soil low in available 
potassium may be somewhat increased.  The addition of potassium may 
decrease absorption of 137Cs in this case; however, this has no 
effect when the available potassium is high [N5, F4].  Root uptake 
of caesium is in general included in the range 0.01 to 1, which is 
the ratio of caesium concentrations in the dry plant material to 
that in dry soil [M3]. 

3.  Transfer to milk

291.  The fractional amount of 137Cs transferred into milk is 
slightly greater than that of potassium.  It has been shown that 
some 10% of orally ingested 137Cs is secreted into the milk of 
dairy cows, corresponding to 1.3 to 1.5% of the amount ingested per 
litre of milk [G1, I1, L2].  The transfer of fallout 137Cs in field 
conditions has been found to be somewhat less, ranging from 0.25 to 
0.86% of intake per litre of milk [P2, M1, S4]. 

4.  Transfer to meat

292.  A correlation between 137Cs concentration in beef and in milk 
has been noted, the ratio of concentrations in meat (Bq/kg) and in 
milk (Bq/1) averaging about 4 for production in the same locality 
[L5, E1, J2].  This would correspond to a transfer of about 4% of 
the daily intake per kilogram of meat [F5].  Equilibrium conditions 
are reached in about 30 days in the cow [F5].  It cannot generally 
be expected that there will be a useful relationship between 137Cs 
in meat and milk, since animals produced for the two purposes are 
frequently provided with different diets and are reared in 
different areas. 

5.  Transfer to diet

293.  For general contamination situations, as for global fallout 
radioactivity, the main contributions to dietary intake of 137Cs 
are generally from grain products, meat and milk.  Fruit and 
vegetables contribute much smaller amounts of 137Cs.  This has been 
the pattern for western diets, such as Denmark [A1] and the United 
States [G4].  In Japan the main contribution has been from cereals 
[U1].  For all foods the transfer from a specific deposited amount 
seems to be rapid, being essentially completed within the first two 
years after deposition. 

294.  Marine food chains are of secondary importance in 
contributing to dietary intake of fallout 137Cs, even in countries 
where fish is widely consumed.  In Japan between 1966 and 1971, 
only about 8% of the 137Cs in diet came from the consumption of 
fish products [U1].  As the contributions from other foods 
decrease, however, the relative contribution from fish can 
increase.  During 1976 in Denmark, 15% of the 137Cs intake was 
attributed to fish [A2].  It is possible that individuals consuming 
large amounts of freshwater fish may acquire 137Cs burdens several 
times greater than individuals eating more diversified diets [G3]. 

295.  The transfer of 137Cs from deposition to diet has been 
studied quantitatively using the following transfer function 
between deposition and diet components in regression analyses of 
reported data [U8]: 

             j         j           j infinite
    Cj(n) = b1 f(n) + b2 f(n-1) + b3 sigma   f(n-m) e-µm
                                     m=1

where Cj(n) is the concentration of 137Cs in the diet component j 
during the year (n) in mBq/gK; f(n) is the deposition density of 
137Cs in the current year (n) and in previous years (n-1) in Bq/m2; 
bjn are the proportionality factors, and µ is the decay constant 
accounting for radioactive decay and reduced availability of 
deposition in all previous years.  The first term of the equations 
is the rate term giving the contribution to dietary level from the 
current year's deposition amount.  The second term is the long term 
contribution expressing separately the contribution from the 
previous year's deposition including storage of foods by market 
practices.  The third term gives the contribution from the 
cumulative deposit of 137Cs in soil. 

296.  The quotient of the time-integrated concentration of 137Cs in 
diet and the integrated deposition density defines the transfer 
factor P23.  The integrals are replaced by summations if, as is 
usually the case, the relevant quantities are assessed over 
discrete intervals of time, such as annual averages 

              infinite
              sigma   C(n)
              n=1
       P23 =  -----------
              infinite
              sigma   f(n)
              n=1

The transfer factor is usually expressed in mBq a/gK per Bq/m2.  It 
may be evaluated for total diet or for dietary components. 

297.  Using the model described above, the evaluation of the 
transfer factor reduces to 

                       infinite -µm               e  
    P23 = b1 + b2 + b3 sigma   e   = b1 + b2 + b3 -----
                       m=1                        1-e

The parameters of the transfer function, obtained by regression 
fits of the 137Cs/K quotients in milk, dietary components and total 
diet from several countries have been reported by UNSCEAR [U8].  
Some of these results are given in Table VII.3. 

298.  The lowest values for the transfer factor for milk are 
obtained for the U.S.A., Denmark and the U.K.  Most of the transfer 
of 137Cs deposition to milk is from direct deposition.  Less than 
15% of the transfer is from uptake of 137Cs from soil.  
Intermediate values of P23milk are obtained from the USSR and 
Argentina.  Transfer from direct deposition is increased and uptake 
from soil is more significant.  The highest values of P23milk are 
obtained for New Zealand, Australia, Norway and the Faroe Islands.  
A high value has also been reported for Finland, which is 15.8 mBq 
a (gK)-1 per Bq m-2 [C2].  These results can be explained by 
efficient transfer of direct deposition and soil conditions which 
allow only low fixation of 137Cs. 

299.  The range of values of transfer factors for milk, 3.4 to 27.5 
mBq a (gK)-1 per Bq m-2, indicates that it is not easy to specify a 
typical transfer situation.  To estimate the 137Cs transfer to milk 
in a particular region requires some indication of soil and pasture 
conditions. 

300.  Using the comprehensive data from Denmark of 137Cs in diet 
[A2], values of the parameters of the transfer function of various 
components of diet and for the total diet have been determined.  
For most foods, the major contribution to the value of the transfer 
factor comes from direct deposition. Only a small transfer is 
attributed to uptake from soil.  The values of the parameters for 
the fit to total diet data are: b1 = 1.6, b2 = 2.2, b3 = 0.04 mBq 
a(gK)-1 per Bq m-2 µ = 0.11 a-1 and P23 = 4.1 mBq a(gK)-1 per 
Bq m-2. Short-term transfer to diet is thus estimated to comprise 
93% of the total transfer, (b1 + b2)/P23. 

301.  The general pattern of 137Cs transfer to diet can be expected 
to be similar to that for Denmark, although increased transfer can 
occur in specific areas, depending on soil conditions or particular 
consumption habits.  UNSCEAR has used a rounded value of P23 of 4 
mBq a(gK)-1 per Bq m-2 to assess the dose commitments from 137Cs 
for the world population [U8]. 


Table VII.3  Parameters of the transfer functions between deposition density and  
137Cs/K in milk
-----------------------------------------------------------------------------------------
Parameter  U.S.A.  Denmark  United   USSR  Argentina  New      Australia  Norway  Faroe
a/                          Kingdom                   Zealand                     Islands
           1973    1976     1977     1972  1976       1977     1973       1976    1976
-----------------------------------------------------------------------------------------
b1         2.3     2.2      1.7      4.4   3.2        6.8      7.3        3.5     6.8
b2         1.3     1.2      1.6      0.2   0.0        0.9      4.8        2.4     5.5
b3         0.0     0.01     0.04     0.3   1.7        3.1      0.3        2.5     2.8
µ          0.05    0.03     0.07     0.2   0.3        0.5      0.2        0.2     0.2
P23        3.4     3.7      3.9      6.2   7.5        13.0     13.4       15.6    27.5
-----------------------------------------------------------------------------------------
a/  The unit for parameters b1, b2, b3 is mBq a(gK)-1 per Bq m-2.
    The unit for parameter µ is a-1.
    The unit for the transfer factor P23 is mBq a(gK)-1 per Bq m-2.

6.  The lichen-caribou-man foodchain

302.  A terrestrial situation which allows much greater than usual 
transfer of caesium to man is the lichen-caribou-man foodchain.  
Caesium depositing on lichens is retained quite effectively.  There 
is a very slow decrease in activity with time, approximately 5 to 
10% of the 137Cs being eliminated annually [U5].  Lichens provide 
the food base for grazing caribou and reindeer during the winter.  
A proportionality factor of 137Cs in lichen to that in reindeer 
meat in northern Sweden during the winter averaged 4.9 ± 0.4 [L4]. 
In summer also grass and herbaceous plants are consumed by the 
animals, and therefore the 137Cs levels in meat show marked 
seasonal variation. 

303.  High concentrations of 137Cs arise in Lapp and Eskimo 
populations who eat the caribou and reindeer meat.  Levels of 740-
1300 Bq per kg of body weight were observed in individuals during 
1964 from fallout [H1, M4], about a factor of 100 greater than 
burdens of individuals from temperate northern hemisphere regions.  
Other fallout radionuclides, such as 54Mn and 55Fe, and the natural 
isotopes 210Pb-210Po are also concentrated along the lichen-
caribou-man foodchain. 

7.  Aquatic behaviour

304.  Caesium in the aquatic environment is strongly adsorbed by 
suspended particulate materials, especially clays. Therefore, the 
amount of caesium in the soluble phase decreases with increasing 
suspended solid concentrations. Potassium is also sorbed, but to a 
much less degree. 

305.  Caesium levels in fish are inversely related to the potassium 
content of the water [K1].  Because of the high concentration of 
potassium in the ocean, the transfer of 137Cs to fish is of primary 
concern in the freshwater environment. The activity of fresh water 
fish may be 100 times that of ocean fish, given the same caesium 
concentrations in water. 

306.  The low mineral content of fresh water also enhances the 
absorption of 137Cs by aquatic plants.  Aquatic plants from fresh 
water areas, which are sometimes important in cattle feed, may have 
increased levels of 137Cs compared to 137Cs which may have 
deposited on nearby pasture ground. 

307.  Caesium in aquatic animals is accumulated primarily from the 
food chain.  Absorption efficiency of potassium and caesium from 
food is high.  In animals the excretion rate of potassium is about 
3 times larger than that of caesium.  As a result, the caesium 
concentration per unit amount of potassium in tissues increases by 
a factor of about 3 with each trophic level [P1]. 

308.  The food web also accumulates caesium from suspended and 
bottom sediments.  Filter feeders may accumulate caesium adsorbed 
to particulate matter.  Benthic invertebrates obtain caesium 
absorbed to ingested bottom sediments.  Fish ingest those 
invertebrates and also some sediment particles along with the prey.  
Absorption efficiency in the fish depends on the caesium fixation 
ability of the sediment minerals. 

309.  The concentration factors (ratio of concentration in organism 
to that in the water) for caesium must be related to the potassium 
concentration in the water and to the turbidity.  From a literature 
review of values for fresh water systems [V1], suggested values of 
the concentration factors are 1000 for algae and plants, molluscs 
and invertebrates in all waters and 5000/Kw and 1500/Kw for non-
piscivorous and piscivorous fish, respectively, in clear waters, 
where Kw is the stable potassium concentration of water in µg/g.  
The factors for fish are a factor of 5 less in turbid waters (> 50 
µg/g suspended solids).  The concentration factors for caesium in 
the ocean are 10 for algae and molluscs, 30 for fish and 50 for 
molluscs [F6]. 

D.  TRANSFER TO MAN

1.  Absorption and distribution in tissues

310.  As a general rule, caesium compounds are soluble in body 
fluids.  Intestinal absorption is complete (100%) under 
experimental conditions [R2, S3], but from normal diets is probably 
less efficient, ranging from 50 to 80% [F5].  In man caesium is 
secreted into the gastrointestinal tract, between the stomach and 
small intestine, and is readily reabsorbed [I2].  One basis for 
therapeutic treatment in internal contamination cases is to 
administer solutions of Prussian blue, which binds with caesium in 
the gastrointestinal tract preventing reabsorption [I2, D1]. 

311.  Caesium migrates rapidly into cells of the body following 
intake and becomes relatively uniformly distributed in soft tissues 
[R2, L3, R4].  The metabolism in mothers and infants has also been 
studied.  There seems to be no placental discrimination, as the 
newborn has 137Cs concentrations about equal to that of the mother 
[B3]. 

312.  Concentrations of caesium and potassium are low in fat 
tissues.  Therefore, for equal 137Cs concentrations in intake, the 
concentrations of 137Cs (Bq per kg body weight) in males are higher 
than in females, due to the higher average proportion of fat tissue 
in the female body.  However, a difference is also expected due to 
longer retention time of caesium in males.  Expressed in Bq 137Cs 
per gK, the difference between males and females is somewhat 
reduced. 

313.  It has been inconclusive for some time whether 137Cs 
concentrates in bone. One study reported that the concentration of 
137Cs in rib bones, which were free of muscle but not of marrow, 
was comparable to the concentration in soft tissues [Y2].  The 
results were variable, and subsequent studies both did and did not 
confirm these results [A3, H3, N4].  From the results of a recent 
study it appears that caesium associated with bone is present in 
the marrow portion with only slight uptake by the hard tissue [H2]. 

314.  A slower turnover of caesium in bone could allow 
concentrations in bone to lag behind those in tissue, causing 
higher relative concentrations in bone during periods of decreasing 
intake.  A longer retention half-time would eventually be noted in 
whole-body measurements.  However, such a component has not been 
identified in over 1000 days of measurement following an acute 
intake case [R4]. 

2.  Retention half-time

315.  A great many investigations of the biological half-time in 
man have been conducted.  Many of the references are collected in 
the discussion by Lloyd [L6]. The half-time in man varies 
considerably, depending on age and other factors. The half-time is 
less in women than in men, and the half-time in children and 
infants is less than in adults.  Pregnant women have shorter 
caesium half-times than in their non-pregnant conditions.  Table 
VII.4 shows the summary of 137Cs retention half-time reported 
recently by the NCRP [N1], using the data of Lloyd et al. [L7] and 
Zundel et al. [Zl]. 

Table VII.4  Retention half-time of  137Cs in the 
human body [N1]
--------------------------------------------------
Subjects         Number   Age        Half-time (d)
--------------------------------------------------
Men              26       23-55 a    105 ± 25
Women            15       20-51 a    84 ± 20
Pregnant women   24       16-39 a    49 ± 16
Children         7        5-17  a    57 ± 20
Infants          5        17-143 d   19 ± 8
--------------------------------------------------

316.  The biological half-time for caesium in man can be considered 
a function of age for juveniles and of sex for adults, but it is 
not determined by body mass [L6].  Half-time and body mass may, 
however, be dependent on some other common factors. The rate of 

caesium turnover may be under hormonal influence or control or may 
reflect the general metabolic rate [L6]. 

317.  Shorter half-time components of 137Cs in man have been 
reported, including one of only 2 to 3 hours [N3].  In general, two 
components of the 137Cs half-time in man have been established:  a 
small fraction (10 to 15%) excreted with a short half-time (1 to 
1.5 days) and the remainder excreted more slowly (50 to 150 days) 
[R1, R3].  The ICRP suggests representative values of retention of 
10% with a half-time of 2 days and 90% with a half-time of 110 
days.  Integral retention is 143 Bq d per Bq intake, contributed 
almost entirely by the long-term component. 

3.  Transfer factor

318.  The value of the transfer factor P34 relating concentrations 
of 137Cs in diet and man can be derived from the integral retention 
by dividing by the potassium content of the body (140 g) and 
multiplying by the daily potassium intake (3.3 g d-1) [I5].  The 
result is 3.4 Bq a (gK)-1 in man per Bq a (gK)-1 in diet. 

319.  The relatively short biological half-time of caesium in the 
body makes it possible to assess the transfer factor P34 from the 
measured 137Cs/K quotients in diet and man integrated over a few 
years.  Using this procedure, an average value of 3 Bq a (gK)-1 per 
Bq a (gK)-1 diet is derived [U6, U7, U8]. 

E.  DOSIMETRY

1.  Dose per unit intake

320.  The dose from 137Cs in tissue is due to the beta particles 
from 137Cs decay and to the photon, x-rays, and conversion and 
Auger electrons from decay of the daughter, 137mBa.  A portion of 
the photon energy will escape from the body, depending on the body 
size.  Calculations have been performed for uniform distributions 
of 137Cs in the body for a range of proportions and masses 
corresponding to infants, children and adults [N1, F1]. 

321.  The average dose rate within the body for a uniform 137Cs 
concentration of one Bq per kg body weight is about 3.5 nGy/d from 
beta particles in both adults and infants plus 3.2 nGy/d from 
photons in the adult and about 1.6 nGy/d in infants.  The totals 
are 6.7 nGy/d per Bq/kg in the adult and 5.1 nGy/d per Bq/kg in the 
infant. 

322.  For 140 gK in the 70 kg adult body, the dose rate 
corresponding to 1 Bq 137Cs per gK would be 

            µGy d-1   140 gK  365 d       µGy/a
    0.0067  --------  ------  ----- = 4.9 -------
            Bq(kg)-1  70 kg   a           Bq/gK

Spiers [S2] also obtains this result and estimates the 
corresponding dose rate in a child weighing 8 kg of 

                µGy/a
          4.1  -------
                Bq/gK

323.  UNSCEAR [U6, U7, U8] assessed the transfer factor between 
tissue and dose, P45, as 4.9 µGy per Bq a (gK)-1.  The value is 
nearly independent of age, being only slightly less for the child.  
For a single uptake of 137Cs, the integral retention per unit 
intake is 143 Bq d per Bq intake (paragraph 317).  The average 
absorbed dose in the body per unit intake is thus 

          µGy/d     Bq d  1
   0.0067 ----- 143 ----- ----- = 1.4 10-8 Gy per Bq intake.
          Bq/kg     Bq    70 kg

324.  This result applies to intake by ingestion.  For inhalation 
the value is less by a factor of 0.63 (for particles of 1 µm size) 
due to fractional deposition in the lungs of inhaled amounts.  
Following inhalation there is only a short retention in the lung 
(half-time = 0.5 d) for the soluble caesium compounds and a small 
dose primarily from the beta particles before the 137Cs becomes 
distributed throughout the body. 

2.  Dose per unit release

(a)    Nuclear explosions

325.  The dose commitment via the ingestion pathway from 137Cs 
released by nuclear explosions can be assessed from the sequential 
product of transfer factors 

          Dc = P23 P34 P45 F

where F is the integrated deposition density.  The values of the 
transfer factors as derived above are:  P23 = 4 10-3 Bq a (gK)-1 
per Bq m-2, P34 = 3 Bq a (gK)-1 per Bq a (gK)-1, and P45 = 4.9 10-6 
Gy per Bq a (gK)-1.  The dose commitment from 137Cs ingestion per 
unit widespread deposition density, such as from nuclear 
explosions, is thus 6 10-8 Gy per Bq m-2. 

326.  The total amount of 137Cs released to the environment by 
nuclear tests, 9.6 1017 Bq, has given a population-weighted 
integrated deposition density of 3100 Bq m-2 in the world as a 
whole [U8].  The world population is 4 109.  With these values, the 
collective dose commitment per unit activity of 137Cs released is 
estimated to be 8 10-13 man Gy per Bq (ingestion). 

327.  The dose commitment via the inhalation pathway is determined 
from the integrated concentration of 137Cs in air, which is 
estimated from the integrated deposition density (Bq m-2) divided 
by the deposition velocity (m s-1).  As for 90Sr, the average 
deposition velocity can be taken to be 2 cm s-1.  Thus, 1 Bq m-2 

integrated deposition density corresponds to 5.8 10-4 Bq d m-3 in 
air.  With the above estimates of integrated deposition density, 
total amount of 137Cs released by nuclear tests, world population, 
breathing rate (22 m3 d-1) and dose per unit intake (8.8 10-9 Gy 
Bq-1), the collective dose commitment per unit activity of 137Cs 
released is estimated to be 1 10-15 man Gy per Bq (inhalation). 

328.  For the external exposure pathway it is assumed that the 
deposited 137Cs becomes exponentially distributed in soil with a 
mean depth of 3 cm.  This gives a dose rate in air of 8.9 109 Gy 
per Bq m-2 [B2].  The mean life of 137Cs in soil is 43.5 years, 
determined by its radioactive decay.  The dose to air from the 
integrated deposition density is, thus, 3.9 10-7 Gy per Bq m-2.  
The dose to tissue is determined by a factor of 0.8 to account for 
the change of material (air to tissue) and back-scatter and 
shielding afforded by other tissues of the body and a factor of 0.4 
to account for building shielding and time spent indoors [U8].  The 
combined factor is 0.8 x 0.4 = 0.32.  The transfer factor P25 
relating integrated deposition of 137Cs to tissue dose is thus 3.9 
10-7 x 0.32 = 1.2 10-7 Gy per Bq m-2.  Using the above values of 
integrated deposition density, total amount of 137Cs released, and 
world population, the collective dose commitment per unit activity 
of 137Cs released is estimated to be 1.6 10-12 Gy per Bq (external 
exposure). 

(b)   Nuclear installations

329.  The contribution of the inhalation pathway to the collective 
dose commitment for effluent releases can be estimated from the 
integrated concentration of 137Cs in air.  This can be derived 
from a dispersion formula or from an estimate of the deposition 
velocity.  In the latter case, the integrated concentration in air 
is the total amount of 137Cs released per unit area of the 
deposition region divided by the deposition velocity, vd.  The 
population affected is the population density deltaN times the area 
of the deposition region.  The areal dependence is removed by the 
product of these quantities.  The collective dose commitment per 
unit activity released is given by the expression 

     c
    S1 = I deltaN phi/vd

where I is the individual intake rate of air and phi is the dose 
per unit activity of 137Cs inhaled.

330.  Using a deposition velocity of 0.5 cm s-1 for near surface 
releases, a population density of 25 man km-2, an air intake rate 
of 22 m3/d and the dosimetric factor of 8.8 10-9 Gy per Bq intake, 
the collective dose commitment for the inhalation pathway per unit 
activity released is estimated to be 1 10-14 man Gy/Bq. 

331.  The contribution of the ingestion pathway from airborne 
effluents to the collective dose commitment per unit activity 
released, Sc1, can be assessed by the expression 

     c
    S1 = P23 P34 P45 deltaN

Using the values for the transfer factors given previously, and 
assuming a constant population density of 25 man km-2 in the region 
of deposition, the collective dose commitment for the ingestion 
pathway per unit activity released is estimated to be 2 10-12 man 
Gy/Bq.  This value assumes that food is locally produced and that 
the production suffices for the population density under 
consideration.  It also applies for the case of 137Cs becoming 
relatively rapidly fixed in soil (P23 = 4 mBq a [gK]-1 per Bq m-2).  
For other types of soil conditions or special consumption patterns 
and also for other population densities, the estimate should be 
adjusted accordingly. 

332.  For the external exposure pathway the transfer factor P25 
relating integrated deposition of 137Cs in soil to the tissue dose 
has been assessed above with regard to nuclear explosions to be 1.2 
10-7 Gy per Bq m-2.  This value is of general applicability.  
Similarly to the ingestion pathway, the collective dose commitment 
per unit activity released, Sc1, can be assessed by the expression 

     c
    S1 = P25 deltaN

Assuming a population density of 25 man km-2, the collective dose 
commitment from external exposure per unit activity released is 
estimated to be 3 10-12 man Gy/Bq. 

333.  For aquatic ingestion pathways from the input of 137Cs to
water bodies the collective dose commitment, normalized per unit 
activity released, can be estimated [U8], using the expression 

     c   ksigma  Nk  Ik  fk  phi
    S1 = -----------------------
            V(lambda + 1/tau)

where V is the volume of the receiving waters, tau is the turnover 
time of receiving waters, lambda is the decay constant of 137Cs, Nk
is the number of individuals  exposed by pathway k, Ik is the 
individual consumption rate of pathway item k, fk is the 
concentration factor for the consumed item in pathway k, and phi is 
the collective dose per unit activity ingested collectively by the 
exposed group. 

                          1        
334.  The quantity  V(lambda+1/tau) is the infinite time integral 
of the water concentration per unit of activity released, while the 
quantity multiplied by fk is the infinite time integral of the 
concentration in the consumed item k.  For radionuclide inputs into 
small volumes of water, the concentrations in water and in fish 

will be high, but the population which can be served with drinking 
water or by fish consumption will be limited.  For inputs into 
larger volumes of water, the concentrations will be smaller, but 
the populations involved will be correspondingly larger. It is 
reasonable, therefore, to assume as a first approximation that the 
quantities Nk/V are relatively constant, independent of V.  The 
values for these quantities as well as values for the other 
parameters of the above expression have been extensively discussed 
by UNSCEAR [U8]. 

335.  A listing of the values used in the assessments presented by 
UNSCEAR is given below [U8]: 

    Parameter                    fresh water   sea water
1.  tau, turnover time of        10 a          1 a
    receiving water
2.  Correction factor for        0.3           1.0
    sediment removal
3.  V, water utilization factor  3 107 1/man   3 109 1/man
    N
4.  fk, concentration factor
    for item k
        drinking water           0.2
        fish                     400           30
        shellfish                              30
5.  Ik, consumption rate for
    item k
        drinking water           440 1/a
        fish                     1 kg/a        6 kg/a
        shellfish                              1 kg/a

(c)    Summary

336.  Table VII.5 summarizes the values obtained above for the 
collective dose commitments per unit of 137Cs activity released in 
airborne and liquid effluents.  These are also the values of the 
collective effective dose equivalent commitments with Sv replacing 
Gy, since the quality factor is one and the dose is assumed to be 
uniform in all tissues.  The largest collective dose commitments 
result from airborne discharges due to the external exposure and 
ingestion pathways.  These estimates are for a generalized release 
situation, and substantial variations could be expected in site-
specific cases. 

Table VII.5  Summary of collective dose 
commitments per unit 137Cs activity 
released [10-14 man Gy per Bq]
------------------------------------------
                              All tissues
------------------------------------------
 Nuclear explosions
  External exposure           160
  Ingestion                   80
  Inhalation                  0.1

 Nuclear installations
  Release to air  a/
    External exposure         300
    Ingestion                 200
    Inhalation                1

  Release to fresh water      
    Fish                      50
    Drinking water            10

  Release to salt water
    Fish                      0.08
    Shellfish                 0.02
------------------------------------------
a/  Assumes population density of 25 
    man km-2.

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VIII.  RADON

A.  INTRODUCTION

337.  Radon is element number 86, but the term generally refers to 
the isotope 222Rn, the decay product of 226Ra.  Other isotopes of 
radon, 220Rn and 219Rn, are generally referred to by their 
historical names, thoron and actinon, respectively.  Radon has a 
half-life of 3.82 days and decays by alpha-particle emission to a 
polonium isotope (218Po), which by further decay through isotopes 
of lead, bismuth, polonium and thallium ends the uranium (238U) 
decay chain with stable lead (206Pb). 

338.  Radon is a chemically inert gas.  It arises from decay of 
radium which occurs in soil and other common materials. Part of the 
radon produced diffuses into the surrounding environment. 

339.  The short-lived decay products of radon, called radon 
daughters, are 218Po (3.05 min), 214Pb (26.8 min), 214Bi (19.7 min) 
and 214Po (1.6 10-4s). They become largely attached to aerosols in 
air and if inhaled they are partly deposited in the human 
respiratory tract.  The radiation doses caused by inhalation of 
radon daughters in air constitute the main part of the natural 
radiation dose to man.  The radiation dose caused by radon itself 
is minor by comparison with that of radon daughters. 

340.  The levels of radon and radon daughters in air depend on the 
source and on the dilution in the air.  The levels are normally 
higher indoors than outdoors.  Reduced ventilation may cause radon 
released from building materials to build up in enclosed spaces.  
The radon levels may be very high in underground mines, 
particularly in uranium mines. 

B.  SOURCES

1.  Outdoors

(a)    Natural radon

341.  The main source of radon outdoors is radium in the earth's 
crust.  The concentration of uranium and radium in the ground 
varies with the types of rocks and minerals.  The concentration of 
radium in rocks and soil is often (but not always) the same as that 
of uranium.  Fractionated dissolution and transport of uranium-234 
and/or radium can cause breaks in the uranium chain [S2]. 

342.  The total amount of radium in the outer 10 km of the earth's 
crust is of the order of 1024 Bq.  Most of the radon produced by 
decay of radium is physically attached to the radium-bearing 
material and only a small part diffuses out into the air.  Other 
relatively less important sources of radon in air outdoors are 
plants, ground water, oceans, etc. The sources contributing to the 
total amount of radon are given in Table VIII.1.  It is assumed 
that the radon exhalation rate from the land (soil) is 0.02 Bq m-2 
s-1 and from the oceans is 70 µBq m-2 s-1. The total production 

rate of radon is of the order of 1020 Bq a-1;  the equilibrium 
inventory in the atmosphere, determined from the total production 
rate divided by the decay constant (66.2 a-1) is estimated to be 
1.5 1018 Bq. 

Table VIII.1  Sources of radon in the global 
atmosphere [H5]
---------------------------------------------
Source                      Radon production
                            per year (Bq)
---------------------------------------------
Soil                        9 1019
Plants and ground water     < 2 1019
Oceans                      9 1017
Houses  a/                  3 1016
Natural gas                 3 1014
Coal                        2 1013
---------------------------------------------
a/  The value for houses is estimated in this 
    document assuming 109 reference houses 
    (see Table VIII.3).  The true value may be 
    between 5 1015-1017 Bq a-1.

(b)    Mines and mine tailings

343.  Sources of radon of local interest include the tailings from 
uranium and phosphate mining and milling and from geothermal power 
stations.  The radon exhalation rate from tailings depends on the 
radium content of the tailings, on the emanation factor (fractional 
release of radon) and on the land reclamation (overburden).  The 
radon exhalation rate from uncovered uranium tailings varies from 
0.5 Bq m-2 s-1 or less to 10 Bq m-2 s-1 or more [S9, U4].  The 
thickness and area of the tailings per unit mass of uranium in the 
ore can also vary depending on the tailings engineering and 
therefore the radon exhalation rate varies in the range of about 
10-100 Bq s-1 per MW(e)a [S9, U4].  By covering the tailings with a 
few meters of soil the radon exhalation rate is reduced by one to 
several orders of magnitude.  The radon exhalation from tailings of 
phosphate mining and milling also depends on land reclamation and 
the type of ore.  A range of exhalation rates of about 0.01-1 Bq 
m-2 s-1 has been reported [R1]. Geothermal power stations may cause 
radon releases into the air by releases of radon from the water 
which is depressurized at the surface.  A radon release of the 
order of 1011 Bq per MW(e)a has been reported [M3].  The radon 
release from coal-fired plants is 3 to 4 orders of magnitude less. 

344.  The sources of radon in underground spaces like mines are 
radium in the rock and minerals of the mine and radon in water.  
The total release of radon into the mine depends on many factors:  
the uranium-radium concentration of the ore; the number and size of 
cracks in the ore; the isolation of abandoned spaces; the random 
concentration and amount of water; the isolation of water; and the 
ventilation system principles.  By using a normal range of total 
ventilation rates (10-1000 m3 s-1) and a normal radon concentration 
(0.1-10 kBq m-3) the range of radon release into the mine may be 

estimated at 1-104 kBq s-1.  Since there is normally an inverse 
proportionality between radon concentration and ventilation rate, 
lower ventilation rates are more often related to higher radon 
concentrations and vice versa.  There is also a radon release from 
the mine to outside air and it can be compared with the radon 
release from uncovered tailings.  Mining of 100 tonnes uranium per 
year results in 5000-15000 m2 of tailings per year depending on the 
percentage of uranium in the ore and the tailings engineering.  The 
radon release will be in the range of 1-10 kBq s-1.  In non-uranium 
mines radon-rich water is often a significant source of radon to 
the mine [S5].  The radon concentration of the water may be in the 
range of 100-1000 kBq m-3. 

2.  Indoors

(a)    Building materials

345.  Radon in houses comes from building materials, the soil under 
the house, the water and the domestic gas.  Radium concentrations 
in building materials have been investigated. The data indicate 
that some materials such as aerated concrete with alum shale and 
phospho-gypsum from sedimentary ores have significantly higher 
radium concentrations than others and cause enhanced radon 
concentrations indoors.  If these materials are excluded, the 
average concentrations of radium in building material is about 100 
Bq kg-1.  Materials with low activity are wood, natural gypsum, 
sand and gravel.  The radon exhalation rate from walls, floors and 
ceilings is dependent on the radium concentration, the emanation 
power, the diffusion coefficient in the material and the qualities 
and thickness of any applied sealant on the surfaces. 

346.  The radon exhalation rate from uncovered building materials 
varies by several orders of magnitude from about 10-6 Bq m-2 s-1 
(gypsumboard, fiberboard, chipboard, bricks) to (0.1-10) 10-3 Bq 
m-2 s-1 for concrete of different origins and qualities [J5, P3,  
S17, M8, W3].  The radon exhalation rate per Bq Ra/kg varies less, 
e.g., (1.6 ± 0.88) 10-5 Bq m-2 s-1 per Bq Ra/kg for some building 
materials in the Federal Republic of Germany [W3].  If the same 
values are normalized to an emanating power of 1%, the radon 
exhalation rate is (4.4 ± 1.9) 10-6 Bq m-2 s-1 per Bq Ra/kg per 
percentage emanating power.  In the first case the standard 
deviation is 54% and in the last case it is 43%.  By painting, 
plastering or application of wall-paper on the wall the radon 
exhalation may be reduced by less than a factor of 5 [W3, M8]. 

(b)    Soil

347.  The contribution of radon from the soil into a building 
depends on the thickness and tightness of the base structure. The 
exhalation from the soil is of the order of 10-2 Bq m-2 s-1 and a 
concrete floor in cellars should normally reduce the radon 
exhalation from soil into the building by a factor of 10 or more. 
Even so, radon from soil may contribute significantly to the radon 
concentrations in a house, particularly in the cellar and in wooden 
houses.  In some areas houses are built on natural uranium deposits 

(Canada) [L2, K2], on phosphate-related land (Florida, U.S.A.) [U3, 
U13], on waste products from uranium industry (Colorado, U.S.A.) 
[C6], and on waste products from alum production from radium-rich 
alum shales (Sweden) [S22].  In these cases the radon exhalation 
rate may be several orders of magnitude higher than from normal 
soil. 

(c)    Water

348.  Another source of radon in a house may be radon-rich water.  
The relative radon release depends on the use of water.  Boiling 
and splashing of the water increase releases and consequently the 
highest radon releases occur in washrooms, at shower-baths and in 
the kitchen during cooking. The resultant radon concentration in a 
house depends on the amount of water used, the volume of the house 
and the ventilation.  Several studies have been made to estimate 
the relative significance of radon from water [P1, C2, N1], and a 
typical value of the air-to-water radon concentration quotient is 
about 10-4.  Measurements of radon in water are most often made in 
areas with suspected high concentrations because of uranium 
deposits and estimates of the weighted average radon concentration 
in water for a country are rare. 

349.  Population-weighted average concentrations of radon in 
drinking water have been estimated for a few countries and are 
found to be 40 kBq m-3 in Finland [A5, C3], 7 kBq m-3 in Sweden 
[S8] and 0.4-4 kBq m-3 in the Federal Republic of Germany [M9].  
The corresponding radon release into a house can be estimated by 
assuming a daily use of 500 1 per person and 10-100% relative 
release from the water. 

(d)    Natural gas

350.  Natural gas containing radon may also be a source of radon in 
houses.  Gas is transported as purified gas in long transmission 
lines and distributed to the homes or bottled under pressure as 
propane for sale as liquified petroleum gas (LPG).  The radon 
concentrations in natural gas at the production wells are found to 
vary from undetectable values up to about 40 kBq m-3 [U12, H6].  
During supply, transit, storage and delivery the radon 
concentration decreases to an approximate average of the order of 1 
kBq m-3 for both natural gas and LPG (in U.S.A.) [U12, B1, G1]. 

(e)    Summary

351.  The relative contribution of different radon sources to the 
total radon input in a house is estimated in Table VIII.2 with some 
typical values of radon concentration and releases. The radon in 
outside air is brought into the house by ventilation.  The volume 
of the house is assumed to be 200 m3 and the inner surface area 350 
m2.


Table VIII.2  The relative significance of different radon sources
in a reference house
----------------------------------------------------------------------
Source             Radon flux  Comments
                   (Bq d-1)
----------------------------------------------------------------------
Building material  70 103      Emanation rate 2 10-3 Bq m-2 s-1

Water              4 103       1000 1 d-1 and 4 kBq m-3, 100% release
                         
Outside air        9 103       Radon concentration outdoors 0.004 kBq
                               m-3; ventilation rate 0.5 per hour

Natural gas        3 103

Liquified          0.2 103
petroleum gas
----------------------------------------------------------------------

C.   BEHAVIOUR IN THE ENVIRONMENT

1.   Release from soil

(a)    Emanation

352.  The mechanism of radon release from rock, soil and other 
materials is not very well understood and is probably not always 
the same.  The main physical phenomena are recoil and diffusion of 
the radon atom through imperfections of the crystalline structures 
of the radium-bearing particle followed by a secondary diffusion, 
which depends on the porosity of the material [A1].  High porosity 
increases the diffusion rate. The release rate from a material 
depends also on its moisture content:  if the moisture content is 
very low the radon release is decreased by the effect of 
re-adsorption of radon atoms on surfaces in the pores.  If the 
moisture content increases slightly, the radon release increases up 
to a certain moisture content, above which the release of radon 
decreases again owing to a decreasing diffusion rate in water-
filled pores [M7]. 

(b)    Diffusion

353.  Once radon has entered the air or water surrounding the 
emanating radium-bearing particle, it is transported by diffusion, 
earth-mechanical and convective flow, percolation of rain water and 
flow of ground water.  The diffusion mechanism can be expressed by 
the equation 

          Cx  =  Co exp [-x/ ´(D/lambda)]

where Cx is the radon concentration at distance x in air or water 
from the emanating surface; Co is the radon concentration at the 
surface; D is the diffusion coefficient (gas kinetic) and lambda is 

the decay constant [G2].  The diffusion constant D is about 10-2 
cm2 s-1 in air and 10-5 cm2 s-1 in water.  This means that it takes 
on the average about 13 days for a radon atom to diffuse 5 m in air 
or 5 cm in water.  In that time the radon would decay by almost a 
factor of 10. Accordingly, long distance transport of radon in air 
and water mainly depends on the other mechanisms mentioned above, 
which are the transport of air and water itself. 

(c)    Exhalation

354.  The radon concentration Cz in soil air at depth z below the 
surface depends on the diffusion coefficient D, the emanating 
factor a (0 < a < 1), the fractional pore space of the soil f, the 
radium activity concentration Cr (per unit volume of soil) and the 
decay constant of radon g according to the equation [J1] 

         a x Cr
    Cz = ------ [1 - exp (- ´(lambda/D) z]
           f

The exhalation rate is expressed by the equation [J1]

           d(Cz)
    R = D [-----]
            dz    z=o

The combination of the above two equations gives

        lambda a Cr
    R = ----------- ´(D/lambda)
             f

If a = 0.1; Cr = 0.07 Bq cm-3; f = 0.3 and D = 0.01 cm2 s-1;
lambda = 2 10-6 s-1, the exhalation rate R is 3 10-2 Bq m-2 s-1.

355.  The diffusion rate and thereby the exhalation rate is 
influenced by meteorological factors such as rainfall, snowfall, 
freezing and variations in atmospheric pressure. An increase in 
these parameters will decrease the exhalation rate.  Measured 
values of radon exhalation rate from soil vary between about 0.0002 
and 0.07 Bq m-2 s-1 [G2, W1].  The radon exhalation from sea water 
per unit area and time is about two orders of magnitude less. 

2.  Dispersion in air

356.  The dispersion of radon in air is influenced by the vertical 
temperature gradient, the direction and strength of the wind and 
the air turbulance.  The dispersion of radon daughters is also 
influenced by precipitation.  The vertical distribution of radon 
and its daughters in air can be calculated from the following 
system of differential equations [J2]: 

   d      C1
   --- (K --) - lambda1 C1 = 0
   dz     dz        

   d      C1
   --- (K --) + lambdai-1 Ci-1 - (lambdai + LAMBDA) Ci = 0
   dz     dz        

where C1 is the concentration of radon atoms in air at the height 
z;  ni is the concentration of radon daughter i in air at the 
height z; lambda1 is the decay constant of radon; lambdai is the 
decay constant of radon daughter i; LAMBDA is the removal rate of 
radon daughters caused by washout and rainout.  Boundary conditions 
to the above equations are 

Ci (z=0) = 0 for i > 1 and Ci(z->infinite) = 0 for i = 1,2,3, ...

By assuming a constant radon exhalation from an infinite plane 
(ground surface) which equals the radioactive decay of the total 
radon content in the atmosphere it is possible to solve the first 
two equations in this paragraph, which in combination with 
different values of the turbulent diffusion coefficient K [J1, J2] 
give the vertical distribution of radon and radon daughters for 
different atmospheric stabilities. 

357.  Measured values of the relative distribution of radon in air 
are shown in Figure VIII.I.  Although measured values in this case 
are found to follow the predicted vertical distribution fairly 
well, the models described above should only be taken to serve as 
rough guidance for the prediction of radon daughter levels.  The 
varying radon exhalation rate on land and on sea and varying 
meteorological conditions may cause distribution patterns different 
from those predicted by the model. 

FIGURE VIII

358.  At ground level the time-variation of radon concentrations 
depends on the variation of the radon exhalation rate and of the 
vertical dispersion of radon. The effect of increased vertical 
dispersion of radon by turbulence during spring, as compared with 
autumn, outweighs the greater exhalation rate of radon during late 
spring and summer, as compared with autumm and winter.  The overall 
effect is a seasonal variation of the radon concentration at ground 
level with a minimum in the spring and summer and a maximum in the 
autumn and winter observed in several measurements [M1, B2, R2, 
Me].  Diurnal variations of the radon concentration in air 
at ground level also occur because of different varying turbulent 
mixing:  the concentrations are maximum in the early morning and 
minimum in the afternoon.  The variations are generally less than 
one order of magnitude [R2, J3]. 

359.  For estimation of the dispersion of radon released from a 
point (for instance, a geothermal plant or a mine ventilation 
outlet) the most commonly used statistical model is the Gaussian 
plume equation [S18, P2, G4].  The estimated concentrations at 
different distances are therefore dependent on local meteorological 
conditions, terrain roughness, etc. However, in the case of a 
continuous release, the daily variations are smoothed out and an 
annual average is obtained, which differs from place to place only 
according to persistent and substantial local differences. 

360.  The dispersion of radon released from extended sources like 
mill tailings can be estimated from dispersion formulas, assuming 
the extended source to consist of a number of small point sources.  
The relative concentration of radon released from a point source is 
approximately inversely proportional to the p-power of the distance 
d from the source.  If the concentration Cd at distance d is 
expressed relative to the concentration C1 at the reference 
distance d1 the expression is 

          d  -p
    Cd = (--)   C1
          d1     

The formula approximately gives the relative concentration at 
distances more than 1 km if the reference distance d1 = 1km and 
p = 1.2 - 1.5. 

361.  The dispersion and relative vertical distribution of the 
radon daughters in air mainly follow the behaviour of radon. Owing 
to deviating atmospheric parameters for radon daughters as compared 
with radon (e.g., precipitation by rainout and washout) there is 
seldom equilibrium between radon daughters and radon and between 
the different radon daughters.  The long-lived decay products of 
radon (210Pb, 210Bi and 210Po) behave in the troposphere as 
aerosols with residence times of the order of ten days and more.  
Because of their long physical half-life there is no simple 
correlation between these nuclides and radon. 


3.   Indoor behaviour

362.  For closed spaces (e.g., a mine or a house) a theoretical 
correlation may be established between radon concentration in air 
and radon input (exhalation and transport by inlet air) and 
ventilation rate.  The change of the radon concentration in the 
enclosure is given by the following equation 

  d C(t) = R (S) + Ak + Colambdanu - C(t) (lambda + lambdanu)
  dt          V    V 

where C(t) is the radon concentration in the air at the time t;  R 
is the radon exhalation rate from unit surface in the room;  S is 
the emanating surface area;  V is the volume of the space, Ak is 
the radon release from the source (water, gas); Co is the radon 
concentration in the inlet air; lambdanu is the ventilation rate 
(h-1); and lambda is the decay constant of radon. 

363.  At equilibrium the radon concentration in the enclosure is 

        R(S) + Ak + Co lambdanu
    C =   V    V               
          lambda + lambdanu

In homes 0.1 < lambdanu < 3 h-1, and since lambda = 7.6 10-3 
h-1 and lambdanu >> lambda, the above equation takes the form
          
          R(S) + Ak
    C =     V    V   + Co
          lambdanu         
               
As long as lambdanu >> lambda and Co is negligible, the radon 
concentration indoors increases in direct proportion to the 
decrease in ventilation rate.  As the ventilation rate increases 
from 0 to 0.1 and to 1 h-1, the radon concentration decreases by 
factors of 13 and 10, respectively. 

364.  In view of the strong influence of the ventilation rate, 
there are great variations of the radon levels as the effective 
ventilation of a room is changed. This is caused by meteorological 
conditions (wind, pressure, temperature) and by human activities 
like opening doors and windows.  There may be variations of the 
radon concentration in air caused by changes of the radon 
exhalation rate from surfaces, which in turn can be caused by 
changes of atmospheric pressure [J6].  Diurnal variation in houses 
have been studied in several long-term measurements [S11, D1, S21, 
H4, J7, S12, M6] and variations of the order of ten and more may 
occur.  Maxima during the night and early morning and minima at 
noon are usually found, but for several reasons that is not always 
the case.  Only a few studies have been reported on the seasonal 
variations of radon concentration indoors.  The variations of the 
monthly averages are found to be less than a factor 3 [F1, S13].  
Examples of measured variations of radon concentration in houses 
are shown in Figure VIII.II [M10]. 

FIGURE VIII.II

365.  In mines and other underground spaces there are also diurnal 
and seasonal variations of the radon concentration. The diurnal 
variations are most often minor if the ventilation is unchanged by 
the seasonal variations may be large with maxima during the summer 
and minima during the winter. This is caused by the change from 
winter to summer in the temperature gradient from outside to inside 
the mine (Figure VIII.III). 

FIGURE VIII.III


4.   Radon daughter concentrations

(a)    Concentration expression

366.  The concentration of radon daughters can be expressed in 
terms of their activity or of their potential alpha energy, the 
latter being the total alpha energy emitted during the decay of the 
atoms present down to 210Pb.  For any mixture of radon daughters in 
air the potential alpha energy is the sum of the potential alpha 
energy of all daughter atoms in the air.  A unit of exposure which 
is used in mines is the working level (WL).  It is defined as any 
combination of short-lived radon daughters per litre of air that 
will result in the emission of 1.3 105 MeV of alpha energy in their 
decay to 210Pb. 

367.  Another quantity of interest in connection with radon 
daughters is the equilibrium factor F defined as the ratio of the 
total potential alpha energy for the given daughter concentration 
to the total potential alpha energy of the daughters if they are in 
equilibrium with radon.  If the unit WL is used, the equilibrium 
factor F can be calculated as 
          
        a Calpha
    F = ---------
          C

where Calpha is the potential alpha-energy concentration in WL of 
radon daughters; C is the radon activity concentration in Bq 1-1; 
and a is a constant (a = 3.7 Bq 1-1/WL). 

368.  For a room having a known ventilation rate lambdanu (air 
changes per hour) it is possible to calculate the equilibrium 
factor F.  The relationship between F and lambdanu is shown in 
Figure VIII.IV.

FIGURE VIII.IV

369.  The product C x F, where C is the radon concentration and F 
is the equilibrium factor is called the equilibrium equivalent 
concentration of radon (EEC); it corresponds to a concentration of 
radon for which the radon daughters in equilibrium with radon have 
the same potential alpha energy as the actual daughter 
concentration of interest. 

(b)    Attachment

370.  The radon daughters in air may be unattached (free atoms or 
ions) or attached to aerosols.  The first daughter, 218Po, is at 
the time of formation an unattached ion or neutral atom.  But 
within a few seconds most of the 218Po becomes attached to an 
aerosol and the subsequent decay products 214Po and 214Bi are 
therefore to a great extent attached to aerosols at their 
formation. 

371.  The attachment rate of a free radon daughter depends 
therefore on the number and size distribution of the aerosols in 
the air.  These parameters vary in different rooms and will thus 
affect the attachment rate.  In a house with normal aerosol 
concentration (ca. 104 cm-3) and size distribution, the attachment 
rate will be about 10-2 s-1, i.e., the mean life of the free radon 
daughter will be about 100 s.  In a mine with higher aerosol 
concentration the corresponding values may be about 0.3 s-1 and 
4 s, respectively. 

372.  Radon daughters in room air will also attach to the surfaces 
in the room.  The deposition rate for radon daughters attached to 
aerosols is dependent on the diffusion rate of the aerosols and the 
proportion between the surface area and the volume of the room.  If 
that proportion is 2 m-1, the mean life of the attached radon 
daughters (as far as deposition is concerned) is of the order of 
one hour.  Unattached radon daughters have much higher diffusion 
rate than aerosols and therefore the deposition rate is also 
higher.  The corresponding mean life is of the order of one minute. 

373.  The fraction of unattached radon daughters in room air is 
also dependent on radioactive decay and ventilation rate. With 
given values of the attachment rate of free atoms and deposition 
rate of unattached atoms, the fraction of unattached daughters in 
air increases with increasing decay constant and ventilation rate.  
This means that the fraction of unattached 218Po atoms (lambda = 
13.6 h-1) is normally higher than, e.g., that of 214Pb (lambda = 
1.6 h-1).  Measured values are in the range of 1-30%. 

(c)   Equilibrium variations

374.  Ventilation and deposition to surfaces both prevent the radon 
daughters reaching equilibrium with radon in air.  Only the 
dependence of the equilibrium factor on the ventilation rate was 
considered in Figure VIII.IV.  However, because of deposition, 
measured values of the factor F are often lower than the predicted 
value from Figure VIII.IV.  The deviation is larger in air with low 
aerosol concentrations.  An approximate expected value of F is 
obtained by multiplying the value in Figure VIII.IV by 0.5. 

375.  Measured values of F in houses show great variation mainly 
due to differing ventilation conditions.  In the UNSCEAR 1977 
report [U12] an average value of F for houses of 0.5 was adopted.  
In outdoor air the equilibrium factor is also dependent on 
meteorological factors.  Measured values indicate an average value 
of 0.6, which was used by UNSCEAR in its 1977 report.  For uranium 
mines with good ventilation a factor of 0.3 may be appropriate. 

D.   TRANSFER TO MAN

376.  The transfer to man of radon and radon daughters occurs from 
the inhalation of air.  A negligible amount arises from decay of 
radium in ingested food and water.  From a dose standpoint, it is 
most important to know the intake amount of radon daughters in air. 

377.  The amount of radon daughters inhaled depends upon the 
concentration in air and on the breathing rate.  The breathing rate 
varies with different levels of physical activity and age.  For the 
adult, the average breathing rates are 20 1 min-1 during light 
activity, 7.5 1 min-1 resting and 12.5 1 min-1 for intermediate 
activity [I1].

378.  To compute the average air intake rate, it will be assumed 
that the time spent indoors per day (19 h) consists of 5.5 h light 
activity, 8 h resting and 5.5 h intermediate activity.  The time 
spent outdoors (5 h) is assumed to consist of 2 h light activity 
and 3 h intermediate activity.  This gives estimated intake rates 
of approximately 15 m3 d-1 indoors and 5 m3 d-1 outdoors. 

379.  The deposition of radon daughters in the respiratory system 
depends on the size distribution of the aerosols to which they are 
attached and on the fraction of unattached radon daughters.  The 
deposition is also influenced by the manner of breathing.  The 
attached radon daughters are deposited in the pulmonary region.  
However, the deposition is not 100 per cent. Some is exhaled and 
some is transported by mucus before decay. An approximate value of 
the fractional retention may be about 50% [U12] but great 
variations have been reported. 

380.  The unattached radon daughters are mainly deposited in the 
upper respiratory tract.  The efficient deposition of unattached 
daughters has been experimentally verified in a model lung by 
Chamberlain and Dyson [C4].  However, a major part of the 
unattached daughters is removed by nasal deposition [U12]. 

E.   DOSIMETRY

1.   Dose per unit exposure

381.  The doses of radon gas in air are negligible in comparison 
with those of the daughters.  On rare occasions, when there is a 
great disequilibrium between radon and radon daughters, the 
relative contribution from radon to effective dose equivalent could 
be significant. 

382.  Inhalation of radon daughters leads to inhomogeneous 
irradiation of the respiratory tract.  The maximum dose is received 
by the basal cells in the epithelium of the upper bronchi in the 
tracheobronchial region of the lungs due to deposition of 
unattached radon daughters.  Normally, the fraction of unattached 
radon daughters in air is small (a few per cent).  Attached radon 
daughters are mostly deposited in the pulmonary region. 

383.  Jacobi [J3] has calculated the doses from radon daughters to 
the tracheobronchial region of the lung (assumed mass 45 g) and to 
the pulmonary region (955 g) as a function of the unattached 
fraction, f, of the potential alpha energy. The absorbed energy 
fractions are 0.03 (1 + 6f) joule in the tracheobronchial region 
per joule inhaled and 0.38 (1 - f) joule in the pulmonary region 
per joule inhaled.  Total energy absorbed in the lung, reflecting 
deposition and clearance for an average aerosol size and breathing 
rate is of the order of 40%. 

384.  The value of the unattached fraction, f, of radon daughters 
in air is usually in the range 0.02 to 0.1 [U12]. For a mean value 
of 0.06, the absorbed doses in the lungs are 0.9 Gy per joule 
inhaled in the tracheobronchial region and 0.4 Gy per joule inhaled 
in the pulmonary region. 

385.  The dose equivalent is obtained by multiplying by a quality 
factor, which for alpha radiation is 20.  The effective dose 
equivalent is obtained by multiplying by the tissue weighting 
factor, which for the lungs is 0.12.  It might be appropriate to 
apply a weighting factor of 0.06 to the dose equivalent to the 
basal cell layer of the tracheobronchial region and 0.06 for the 
pulmonary region.  The contributions to the effective dose 
equivalent to the lungs per unit energy inhaled are thus 1.1 Sv J-1 
(tracheobronchial region) and 0.5 Sv J-1 (pulmonary region). 

386.  The dosimetry of radon daughters in the lungs is, at the 
moment, under review.  The comparisons of various dosimetric models 
may lead to recommendations for adjustments of some of these 
results.  For this report a value of 1.6 Sv per joule inhaled will 
be used for the total effective dose equivalent to the lungs, 
applicable to exposures indoors and outdoors and also to 
occupational exposures in mines. 

387.  The potential alpha energy of an atom in the decay chain of 
radon is the total alpha energy emitted during decay of the atom up 
to 210Pb.  Dividing by the decay constant of the radionuclide gives 
the potential alpha energy per unit of activity.  The values for 
radon and its short-lived daughters are given in Table VIII.3. 

388.  The measured concentration of radon in air must be multiplied 
by the equilibrium concentration (EEC).  For an EEC of 1 Bq m-3 and 
a breathing rate of 20 m3 d-1 (7300 m3 a-1), the effective dose 
equivalent rate is 

  Bq     m3                     J      Sv
1 -- 7300   34500 MeV 1.6 0-13 --- 1.6 -- = 6 10-5 Sv a-1
  m3     a                     MeV     J

The effective dose equivalent per unit integrated concentration of 
radon in air is 6 10-5 Sv per Bq a m-3. 

Table VIII.3  Potential alpha energy of radon and 
short-lived decay products
-----------------------------------------------------
Radionuclide  Energy per atom   Energy per unit
              (MeV)             activity (MeV Bq-1)
-----------------------------------------------------
222Rn         19.2              9.15 106
218Po         13.7              3620
214Pb         7.69              17800
214Bi         7.69              13100
214Po         7.69              0.002
-----------------------------------------------------
Total a/                        34500
-----------------------------------------------------
a/  The total is the sum of the potential energies 
    of the daughters only.

389.  In terms of WLM for occupational exposures, the potential 
alpha energy inhaled is calculated as follows: 

    1.3 105 MeV 1-1 1200 1 170 h 1.6 10-13 d  
            WL           h     M           Mev

    =  4.2 10-3 J (WLM)-1

where the breathing rate is 1200 1 h-1 (20 1 min-1) during an 8 h 
working day.  The effective dose equivalent per unit exposure is 

          4.2 10-3  d  1.6 Sv = 7 10-3 Sv (WLM)-1
                   WLM     J

The results are summarized in Table VIII.4.

Table VIII.4  Effective dose equivalent per unit exposure from
short-lived radon decay products
----------------------------------------------------------------
Public (indoors and outdoors)    5.10-5 Sv (Bq a m-3)-1
Workers                          7 10-3 Sv (WLM)-1
----------------------------------------------------------------

390.  The effective dose equivalent caused by inhalation of radon 
without daughters at a concentration of 1 Bq m-3 is about 7 10-7 Sv 
a-1 [H8].  This is only about one per cent of the effective dose 
equivalent caused by inhalation of radon daughters in equilibrium 
with radon. 

391.  Radon in water may cause a radiation dose to man by ingestion 
of water and by inhalation of the radon daughters produced by decay 
of the radon released to air.  Consumption of 0.5 1/day of radon-
rich water with a radon concentration of 1 kBq 1-1 will lead to an 
effective dose equivalent of 0.5 mSv a-1 (by ingestion) [S10]. 

392.  The levels of radon have been measured repeatedly in 
different places of the world.  The levels found in outdoor air 
vary between about 0.1 to 10 Bq m-3, lower values having been found 
over oceans and islands, higher values over continents.  For the 
estimation of an average dose equivalent, a radon concentration in 
air of 3.7 Bq m-3 might be used, corresponding to an equilibrium 
equivalent concentration of 2.2 Bq m-3 (equilibrium factor = 0.6).  
If it is assummed that people are outdoors 20% of the time as an 
annual average (i.e., the occupancy factor is 0.2) the 
corresponding average effective dose equivalent would be 1.5 10-2 
mSv per year from radon daughters. 

393.  In houses the radon levels may be high because of radium-rich 
building materials and/or poor ventilation. However, normal values 
are of the order of 10 Bq m-3. An inhabitant-weighted average value 
for countries in which comprehensive measurements have been carried 
out is an equilibrium equivalent concentration of radon of 14 Bq-3. 
Using the effective dose equivalent per unit exposure given in 
Table VIII.4 for the public, 0.06 mSv (Bq a m-3)-1, and an 
occupancy factor of 0.8,  an equilibrium equivalent concentration 
of 14 Bq m-3 will correspond to an effective dose equivalent of 0.7 
mSv per year.  A summary of the dose estimates for indoor and 
outdoor exposures is given in Table VIII.5. 

2.   Dose per unit release

394.  Estimates of the dose per unit release of radon are quite 
variable, depending on the location of the release, the population 
density, and the conditions of the dispersion.  The most 
generalized estimates of the collective effective dose equivalents 
to the world population (4 109 person) may be made by combining 
dose values of Table VIII.5 with the total amounts of radon 
released of Table VIII.1.  For a release of radon outdoors (1020 Bq 
a-1 world-wide) the estimate is 1 10-15 man Sv per Bq, and for a 
release indoors (3 1016 Bq a-1 world-wide) the result is 9 10-11 
man Sv per Bq. 

Table VIII.5  Average effective dose equivalent from inhalation
of radon daughters in air
----------------------------------------------------------------
            Equilibrium     Occupancy  Dose factor   Effective
            equivalent      factor     (mSv per Bq   dose
            concentration              a m-3)        equivalent
            (Bq m-3)                                 (mSv a-1)
----------------------------------------------------------------
Outdoors    2.2             0.2        0.06          0.03

Indoors     14.0            0.8        0.06          0.7
----------------------------------------------------------------

395.  Releases of radon from nuclear installations result from 
mining and milling operations and from the residual tailings. These 
generally occur in areas of low population density - from 3 man km-2 
in mining areas to about 25 man km-2 around the mills [U12].  The 
local collective dose per unit release can be estimated by an 
integration over the distance from 1 to 100 km, assuming an 

atmospheric dispersion factor of 5 10-7 s m-3 at 1 km from the 
release and a reduction in concentration inversely proportional to 
the 1.5 power of the distance expressed in kilometres [U12].  Using 
the dose factor of 0.06 mSu per Bq a m-3 and population densities 
of 3 and 25 man km-2, the range of estimates of the collective 
effective dose equivalent commitment is 3 10-16 to 3 10-15 man Sv 
per Bq.  This does not yet account for limited outdoor occupancy. 
Thus, an average estimate of the collective dose per unit release 
is roughly the same as the generalized estimate obtained above for 
an outdoor release, namely 1 10-15 man Sv per Bq. 

F.  REFERENCES

A1  Andrews, J.N. and D.F. Wood.  Mechanism of radon release in 
    rock matrics and entry into groundwaters.  Trans. Inst. Min. 
    Metall, (IMM) Sect. B:  Applied Earth Science 81: 198-209 
    (1972). 

A5  Asikainen, M. and H. Kahlos.  Natural radioactivity of ground- 
    and surface-water in Finland.  Institute of Radiation 
    Protection report STL-A24.  Helsinki, 1977 (in Finnish). 

B1  Barton, C.J., R.E. Moore and P.S. Rohwer.  Contribution of 
    radon in natural gas to the natural radioactivity dose in 
    homes.  Oak Ridge National Laboratory report 0RNL-TM-4154 
    (1973). 

B2  Blifford, J.H. Jr., H. Friedman, L.B. Lockhart, Jr. et al.  
    Geographical and time distribution of radioactivity in the air.  
    J. Atmos. Terr. Phys. 9:  1-17 (1956). 

C2  Castrén, O., M. Asikainen, M. Annanmäki et al.  High natural 
    radioactivity of bored wells as a radiation hygienic problem in 
    Finland,  in Proceedings of the Fourth International Congress 
    of the International Radiological Protection Association 
    (IRPA).  Paris, April 1977. 

C3  Castrén, P.  The contribution of bored wells respiratory radon 
    daughter exposure in Finland.  p. 1364-1370  in Natural 
    Radiation Environment III.  CONF-780422 (1980). 

C4  Chamberlain, A.C. and E.D. Dyson.  The dose to the trachea and 
    bronchi from the decay products of radon and thoron. Br. J. 
    Radiol. 29:  317-329 (1956). 

C6  Culot, M.V., H.G. Olson and K.J. Schiager.  Radon progeny 
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IX.  PLUTONIUM

A.  INTRODUCTION

396.  Plutonium, the element of atomic number 94, is a member of 
the actinide series of elements, those of atomic number 89 
(actinium) through 103.  The actinide elements have similar 
chemical properties and are also similar to the lanthanide or rare 
earth elements of atomic number 57 (lanthanum) through 71.  The 
actinides are considered a second rare earth series. Elements 
beyond uranium (atomic number 92) in the periodic table are called 
transuranium elements.  Some environmental information on another 
transuranium element, americium (atomic number 95), is also 
available and will be considered here. 

397.  Plutonium occurs naturally in very small quantities.  It is 
formed continuously in uranium ores by neutron capture, the 
neutrons being produced by spontaneous fissioning of uranium. The 
uppermost layers of the earth's crust contain a few kilograms of 
239Pu and about the same amount of primordial 244Pu.  Plutonium 
naturally occurring can only be detected in the richest uranium 
ore. 

398.  The most important of the 15 plutonium isotopes, 239Pu, has a 
half-life of 24065 years and is produced from uranium in nuclear 
reactors: 

    238U (n,gamma)  239U  beta  239Np  beta  239Pu
                          ->           ->       

One of the plutonium isotopes, 241Pu, decays by beta-particle 
emission with a 14.4 year half-life to americium-241 which has a 
half-life of 432.2 years.  The decay properties of 238Pu, 239Pu, 
240Pu, 241Pu, and 241Am are listed in Table IX.1.  Americium is 
produced by successive neutron capture reactions by plutonium 
isotopes in a reactor: 

   239Pu (n,gamma)  240Pu (n,gamma)  241Pu  beta  241Am
                                             ->

399.  Plutonium can exist in four valence states in aqueous 
solutions:  III,IV,V and VI.  The IV state is the most common under 
physiological conditions where it will exist in solution only as a 
strongly complexed ion.  Weak complexes of Pu(IV) in neutral 
solutions will form polymeric hydroxides.  Plutonium oxidizes 
rapidly and, thus, the very insoluble PuO2 is the most common form 
in the environment, although Pu(VI) has been reported in oceans 
[N1] and drinking water [L1].  Some plutonium will be complexed 
with biological ligands and incorporated in micro-organisms or in 
plant or animal tissues. 

400.  Americium can exist in three valence states in aqueous 
solutions:  III, IV and VI.  The trivalent state is the stable form 
under physiological conditions.  The most common oxide is AmO2, 
which is more soluble than PuO2.  Americium appears to be more 
readily incorporated into biological materials when dispersed to 
the environment than is plutonium. 

401.  Plutonium isotopes, 238Pu, 239Pu and 240Pu and the americium 
isotope 241Am, emit alpha radiation.  Since the x rays accompanying 
the alpha emissions are low energy, concentrations of these 
isotopes that might occur in the environment would not cause 
biological effects unless they are incorporated in biological 
material.  Deposition in the lungs and absorption from the gastro-
intestinal tract following ingestion are the most important routes 
of entry into the bodies of animals and human beings. 

Table IX.1  Decay information for plutonium and 
americium isotopes
---------------------------------------------------
Isotope  Half-life   Decay   Energy a/   Intensity
         (years)     mode    (MeV)       (%)
---------------------------------------------------
238Pu    87.74       alpha   5.59        0.716
                             5.55        0.283
                             5.45        0.001

239Pu    24065       alpha   5.24        0.739
                             5.23        0.152
                             5.19        0.107

240Pu    6537        alpha   5.26        0.734
                             5.21        0.265
                             5.11        0.0009

241Pu    14.4        beta    5.24 10-3   1.0
                             (average)

241Am    432.2       alpha   5.58        0.852
                             5.54        0.128
                             5.48        0.014
                             5.64        0.003
                             5.61        0.002
---------------------------------------------------
a/  Principal transitions.

B.   SOURCES

1.  Nuclear explosions

402.  The major source which has introduced plutonium into the 
environment has been atmospheric nuclear testing.  Of nearly 1.5 
1016 Bq of 239, 240Pu activity released, about 1.2 1016 Bq has been 
dispersed and deposited world-wide [H2]. The remainder has been 
deposited locally at the sites of the tests.  The amount deposited 
in the northern hemisphere, where most of the tests were conducted, 
was three times that deposited in the southern hemisphere. 

403.  The amounts of the globally dispersed plutonium isotopes 
produced in all nuclear tests are listed in Table IX.2.  Of the 
total mass of plutonium released world-wide (4 106 g), 96% is 
comprised of 239Pu and 240Pu.  These two isotopes are not 
separately distinguished in alpha spectrometry and the combined 

amounts are usually reported.  The production data and also 
analysis of environmental samples indicate that of 239, 240Pu total 
amounts, 85% by mass or 60% by activity is due to 239Pu. 

Table IX.2  Production of globally dispersed plutonium 
isotopes in atmospheric nuclear testing
---------------------------------------------------------
                     Specific      Amount produced
Isotope  Half-life   activity   -------------------------
         (a)         (Bq/g)     Activity (Bq)   Mass (t)
---------------------------------------------------------
238Pu    87.74       6.3 1011   3.3 1014        0.00051

239Pu    24065       2.3 109    7.4 1015        3.26

240Pu    6537        8.4 109    5.2 1015        0.59

241Pu    14.4        3.8 1012   1.7 1017        0.041

242Pu    376000      1.5 108    1.6 1013        0.11
---------------------------------------------------------

404.  Most of the alpha activity of plutonium produced in nuclear 
explosions is due to 239, 240Pu.  241Pu is a beta-emitter which 
decays to 241Am.  Little 241Am was produced directly in the tests, 
but the activity amounts are accumulating in the environment as 
241Pu decays.  The activity of fallout 241Am in soil is currently 
about 25% of that of 239, 240Pu.  Decay of 241Pu this far and 
subsequent decay will result in a total production of 241Am from 
nuclear tests of 5.5 1015 Bq [B4]. 

2.   Nuclear fuel cycle

405.  There are about 240 nuclear reactors used to generate 
electric power throughout the world.  Upon fuel discharge, for each 
MW(e)a of electricity produced, it can be calculated that 1.2 1012 
Bq of 239, 240Pu are produced, 4.0 1012 Bq of 238Pu, 2.1 1014 Bq of 
241Pu and 2.8 1011 Bq of 241Am [H4, K1].  Generally more than half 
of the 239Pu produced during reactor operation undergoes fission, 
thus contributing to the energy produced by the reactor.  Routine 
operation of the reactors probably has not resulted in the release 
to the environment, world-wide, of more than trace amounts of 
plutonium and americium. 

406.  Discharge of plutonium to rivers and oceans from fuel 
reprocessing plants can be much more significant.  It has been 
estimated that 0.1-1% of the plutonium throughput is released in 
liquid effluent from the Windscale plant in the U.K. [M1]. Little 
plutonium is released to the air from reprocessing plants.  For 
example, it has been estimated that future reprocessing activities 
may result in the airborne release of about 40 to 4000 Bq of 
plutonium and 4 to 40 Bq of americium per MW(e)a [K1]. 


3.   Other sources

407.  Processes involved in the production of nuclear weapons have 
resulted in the release of plutonium to the surrounding 
environment.  Releases occur from fabrication and particularly from 
reprocessing plants.  Leakage from an oil storage area at Rocky 
Flats plant in Colorado resulted in release of about 2 1011 Bq of 
239Pu, half of it offsite but in the near vicinity of the plant.  
At Mound Laboratory in Miamisburg, Ohio, 4 1011 Bq of 238Pu were 
washed into an abandoned canal and another 2 1010 Bq was estimated 
to have been released to the air.  Plutonium has been released to 
onsite disposal areas at several laboratories. 

408.  A few accidents involving nuclear weapons have been reported 
that introduced plutonium into the environment.  The crash of an 
airplane at Thule, Greenland resulted in about 9 1011 Bq of 
plutonium being deposited on the shore and in the bottom sediments.  
A much smaller quantity is believed to have been carried by winds 
away from the accident site [F1]. The collision of two military 
planes resulted in plutonium from two weapons being dispersed at 
Palomares, Spain.  Much soil was removed in an attempt to clean up 
the plutonium. Three aberrant missiles were deliberately destroyed 
in flight and another burned on the launch pad at Johnson Island in 
the Pacific [F1].  Although the launch pad was decontaminated, 
undoubtedly several kilograms of plutonium were dispersed to the 
ocean from the three accidents.  It is estimated that from 4 1011 
to 40 1011 Bq of plutonium remain available to be incorporated into 
biological systems from these accidents [F1]. 

409.  The use of plutonium in thermoelectric generation systems of 
spacecraft has resulted in a relatively small amount of 238Pu and 
239Pu being introduced into the environment.  Of nearly 350 1014 Bq 
of 238Pu with an accompanying 260 1011 Bq of 239Pu carried into 
space by 19 U.S. spacecraft, 6.3 1014 Bq of 238Pu and 4.8 1011 Bq of 
239Pu were dispersed into the environment when one spacecraft re-
entered the atmosphere and burned over the Indian Ocean. Nearly 80% 
of this was dispersed in the stratosphere of the southern 
hemisphere and about 20% in the stratosphere of the northern 
hemisphere [H1, P5].  Another 16.5 1014 Bq of 238Pu and 12 1011 Bq 
of 239Pu entered, with containers intact into the Pacific Ocean as 
a result of an aborted flight.  In another aborted flight, the 
plutonium source was recovered intact from the ocean floor [D4]. 

410.  Large usage is now being made of smoke detectors containing 
241Am as an ionization source, with an average of 1 105 Bq of 241Am 
in each detector. About 28 1011 Bq of 241Am has been distributed 
throughout the U.S.A. in this form [U2]. Since the useful life of 
these detectors is estimated to be 10 years, it can be assumed that 
some of these detectors have already been disposed of in sanitary 
landfills, by incineration, and by other means. 

411.  Other uses of transuranium elements are made in consumer or 
medical devices, notably 238Pu in heart pacemakers.  These are 
sealed sources, which when handled properly in normal 
circumstances, should not allow the contents to be released to the 
environment. 

C.  BEHAVIOUR IN THE ENVIRONMENT

1.   Movement in soil

412.  When plutonium enters soil as fallout, or is added in 
solutions containing hydrolyzable Pu(IV), it is usually highly 
insoluble, regardless of soil type [W4]. Diffusion coefficients for 
surface soils [G1] are universally low (approximately 10-7 cm2 
s-1).  Therefore, the major inventory normally remains in the top cm 
of undisturbed soils, even with considerable water percolation 
[E6].  The relative immobility of plutonium in soils under these 
circumstances may be attributed to the initial low solubility of 
fallout particles and interaction of Pu(IV) hydrolysis products 
with soil, mineral and organic surfaces [W4]. 

413.  A small fraction (< 0.1%) of plutonium in soils is soluble, 
accounting for limited plant uptake from soil and chemical mobility 
under certain conditions in subsoils.  This may be due to the 
presence of complexing agents or valence states less subject than 
Pu(IV) to hydrolysis and insolubilization.  When Pu(IV) is added to 
soil as a synthetic or natural organic complex, solubility is 
initially greatly increased (several orders of magnitude) because 
of reduced hydrolysis;  subsequent mobility is a function of 
complex stability, competition with other ligands, and resistance 
of the ligand to chemical and microbial degradation [W6]. Empirical 
evidence suggests the presence of Pu(III) under reduced conditions 
in ground waters, and Pu(V)-Pu(VI) have been identified in certain 
natural waters [W4, B6]. 

414.  Physical processes also account for some vertical movement in 
soil.  Cultivation results in redistribution within the plow layer 
(to 30 cm) and longterm field studies have traced plutonium 
migration to 30 cm in undisturbed arid soil [N4].  In the latter 
case, the increased mobility over that predicted by diffusion alone 
has been attributed to biological transport and particle movement. 

415.  Since most plutonium is strongly absorbed on surface soils, 
wind and water erosion become primary environmental transport 
mechanisms [W4].  Transport distance will generally be a function 
of the size of the particle with which plutonium is associated.  
Particles in the fine silt-clay size range are the most likely to 
contain the highest concentrations of plutonium, to be transported 
the greatest distance by wind and water, and to remain attached to 
biological surfaces. 

416.  Detailed investigations of the behaviour of americium in 
soils are lacking.  In contrast to plutonium, disproportionation 
does not occur readily, and Am(III) would be the expected stable 
species [W4].  Hydrolysis reactions also influence the behaviour of 
Am(III) in soil, but the products of Am(III) hydrolysis are more 
soluble than those of Pu(IV) [R3]. 

2.  Transfer to plants

417.  Principal mechanisms of plutonium and americium transport to 
vegetation are foliar interception and root uptake.  Foliar uptake 

is dependent upon chemical form and size of the particle 
intercepted, residence time and weathering reactions of the leaf.  
Translocation to the seeds and roots after deposition on the leaf 
of soyabeans is approximately 10-5 of intercepted amounts [W4].  
The primary mode of entry into plants is root uptake, with reported 
soil-to-plant concentration ratios for plutonium ranging from 10-3 
to 10-8 [E5].  Increasing evidence suggests that the solubility in 
soil, rather than discrimination at the plant root level, is the 
limiting factor in plutonium uptake by plants [W6]. 

418.  Evidence suggests that plutonium is transported across the 
root as Pu(IV) [W6, D3].  Complexed Pu(IV) is probably the major 
translocated species in plants, and several anionic and cationic 
complexes of Pu(IV) have been determined in the xylem of plants 
supplied with Pu(IV) [B3].  Plutonium is not uniformly distributed 
in the plant.  The plutonium concentration decreases up the stem of 
soyabeans, and lowest plutonium concentrations occur in the seeds 
of barley and soyabeans [W6].  Systematic investigations of 
americium translocation and deposition in the plant after root 
uptake have not been conducted. 

3. Transfer to animals

419.  The primary sources of plutonium and americium to domestic 
animals are inhalation and consumption of plant tissues containing 
plutonium in surface-absorbed particles or in tissues.  In grazing 
herbivores, plutonium is primarily associated with the gastro-
intestinal tract and pelt, and to a lesser degree, with the lungs 
[W4].  Gastro-intestinal absorption requires the presence of 
soluble plutonium and hydrolysis/complexation reactions are likely 
to govern solubility.  The reducing potential of the gut appears 
sufficient to maintain principally the Pu(IV) state, which is 
subject to insolubilization by hydrolysis.  The effect is more 
pronounced in the presence of additional reducing substances, such 
as food residue [S8].  The fraction of ingested amount absorbed and 
deposited in the bone and liver is approximately 10-4 [W6].  
However, administration of Pu(VI) in solution to starved animals, 
or Pu(IV) in complexed forms (synthetic chelates or plant tissues) 
may increase gut absorption [W6, W5, B2]. 

420.  The gastro-intestinal absorption of americium from gavaged 
solutions is slightly greater than that of plutonium perhaps 
reflecting the reduced tendency of americium for hydrolysis.  
Information on absorption of americium incorporated in plant 
tissues is not yet available. 

4.  Transfer to diet

421.  The transfer of plutonium and americium to diet from fallout 
has not been as extensively studied as the transfer of 90Sr and 
137Cs.  A complete diet sampling, conducted annually in New York, 
was analysed for 239, 240Pu in 1972 [B5] and 1974 [B4].  The 1974 
samples were also analysed for 241Am.  A few samples of selected 
foods from 1963 and 1964 were also analysed for 239, 240Pu. 

422.  The highest concentrations of 239, 240Pu and 241Am were found 
in shellfish, followed by grain products and fresh fruits and 
vegetables.  The lowest concentrations were in meat, milk, eggs, 
fresh fish and in processed foods.  The values indicate that 
external contamination is a factor in the occurrence of plutonium 
in foods.  For the shellfish sample, comprising clams and shrimp, 
most of the plutonium and americium were found in the clams.  The 
muscle in the fresh fish sample, comprising halibut, snapper and 
flounder, had a 239, 240Pu concentration 10 times less and a 241Am 
concentration 50 times less than the shellfish sample. 

423.  Based on the New York sampling, the intake by ingestion 
during 1974 was estimated to be 60 mBq a-1 for 239, 240Pu and 16 
mBq a-1 for 241Am. The ratio of americium to plutonium was 0.27 in 
the total diet indicating little increase of americium relative to 
plutonium compared to the americium and plutonium in the soil. 

424.  A 239, 240Pu dietary intake record has been calculated based 
upon the annual fallout deposition rate and the cumulative deposit 
in soil and compared with the New York food sample surveys 
conducted in 1963, 1964, 1972 and 1974 [B4]. Assuming no further 
atmospheric injections, the 239, 240Pu dietary intake will remain 
relatively constant at 0.03 Bq a-1 owing to uptake from the 81 Bq 
m-2 cumulative deposit in soil.  For 241Am the calculation 
indicates that the uptake of 241Am from the cumulative deposit in 
soil is a factor of two greater than plutonium.  The estimated 
241Am dietary intake continues to increase as the cumulative 
deposit in soil increases owing to ingrowth from 241Pu decay.  When 
the cumulative deposit reaches its projected maximum of 29 Bq m-2, 
the dietary intake will also be at a maximum of 0.03 Bq a-1. 

425.  The cumulative transfer of plutonium and americium to diet 
depends very much on the assumed residence times in soil.  These 
times are no doubt shorter than the radioactive mean lives due to 
leaching and fixation in soil.  Extremes of transfer estimates are 
obtained by taking the mean residence times to be 50 years in one 
case and the radioactive mean lives in the other case.  The 
geometric mean of these extremes then gives a tentative estimate of 
the transfer factor P23 from deposition density to diet.  The 
results are 0.6, 0.3 and 0.2 Bq per Bq m-2 for 239Pu, 240Pu and 
241Am, respectively.  For 238Pu and 241Pu the estimates of P23 are 
based on the radioactive mean lives.  Most of the transfer is 
attributed to direct deposition [B4].  The values are 0.08 and 0.04 
Bq per Bq m-2 for 238Pu and 241Pu, respectively. 

5.  Aquatic behaviour

426.  Plutonium is mobilized off watersheds to rivers and coastal 
waters [H1, M5, H3, S3, S2].  Estimates available for plutonium 
indicate input ranges from 0.05% per year for heavily cultivated 
watersheds [M5, S2] to 0.005% per year for a heavily forested 
watershed, indicating a residence time of 103 to 2 104 years [W4]. 

427.  Environmental studies have shown that in a variety of 
comparatively shallow bodies of water, both fresh water and marine, 
more than 96% of the total plutonium released to or deposited on 

these environments is rapidly transferred to sediment [E1, L2, H5, 
H6, P6, S1, N2, H3, E4, P1, W3, S2]. However, in the deep oceans 
there is only slow transfer of the total plutonium in the ocean 
water column to deposited sediments.  It is estimated that this may 
represent about 30% by 1980 of deposited fallout plutonium [B8]. 

(a)    Freshwater systems

428.  The behaviour of plutonium and americium has been studied in 
a wide range of fresh water systems [S2, E3, D1, R1, B3, W3, B7].  
The concentrations of plutonium in the water of these systems 
varied by more than four orders of magnitude [W4].  Higher 
concentrations of plutonium have been observed in lakes with low 
pH, lakes with high sulfate concentrations and other acidic lakes 
[W1].  Chemical analyses indicate that while the plutonium in Lake 
Michigan in the U.S.A. was predominantly in the Pu(V) and (VI) 
states, in all other lakes studied Pu(III) and (IV) predominated 
[N1, W2].  The results strongly suggest that the solubility of 
plutonium is governed by different complexing agents.  In waters of 
high pH, the concentration of CO3-2  and HCO3- is relatively high, 
and carbonate complexes can form.  In waters of low pH, such 
complexes cannot exist, and the solubility must be due to 
complexing with other ligands, such as natural organic compounds. 

429.  A relationship has been shown between the concentration of 
plutonium in water and the concentration in sediments or 
particulate matter [N1].  Values for the distribution constant, KD, 
vary between 104 and 5 105, with most values not varying more than 
fivefold.  Considering the wide variety of the systems, including 
sediment types, size, source terms, etc., this small range in 
values suggests a commonality in the behaviour of plutonium in 
these systems.  There is some evidence that the plutonium absorbed 
by sediment particles is predominantly in the (III) and (IV) 
states, yet on re-equilibration of sediment with water, it has been 
shown that there is a conversion of Pu(III) and Pu(IV) back to 
Pu(V) or (VI).  This strongly supports the hypothesis that the 
concentration of plutonium in many fresh water lakes and rivers is 
controlled by an equilibrium between water and sediment [B7]. 

(b)    Marine systems

430.  The behaviour and fate of transuranic elements in the marine 
environment were given very little attention before the mid-1960s.  
By far the greatest effort for the next decade was applied to 
determine the residence time of plutonium in the oceans [B8, M4].  
Comparison with 90Sr and 137Cs indicates that the residence time of 
plutonium in the water column is less than that of both 90Sr and 
137Cs.  The observed distribution has been explained in terms of a 
distribution of particles settling at various velocities [N2]. 

431.  In the Irish Sea, the concentration of dissolved 239, 240Pu 
is only 6% as large as that of 137Cs, normalized to a unit 
discharge rate;  the value for 241Am is even lower at 3%.  These 
values suggest that plutonium and americium leave the water phase 
very rapidly.  Measurements in the Irish Sea indicate that 

plutonium is in solution predominantly as Pu(VI) and on particles 
as Pu(III) and Pu(IV) [Nl], a situation similar to that which 
exists in the Great Lakes in the U.S.A. 

432.  Distribution coefficients, KD, between water and suspended 
sediments for plutonium in the oceans are similar to those in the 
Great Lakes and do not appear to be source related.  Similar KD 
values (104 to 105) for sediments from the Irish Sea and Enewetak 
Lagoon suggest that similar chemical reactions are occurring.  In 
the Irish Sea, the overall order of KD values for transuranic 
nuclides is 241Am > 242Cm and 244Cm > 239, 240Pu [P2]. 

(c)    Bioaccumulation

433.  Trophic-level studies in freshwater and marine environments 
indicate that plutonium concentration factors for organisms 
relative to water generally decrease at higher trophic levels [H5, 
H6, B7, D2, E2, P4].  Typical values for plutonium in the edible 
portions used for assessment purposes are 10 for fish, 100 for 
crustacea and 1000 for molluscs and algae [I1, N3].  Whole 
organisms values may be 10 to 50 times higher depending upon the 
degree of contamination by sediment.  Limited field data indicate 
increased concentration factors for 241Am over plutonium in lower 
trophic levels and in fish [P3, W1, P4]. 

434.  Laboratory studies indicate that marine teleost fish can 
absorb Pu(VI) by direct uptake from sea-water with limited 
absorption across the gut from labelled food or sediment. Marine 
elasmobranchs on the other hand appear to absorb plutonium across 
the gut relatively easily [P3].  Crustacea, such as crabs, have 
high assimilation efficiency for plutonium when fed labelled food 
[F3, G2], and some biomagnification in a simple laboratory 
invertebrate food chain has been observed [F2]. 

D.  TRANSFER TO MAN

435.  Information on the transfer of plutonium to man is available 
from autopsy measurements on persons exposed to plutonium from 
weapons test fallout, from occupational sources, and from 
intentionally administered plutonium in terminally ill patients.  
The data from fallout plutonium are most pertinent to general 
environmental considerations.  In addition, the results from many 
animal studies provide supporting data on the probable magnitude of 
the biokinetic parameters that determine this transfer. 

436.  Extensive data on the fallout plutonium content of persons in 
the general population have been published, including data from 
several areas in the U.S.A. [M2, W7], from Finland [M3] and Japan 
[O1].  Although great variability was noted from sample to sample, 
particularly where only small quantities of tissue were available, 
the average results agreed reasonably well with computed tissue 
plutonium burdens [B5], based on estimated plutonium intake by 
inhalation (ingestion was shown to be insignificant relative to 
inhalation), assuming metabolic parameters as employed by ICRP. 

437.  Considering all of the human data now available, there would 
seem to be no reason to alter the ICRP assumption of an equal 
distribution of systemic plutonium between skeleton and liver (45% 
in each) [I4].  The deposition and retention of fallout plutonium 
in the lung seems also to be well described by the ICRP lung model, 
assuming behaviour as a Class Y compound [I5].  The ICRP lung model 
appears to overestimate substantially the transfer of fallout 
plutonium to tracheobronchial lymph nodes.  From extensive studies 
in experimental animals it is known that the extent of 
translocation from lung to lymph nodes varies widely with the 
chemical and physical form of the particles inhaled [B1].  The 
human data also suggest that gonadal deposition may be slightly 
higher than the fraction of 10-5 g-1 assumed by ICRP [I2].  More 
precise analytical data are required, however, to support such a 
conclusion.  Studies of plutonium deposition in the gonads of 
several species of experimental animals support the ICRP assumption 
[R2]. 

438.  Based on comparative studies in experimental animals, ICRP 
has assumed that inhaled americium will behave in humans in a 
manner identical with plutonium [I4], except that all americium 
compounds are assumed to follow Class W lung model kinetics [I2];  
i.e., americium is more rapidly cleared from the lung and more 
efficiently translocated to bone and liver than is plutonium oxide.  
Applying these assumptions to the estimated intake by inhalation of 
fallout americium and 241Pu (which will decay to 241Am), estimates 
have been made of human americium burdens [B4].  These indicate a 
241Am/239, 240Pu ratio of 0.24 in l978, which will increase to 0.38 
by the year 2000 because of further decay of deposited 241Pu.  
Pooled samples from 18 autopsies done from 1970 to 1974 showed a 
measured 241Am/239, 240Pu ratio, in vertebrae, of 0.22 [B4]. The 
close agreement between measured and calculated ratios lends 
support to the ICRP assumptions. 

439.  Data on the gastro-intestinal absorption of plutonium and 
americium are available only from studies in experimental animals.  
Such studies were summarized by an ICRP task group in 1972, which 
led to assumed values for the fraction absorbed of 10-6 for 
plutonium oxide and 3 10-5 for other commonly occurring compounds 
of plutonium [I4].  It was recognized that a much higher absorption 
might be expected for complexed forms of plutonium.  More recent 
data obtained in a variety of animal species [S4, S6] resulted in a 
modification of ICRP estimates to 10-5 for oxides and hydroxides of 
plutonium, 10-4 for other commonly occurring plutonium compounds, 
and 5 10-4 for all compounds of americium [I2].  Qualitative 
support for the plutonium numbers is provided by autopsy data from 
a group of reindeer-herding northern Finns who ingest large 
quantities of plutonium-rich reindeer liver [M3]. Their plutonium 
burdens seem to be no higher than those of southern Finns, who do 
not ingest these relatively large quantities of plutonium.  A 
difference should have been apparent if the fraction absorbed from 
the gastro-intestinal tract had been much greater than 10-4. 

440.  There is evidence from animal studies that plutonium 
incorporated into alfalfa [S7] or liver [S4] may be absorbed to a 
greater extent than inorganic plutonium; the effect is not large, 

however, and is reversed in the case of americium [S4]. Concern has 
been expressed that the gastro-intestinal absorption of plutonium 
in the hexavalent state, such as may be produced by chlorination of 
water supplies, may be markedly increased as compared to 
tetravalent plutonium [L1].  It has since been shown, however, that 
under normal conditions in the gastro-intestinal tract of 
experimental animals, no such increase of gastro-intestinal 
absorption is observed [S8]. 

441.  A marked increase in the gastro-intestinal absorption of 
plutonium and other actinides has been reported in neonatal animals 
of several species [I4, S5].  This increase may be as much as 
several hundredfold in the case of rats and several thousandfold in 
the case of miniature swine.  In addition to increased absorption, 
there is a prolonged retention of the actinide within the mucosa of 
a small intestine [S5].   It must be assumed that the human infant 
will also show an increased absorption, although the magnitude of 
this increase and its duration is unknown. 

442.  In addition to the inhalation and ingestion routes, actinides 
may under unusual circumstances of occupational exposure, enter the 
body by absorption through the intact or punctured skin [I4, B1].  
Normally, however, intact skin is an effective barrier to plutonium 
entry, and this route of entry should not be of concern for general 
environmental exposure. 

443.  Once deposited systemically, plutonium is tenaciously 
retained.  This fact is qualitatively evident from extensive data 
on the excretion of plutonium by occupationally exposed humans 
[V1].  It has also been quantitatively evaluated in a few human 
cases  and in a variety of experimental animals [I4, D5].  Based on 
these data, ICRP has employed a biological half-time of 100 years 
for plutonium in the skeleton, and a half-time of 40 years for 
plutonium in liver; plutonium in gonads is assumed to be retained 
without loss [I4, I2].  The same parameters are assumed to apply in 
the case of americium. 

E.  DOSIMETRY

1.   Dose per unit intake

444.  The doses to the various tissues following inhalation or 
ingestion or plutonium and americium are determined using the 
parameters and models suggested by ICRP.  A variety of estimates 
are possible, depending on particle size of inhaled particles and 
the solubility class of both inhaled and ingested material.  The 
values given below are for the representative 1 µm aerosol size and 
for the insoluble oxide or hydroxide forms of plutonium. 

445.  The ICRP lung model divides the respiratory system into three 
compartments. Deposition fractions in each region for the 1 µm 
particle size are 29% in the nasopharyngeal region, 8% in the 
tracheo-bronchial region and 23% in the pulmonary region.  Smaller 
particles have a progressively greater deposition fraction in the 
pulmonary region and less retention in the nasopharyngeal region. 

446.  Inhaled, insoluble plutonium particles are assigned Class Y 
parameters, retention in the lungs of the order of years (500 d 
half-time for 60% of the pulmonary deposition). Because of greater 
mobility of americium, all of its compounds are assigned Class W 
parameters, with retention in the lungs of the order of weeks (50 d 
half-time for 60% of the pulmonary deposition).  In both cases, 40% 
of the pulmonary deposition is cleared with a half-time of 1 d by 
mucocilliary action through the tracheo-bronchial region.  
Translocation from the naso-pharyngeal and tracheo-bronchial region 
is rapid, within one day, with most of the material swallowed and 
small amounts absorbed to blood. 

447.  For both inhaled and ingested material reaching blood, 
fractional transfer to blood and liver is assumed to be 0.45 each 
and 3.5 10-4 to gonads (testes) for both plutonium and americium.  
Uptake to blood following ingestion of insoluble plutonium is 10-5 
and 5 10-4 for all compounds of americium. 
Table IX.3  Absorbed dose equivalent commitments per unit
intake of plutonium and americium (Sv per Bq)
---------------------------------------------------------------------------
Isotope     Lung       Liver     Gonads     Red bone   Bone       Effective
                                            marrow     lining     a/
                                                       cells
---------------------------------------------------------------------------
 Inhalation

238Pu      3.2 10-4   1.8 10-4   b/         6.6 10-5   8.3 10-4   8.2 10-5
                                     
239Pu      3.2 10-4   2.1 10-4   b/         7.6 10-5   9.5 10-4   8.9 10-5
                                     
240Pu      3.2 10-4   2.1 10-4   b/         7.6 10-5   9.5 10-4   8.9 10-5

241Pu      3.2 10-6   4.4 10-6   2.8 10-7   1.7 10-6   2.1 10-5   1.6 10-6

242Pu      b/         5.5 10-4   3.2 10-5   2.0 10-4   2.5 10-3   1.4 10-4
               
 Ingestion      
               
238Pu      b/         4.0 10-8   2.3 10-9   1.5 10-8   1.8 10-7   1.5 10-8
               
239Pu      b/         4.4 10-8   2.6 10-9   1.6 10-8   2.1 10-7   1.6 10-8
               
240Pu      b/         4.4 10-8   2.6 10-9   1.6 10-8   2.1 10-7   1.6 10-8
               
241Pu      b/         8.6 10-10  5.7 10-11  3.4 10-10  4.2 10-9   2.5 10-10
               
242Pu      b/         2.3 10-6   1.4 10-7   8.4 10-7   1.1 10-5   5.9 10-7
---------------------------------------------------------------------------
a/  Effective dose equivalent commitment per unit intake (Sv per Bq).
b/  Doses which contribute < 10% to the effective dose equivalent commitment.
    Note:  for plutonium - Class Y and f1 = 10-5
           for americium - Class W and f1 = 3 10-4.
    Reference [I3].

448.  The estimates of the absorbed dose equivalents per unit 
intake are given in Table IX.3.  These are computed over a 50 year 
period following intake.  They correspond to the transfer factors 
P25 and P35 relating intake in air and diet, respectively, to 
tissue dose.  For these dose equivalent values, the quality factor 
of 20 has been used for the alpha-emitters and 1 for the beta-
emitters (241Pu). 

2.  Dose per unit release

(a)    Nuclear explosions

449.  The dose equivalent commitments for plutonium and
americium released in nuclear explosions can be assessed using
the expressions

        Dc  =  P25 Ia      (inhalation)

        Dc  =  P23 P35 F   (ingestion)

where Ia is the cumulative intake from air (the integrated 
concentration in air (Bq a m-3) times the breathing rate (22 m3 
d-1) and F is the integrated deposition density.  The values of 
the transfer factors were discussed above. 

450.  For the past pattern of nuclear tests the population-weighted 
global integrated concentration in air and deposition density can 
be determined from comparisons with 90Sr [U1], from Pu-90Sr 
production ratios, and from 241Pu decay considerations [B4].  The 
results are for air:  1.5, 37, 25, 830 and 1.7 µBq a m-3 and for 
deposition density:  0.85, 21, 14, 470 and 16 Bq m-2 for 238Pu, 
239Pu, 240Pu, 241Pu and 241Am, respectively.  With this 
information, the dose equivalent commitments can be determined by 
using the above expressions, and the values per unit activity 
released can be obtained by dividing the results by the estimated 
input amounts from nuclear testing (Table IX.2). 

451.  Finally, the collective dose equivalent commitments are 
determined by multiplying by the appropriate global population.  
For inhalation, the activity in air has by now been nearly 
depleted, so the present global population applies (4 109 persons).  
For ingestion of the short half-life 241Pu, the present population 
applies and for the other isotopes, the long-term transfer can be 
assumed to apply to the projected equilibrium value of 1010 
persons.  The results for the collective effective dose equivalent 
commitment per unit release are included in the summary Table IX.4. 

(b)    Nuclear installation

452.  The contribution of the inhalation pathway to the collective 
dose commitments for plutonium and americium in airborne effluents 
from nuclear installations can be estimated from the integrated 

concentrations in air in the dispersion region.  It has previously 
been shown that the appropriate formula is 

     c   I deltaN P25
    S1 = ------------
            vd

where I is the air breathing rate (22 m3 d-1), deltaN is the 
population density in the region (25 man km-2), P25 is the dose per 
unit intake factors (Table IX.3) and vd is the deposition velocity 
(0.5 cm s-1).  The integrated concentration in air is determined by 
the amount released (unit activity) per unit area of the deposition 
region divided by the deposition velocity.  The areal dependence is 
removed by multiplying by the population of the region (times the 
area).  The results from evaluating this expression are summarized 
in Table IX.4. 

Table IX.4  Summary of collective effective dose equivalent 
commitments per unit activity released of plutonium and 
americium (10-14 man Sv per Bq)
---------------------------------------------------------------
                        238Pu   239Pu   240Pu   241Pu   241Am
---------------------------------------------------------------
 Nuclear explosions

Inhalation              1000    1000    1000    30      5 a/
Ingestion               3       30      10      0.01    300

 Nuclear installations
Release to air b/
  Inhalation            10000   10000   10000   200     20000
  Ingestion             3       20      10      0.03    300

 Release to fresh water
  Drinking water        20      20      20      0.2     900
  Fish                  5       5       5       0.06    200

 Release to salt water
  Fish                  0.01    0.01    0.01    0.0001  0.4
  Shellfish             0.2     0.2     0.2     0.002   6
---------------------------------------------------------------
a/  Per Bq of 241Pu released.
b/  Assumes population density of 25 man km-2.

453.  The contribution of the ingestion pathway from airborne 
effluents to the collective dose commitment per unit activity 
released, Sc1, can be determined by using the expression 

           c
          S1 = P23  P35  SN

Using the values of the transfer factors given previously and a 
population density, deltaN, of 25 man km-2 in the region of 
deposition, the values summarized in Table IX.4 are obtained. 

454.  For the aquatic ingestion pathways the generalized UNSCEAR 
model is utilized [U1] 

     c   Nk  Ik  fk  P35
    S1 = -----------------
         V(lambda + 1/tau)

to be evaluated for each pathway k.  The quotient of water 
receiving volume, V, and the population involved, Nk, is the water 
utilization factor [U1], assumed to be a constant for each pathway.  
A summary of the values used in the assessments by UNSCEAR is given 
in the following listing: 

    Parameter                     fresh water    sea water
1.  tau, turnover time of         10 a           1 a
    receiving water
2.  Correction factor for         1.0            1.0
    sediment removal
3.  V, water utilization factor   3 107 1/man    3 109 1/man
    N
4.  fk, concentration factor
    for item k
        drinking water            0.1
        fish                      10             3
        shellfish                                300
5.  Ik, consumption rate for
    item k
        drinking water            440 1/a
        fish                      1 kg/a         6 kg/a
        shellfish                                1 kg/a

455.  Due to insufficient data, the values of the above listing 
will be assumed to apply to all isotopes of plutonium and 
americium.  The values of the dose factor, P35, are given in Table 
IX.3.  A summary of the evaluated results is given in Table IX.4. 

456.  Assessments using the above model only account for that 
portion of the dose given during the mean residence time of the 
water in the receiving area.  These are essentially complete 
collective dose commitments for the release to fresh water and for 
the shorter-lived isotopes (241Pu).  The removal of the longer-
lived isotopes to the sediments of the deep ocean will largely 
preclude any further contributions to the dose estimates. 

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X.   RADIATION EFFECTS

457.  The interaction of radiation with matter results in the 
liberation of energy carried by the alpha-particle, beta-particle, 
gamma or x-rays and the ionization or excitation of the irradiated 
material.  In biological material, damage may be caused directly to 
cell components by the radiation interactions or indirectly by the 
actions of free radicals, the charged fragments of ionization 
events. The damage may result in cell death or in cell 
transformations, which at some later time may cause harmful effects 
in the irradiated individual or in his offspring. The amount of 
damage depends on the amount of radiation, which may be from 
external irradiation or from radionuclides within the body, on the 
type of radiation and on the sensitivity of the tissue. 

458.  Radiation effects in man are usually classified as somatic 
and genetic or hereditary, according to whether they affect somatic 
or germinal cells.  Somatic damage is expressed therefore by 
definition within the lifetime of the irradiated individual, while 
genetic damage is expressed at some stage in his progeny.  Somatic 
effects are - somewhat loosely - further distinguished as immediate 
or late depending on the time of their appearance. 

A.   SOMATIC EFFECTS

1.  Early somatic effects

459.  The immediate or early somatic effects are expressed in man 
within a few days or a few weeks after exposure as a result of 
damage to one or more of the self-renewing tissues. These effects 
are also called functional because they are due to the inactivation 
of a great number of functional cells of a given differentiative 
line. Selective irradiation of a given tissue (as in the case of 
exposure to some internal emitters) usually leads to effects 
localized in that tissue;  whole-body irradiation above a given 
dose results, on the contrary, in the appearance of generalized 
effects, usually under the form of a specific syndrome which 
depends on dose.  The clinical severity of the immediate effects 
changes considerably with the dose, dose rate, type and energy of 
the absorbed radiation and part of the body irradiated. 

460.  The nature of immediate effects is non-stochastic or 
deterministic, in the sense that these effects are expected to 
occur in an exposed individual absorbing doses in excess of a 
minimum amount of radiation referred to as the threshold dose.  The 
threshold dose is extremely variable depending on the effect and 
the tissue considered.  It is of the order of a few tenths of a Gy 
for most functional effects of importance for radiation protection.  
The existence of a threshold of dose is an important characteristic 
of these effects;  it makes it virtually impossible for them to 
appear for doses below the threshold, thus allowing their complete 
avoidance. 

461.  Another important aspect of early effects to be emphasized in 
relation to irradiation from internal emitters is the variability 
of the dose threshold as a function of dose rate.  Low dose rate 
irradiation is normally a condition leading to an increase of the 
threshold.  Since irradiation from radionuclides (particularly at 
the levels usually present in the environment) takes place at very 
low doses and dose rates it is virtually impossible, except under 
exceptional conditions of emergency, that immediate effects might 
be observed.  They do not deserve therefore any more extended 
treatment. 

2.  Late somatic effects

462.  Late somatic effects are those appearing in irradiated 
individuals after a latency period and are expressed mainly in the 
form of leukaemias or solid tumours.  These effects are stochastic 
or statistical in nature, in so far as it is impossible to identify 
a causal relationship for them in any given case.  The correlation 
between radiation dose and induction of these conditions may only 
be shown on large populations of irradiated individuals as an 
increase of the above diseases over their apparently spontaneous 
background incidence. 

463.  Since it is impossible to establish with any certainty the 
shape of the dose-induction relationship for late tumorous effects, 
particularly at low doses and dose rates, it is usually assumed 
that the frequency of their occurrence is linear with dose and 
without threshold.  The clinical severity of these conditions is 
variable, but it is commonly assumed for purposes of radiation 
protection that they might be of a uniform and maximum severity, 
namely the death of an individual. 

B.   GENETIC EFFECTS

464.  Radiation-induced hereditary effects may appear in the 
progeny of irradiated individuals within the first generation 
following irradiation - in which case the damage is called dominant 
- or within subsequent generations when the genes that carry the 
same mutation in the male and in the female genetic complement 
happen to match in the genome of the zygote.  In this latter case 
the genetic damage is called recessive.  Clinically, radiation-
induced hereditary conditions have a large spectrum of severity 
from the relatively trivial to the very harmful. 

465.  As in the case of the somatic late effects, radiation 
protection usually makes reference to the most severe hereditary 
diseases which are either incompatible with life or very disabling 
for the individual.  Radiation protection also assumes that for 
this class of effects a linear non-threshold induction relationship 
with dose may apply. 

C.   DOSE-RESPONSE RELATIONSHIPS

466.  In experimental animals and in man late somatic and 
hereditary effects may exhibit different shapes of the dose-effect 
relationships, according to a large number of physical and 

biological variables operating in each particular system. Linear, 
linear-quadratic, quadratic or complex relationships have been 
described in various circumstances. No generalization may be gained 
by the consideration of all existing experience, except perhaps 
that each specific system responds according to different kinetics 
of action and that biologically complex effects usually correspond 
to more complex types of relationships.  It would be impossible to 
set up a rationale for a system of radiation protection by 
considering each case separately.  To overcome this difficulty the 
assumption is made that late somatic and hereditary effects of 
irradiation follow a non-threshold linear function of dose.  This 
assumption is simple and there is evidence that it is also a 
conservative assumption in most cases. 

467.  It is important to stress the meaning of the assumption of 
non-threshold linearity.  It postulates, on the one hand, that 
there is no dose, however small, that may in principle be 
considered safe and no dose increment, however small, which could 
not produce a corresponding increase of effect and therefore of 
risk.  The summation of doses taken as a measure of total risk and 
calculations of collective doses as expression of the total 
detriment in the exposed population have little meaning outside the 
assumption of non-threshold linearity. 

468.  On the other hand, adoption of the assumption of linearity 
involves implicitly the adoption of other important principles.  
Actually, if there is a linear relationship between the dose and 
the induction of stochastic effects, it becomes possible to use the 
average dose received by a given organ or tissue as a significant 
reference quantity.  Under these conditions it becomes unnecessary 
to consider the dose variability within the given organ or tissue 
because the response of the component cells (taken to be of uniform 
sensitivity) will in any case be linear with the absorbed dose.  
The reactions of the component cells will in any case sum-up to 
produce an overall effect corresponding to that expected from the 
mean dose in that organ or tissue. 

469.  In case of internal irradiation there may be problems related 
to the presence of point sources.  The following considerations 
apply in these cases.  Firstly, with regard to non-stochastic 
effects, cell death resulting from high doses within microscopic 
volumes of the tissue are expected to produce less harm than for 
the same dose uniformly distributed within large volumes:  this is 
because killing of transformed cells by high doses would be 
expected to lead to inactivation of potentially transformed cells.  
Moreover, loss of cells around the zones of highest dose absorption 
would not be expected to result in significant decrease in tissue 
or organ function, unless the functional reserve of the organ or 
tissue is impaired for other reasons. 

D.  RISK ESTIMATES

470.  UNSCEAR has extensively reviewed in its 1977 report most 
information on the subject of tumour induction in man by ionizing 
radiation.  The Committee concluded that the risk for all fatal 

malignancies after whole-body irradiation at low doses and dose 
rates of low-LET radiation - as an average of both sexes and all 
ages - is in the region of 10-2 Sv-1.  The risk of inducing non-
fatal malignancies under the same conditions would probably fall in 
the same range. 

471.  It should be emphasized that for doses of the order of those 
received annually from natural sources no direct information is 
available.  The above estimate is derived predominantly from 
observations conducted on people exposed to absorbed doses of over 
1 Gy.  While the rate per unit dose or doses in the region of the 
natural background would unlikely be higher, it could, however, be 
substantially lower.  There is no evidence that irradiation from 
internal sources would produce rates of tumour induction differing 
from those from external irradiation, if account is taken of the 
mean absorbed doses in tissues. 

472.  Concerning hereditary effects, the Committee estimated that 
when a population is continuously exposed to low doses of low-LET 
radiation at rates of the order of 10-2 Gy per generation (a 
generation corresponds to about 30 years) about 50 genetic diseases 
might be expected to occur per one million first generation 
progeny. At equilibrium, the total genetic damage expressed over 
all generations (or the value in each generation reached after 
prolonged continuous exposure) would be of the order of 150 cases 
per million progeny. 

473.  It is often convenient to be able to estimate the total 
detriment to the individuals from irradiation of specific organs or 
tissues, taking into account the various types of effects from 
various irradiations.  This could be a difficult procedure, 
however, in cases when different irradiation modalities lead to 
different effects.  Such a situation applies for irradiation by 
internally deposited radionuclides, when the nuclides produce 
different doses in tissues of varying sensitivity and the effects 
must be added to the effects of whole-body irradiation. 

474.  In order to overcome this difficulty, ICRP, in its 
publication 26, has designed a system allowing combined estimates 
of risk in various organs and tissues, based on their 
susceptibility to various effects.  These risks of effects of 
irradiations of specific tissues are weighted relative to the total 
effect from whole-body irradiation.  The risks apply to one 
individual or the whole exposed population, making use of the 
hypothesis of linearity.  It is realized, of course, that the 
applicability of risk estimates may vary according to the 
characteristics of the individual (genetic make-up, sex, age, etc.) 
or to the structure of the exposed population.  But it is also held 
that an acceptable level of precision may be reached by assuming an 
average risk value to be applied to all members of the population 
irrespective of the above mentioned variability. 

475.  The mortality risks and weighting factors recommended by ICRP 
are shown in Table X.1.  These factors have been derived for the 
protection of workers, but may also be applied to large population 

groups, provided allowance is made for the hereditary effects which 
would be expected to appear in all generations subsequent to the 
second. 

476.  It must be emphasized that the risk estimates for induction 
of somatic and genetic effects should be regarded as the best 
possible numerical conclusions to be drawn from a very 
heterogeneous data base affected by various types of dosimetric and 
epidemiological uncertainties.  Although it is felt that such 
estimates are reasonably precise for the purpose of radiation 
protection they are to be interpreted and used in a statistical 
sense simply as illustrations of the order of magnitude of 
potential risks.  The actual validity of these estimates could not 
possibly be tested empirically under normal circumstances owing to 
the small levels of the risks compared with the far higher 
"spontaneous" background of similar conditions in the general 
population. 

Table X.1  Mortality risk and weighting factors for 
different organs (from ICRP publication 26)
------------------------------------------------------------
Tissue                           Mortality risk   Weighting
                                 (Sv-1)           factor
------------------------------------------------------------
Breast                           0.25 10-2        0.15
Red bone marrow                  0.2  10-2        0.12
Lungs                            0.2  10-2        0.12
Thyroid                          0.05 10-2        0.03
Bone surface                     0.05 10-2        0.03
Remainder                        0.5  10-2        0.30
Gonads (hereditary effects in
 the first two generations)      0.4  10-2        0.25
------------------------------------------------------------
Total                            1.62 10-2        1.00
------------------------------------------------------------

XI.   CONCLUSIONS

A.   RADIONUCLIDES AND THE ENVIRONMENT

477.  Radiation is a natural feature of man's environment - from the 
high energy charged particles which make up cosmic radiation to the 
radioactive decay of radionuclides in the earth's crust and in the 
biosphere.  Several of the radionuclides considered in this 
document have important natural sources, including tritium, carbon-
14, krypton-85 and radon.  The first three are produced mainly by 
cosmic ray interactions in the atmosphere.  Radon arises from the 
decay of radium present in the earth's crust.  All of these 
radionuclides are widely dispersed in air.  Tritium and 14C enter 
more general bio-geochemical cycles, the hydrological cycle and the 
carbon cycle, respectively. 

478.  Several activities of man result in the production of 
radionuclides and contribute to the radiation background.  An 
important source has been the testing of nuclear weapons in the 
atmosphere.  Large scale atmospheric testing was completed prior to 
the Test Ban Treaty of 1963, but additional tests by some countries 
have continued.  The radionuclides produced in atmospheric nuclear 
explosions become widely dispersed in the atmosphere, primarily of 
the hemisphere in which the test was conducted, but with some 
interhemispheric mixing contribute to the global exposures to 
fallout radioactivity. Important fission radionuclides in terms of 
the doses delivered include 131I, 90Sr and 137Cs.  In the long-
term, the radionuclides 14C and isotopes of plutonium become 
important contributors to the radiation dose. 

479.  The generation of electricity using nuclear power reactors 
also results in the production of radionuclides and in some 
radiation exposure of man.  The radioactive materials are largely 
contained within the fuel elements in the reactor or in waste 
treatment systems at the fuel reprocessing plant. Releases of 
controlled amounts occur in liquid and airborne effluents from the 
nuclear installations.  Accidents could result in potentially 
greater releases of radioactive materials. 

480.  The natural and man-made sources of radionuclides in the 
environment have been discussed for the various radionuclides 
considered in this document.  The amounts of radioactive materials 
released from the various sources have been reviewed and the 
relevant values have been shown to depend very much on specific 
past practices. Even the normalized release amounts, for example 
the amounts per Mt in nuclear explosions or per MW(e)a of 
electricity generated, have limited validity.  The amounts released 
in atmospheric nuclear testing depend on the types of devices 
tested and on the geographic pattern of past testing.  Releases 
from nuclear power installations depend on the efficiency and 
integrity of present designs and on the specific waste management 
practices currently utilized. 

481.  The behaviour of radionuclides in the environment has been 
studied rather extensively, so that by now, the dispersion of these 
pollutants in the environment is fairly well understood.  Values 

have been derived for the various transfer factors which describe 
the transport of radionuclides in the environment and their 
transfer to man.  For the purpose of dose assessment it is 
necessary to consider only the more specific aspects of particular 
release situations and to adjust the more generally valid 
parameters to the local conditions. 

B.   DOSE ASSESSMENTS

482.  Exposure of man to radionuclides in the environment occurs by 
inhalation of amounts in air, ingestion of amounts incorporated 
into diet, or from external exposure to radionuclides in air or 
deposited on the ground.  The dose assessments for the 
radionuclides considered in this document have been directed toward 
obtaining estimates of the collective effective dose equivalent 
commitments per unit amount of activity of the radionuclide 
released.  This expression of dose gives the absorbed dose, 
weighted for radiation type and sensitivity of irradiated tissues 
to the entire population and for as long as the exposures from a 
specific release continue. This quantity is expected to be most 
directly related to the total health detriment which may result.  
The results of the dose assessments are summarized in Table XI.1. 

483.  The degree of transfer of the radionuclides to man and thus 
the dose estimates vary depending on the source of release of the 
radionuclides.  Natural production of tritium, 14C and 85Kr occurs 
primarily in the upper atmosphere, following which there is 
widespread dispersion, and for 3H and 14C generalized cycling 
throughout the environment.  Nuclear explosions conducted in the 
atmosphere, particularly large scale tests, result in injection of 
debris into the stratosphere, from where the radionuclides are 
globally dispersed.  Releases of radionuclides from nuclear 
installations are near surface emissions in airborne effluents or 
in liquid effluents to rivers and lakes or to the marine 
environment.  Exposures are primarily to the local and regional 
populations, although the longer-lived radionuclides may also 
become more widely dispersed. 

484.  Release of unit activity of a radionuclide to the environment 
generally results in the lowest collective effective dose 
equivalent commitment when the release is to the marine 
environment.  In this case, for the generalized situation 
considered here, transfer of radionuclides to man occurs only 
through ingestion of fish and shellfish. There may, however, be 
important specialized pathways of transfer in more specific release 
circumstances. 

485.  Somewhat greater collective doses result from releases to 
freshwater systems, particularly if the water is subsequently used 
for drinking.  Transfer of radionuclides to fish may also be 
somewhat more in fresh water than in the ocean. 

Table XI.1  Collective effective dose equivalent commitments
per unit activity released (10-12 man Sv per Bq)
-------------------------------------------------------------
                          Source                             
                                   Nuclear installations     
Radio-            Nuclear     Airborne  Release to   Marine
nuclide  Natural  explosions  release   fresh water  release
-------------------------------------------------------------
3H       0.0005   0.0008      0.0009    0.003        0.0008
14C      120      120         120       120          120
85Kr     0.0002   0.0002      0.0002    0.0002       0.0002
90Sr     -        0.6         1         1            0.00006
129I     -        30000       30000     30000        30000
131I     -        0.00009     0.4       0.009        0.00006
137Cs    -        2           5         0.6          0.001
222Rn    0.001    -           0.001     -            -
238Pu    -        10          100       0.3          0.002
239Pu    -        10          100       0.3          0.002
240Pu    -        10          100       0.3          0.002
241Pu    -        0.3         2         0.pp3        0.00002
241Am    -        3           200       10           0.06
-------------------------------------------------------------
These values are only approximations of a very generalized nature.  
The text contains a full discussion of limitations and of the 
specific assumptions utilized. 

486.  The largest doses generally result from airborne releases.  
The radionuclides in air may be inhaled, although the contributions 
to the dose from this pathway are usually small.  More importantly, 
the deposited radionuclides are available for incorporation into 
the general diet.  Long-term transfer of the radionuclides to man 
may then occur. 

487.  The greatest contributors to the collective dose commitments 
per unit activity released are the longest-lived radionuclides.   
Specifically, the very low dose rates from 129I result in 
substantial collective doses over the millions of years mean 
radioactive lifetime of this radionuclide.  The same is true of 14C 
and some isotopes of plutonium, for which the mean radioactive 
lifetimes are of the order of thousands of years.  Of course, the 
validity of assessments which involve such long-term projections 
must be questioned.  These results can not be related to 
equilibrium dose rates to the world's population.  For this 
purpose, limiting the integration periods to the duration of the 
practices, (for example a few hundred years at the most for nuclear 
power production) obtaining the incomplete collective dose 
commitments, has some merit. 

C.   EFFECTS EVALUATION

488.  The final step in evaluating the consequences of releasing 
radionuclides into the environment, once exposures have been 
determined, is to estimate the health effects. A great deal of 
study has been made of radiation exposure-response relationships 
and risk estimates have been formulated for the various effects.  

There is uncertainty, however, in knowing whether the risk 
estimates, which are generally only obtainable at higher doses and 
dose rates, will be valid for the low level chronic exposures of 
environmental situations. 

489.  It is not the intention of this document to provide detailed 
guidance regarding effects evaluations.  This requires more 
detailed specifications of the radionuclides present in the sources 
and the amounts released and closer consideration of the 
environmental conditions and the consequent exposures to the 
population involved. Only a brief summary has been presented 
(Chapter X) of the main aspects of radiation effects and of the 
considerations involved in assigning risk estimates. 

490.  In the generally accepted philosophy of radiation protection, 
all exposures are considered to increase the risk of harmful 
effects.  Increased risks are only justified when balanced by net 
positive benefits from the radiation operations.  The conceptual 
basis for measuring benefits and accounting for acceptability of 
risks are topics under review by national and international bodies. 

491.  The first part of the procedure of assessing the consequences 
of releasing radionuclides into the environment - the specification 
of sources, of environmental behaviour of radionuclides and of the 
consequent exposures - appears, in general, to be well founded, as 
shown by the reviews of radionuclide behaviour and dosimetry of 
this document.  This contributes in a general sense to the 
establishment of health criteria for these radionuclides.  The 
usual precautions must be expressed in applying general criteria to 
specific situations.  Considerable judgement is always required to 
make meaningful evaluations to serve as proper guides for future 
actions. It should also be appreciated that the results of 
radiological assessments are but one aspect of the considerations 
on which decisions must be made.  The better knowledge that has 
developed on these aspects should not unduly condition other 
factors, such as socio-economic considerations, which are also of 
importance in the rational selection of various possible options. 

ANNEX I

Excerpts from "Basic Safety Standards for Radiation Protection 1982 
Edition", (Safety Series No. 9) IAEA



    See Also:
       Toxicological Abbreviations