SOURCES AND EFFECTS
OF IONIZING RADIATION
United Nations Scientific Committee on the Effectsof Atomic Radiation
UNSCEAR 2000 Report to the General Assembly,
with Scientific Annexes

VOLUME I: SOURCES

UNITED NATIONSNew York, 2000

NOTE
The report of the Committee without its annexes appears as Official Records of theGeneral Assembly, Fifty-fifth Session, Supplement No. 46 (A/55/46).
The designation employed and the presentation of material in this publication do not
imply the expression of any opinion whatsoever on the part of the Secretariat of the United
Nations concerning the legal status of any country, territory, city or area, or of its authorities,
or concerning the delimitation of its frontiers or boundaries.
The country names used in this document are, in most cases, those that were in use
at the time the data were collected or the text prepared. In other cases, however, the names
have been updated, where this was possible and appropriate, to reflect political changes.

UNITED NATIONS PUBLICATION
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ISBN 92-1-142238-8

ANNEX A
Dose assessment methodologies

CONTENTSPage

INTRODUCTION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20
I.

GENERAL CONSIDERATIONS FOR DOSE ASSESSMENTS . . . . . . . . . . . . .
A. DOSIMETRIC QUANTITIES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1. Definitions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2. Age groupings . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
B. ENVIRONMENTAL BEHAVIOUR OF RADIONUCLIDES . . . . . . . . . . .
1. Transfer processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2. Parameters for dose estimation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3. Atmospheric dispersion from a near-surface release . . . . . . . . . . . . . .

21
21
21
22
23
23
23
24

II.

EXTERNAL IRRADIATION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
A. COSMIC RAYS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
B. NATURALLY OCCURRING RADIONUCLIDES . . . . . . . . . . . . . . . . . . .
1. Exposure processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2. Methods for estimating exposures . . . . . . . . . . . . . . . . . . . . . . . . . . .
C. RADIONUCLIDES IN AIR: CLOUD SHINEAND IMMERSION EXPOSURE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1. Exposure processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2. Methods for estimating exposures . . . . . . . . . . . . . . . . . . . . . . . . . . .
D. RADIONUCLIDES DEPOSITED ON SOIL . . . . . . . . . . . . . . . . . . . . . . .
1. Exposure processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2. Methods for estimating exposures . . . . . . . . . . . . . . . . . . . . . . . . . . .

27
27
28
28
28

III. INHALATION EXPOSURE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
A. NATURAL RADIONUCLIDES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
B. RADIONUCLIDES RELEASED TO THE ATMOSPHERE . . . . . . . . . . . .
1. Exposure processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2. Methods for estimating exposures . . . . . . . . . . . . . . . . . . . . . . . . . . .

35
35
36
36
36

IV. INGESTION EXPOSURE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
A. NATURAL RADIONUCLIDES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
B. RADIONUCLIDES RELEASED TO THETERRESTRIAL ENVIRONMENT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1. Transfer processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2. Food and water consumption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3. Methods for estimating exposures . . . . . . . . . . . . . . . . . . . . . . . . . . .

37
38

29
29
29
32
32
32

38
38
41
42

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

20

Page

C.

V.

RADIONUCLIDES RELEASED TO THE AQUATIC ENVIRONMENT . 46
1. Transfer processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46
2. Methods for estimating exposures . . . . . . . . . . . . . . . . . . . . . . . . . . . 47

GLOBALLY DISPERSED RADIONUCLIDES . . . . . . . . . . . . . . . . . . . . . . . . .
A. TRITIUM . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
B. CARBON-14 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
C. IODINE-129 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

49
49
51
53

CONCLUSIONS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 55
Tables . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 56
References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 79

INTRODUCTION
1. The estimation of exposures of human populations
from the various sources of radiation is an important andcontinuing goal of the Committee. In its previous assessments, the Committee took many different approaches todose estimation, depending mainly on the availability ofdata. These methods have been documented in the
UNSCEAR reports. To ensure that the methods are
relevant for continued use, the assumptions and parametersmust be reviewed from time to time and, if necessary,
updated for improved accuracy. The objective of thisAnnex is to provide such a review of dose estimation
procedures.
2. The initial work of the Committee involved evaluating the doses from natural background sources and fromthe testing of nuclear weapons in the atmosphere. In eachcase, the starting point of the calculations was where thefewest steps or assumptions would be needed, for example,
the concentrations of radionuclides in the body or thedeposition of radionuclides on the ground. To evaluate theexposures from nuclear power production, generic models
had to be used to estimate the dispersion of radionuclides
in the environment, the transfer to humans and the doses
from various pathways, since the concentrations or depositions were not measurable at the point of interest. To
evaluate the exposures resulting from the Chernobyl
accident, some of the dose estimation procedures weremodified to account for seasonal and other features indicated by available measurements.
3. In most cases, the Committee has been interested in
evaluating the average annual doses from the naturally
occurring levels of radionuclides in the environment andfrom the releases due to man-made practices or events.
There has been little need for detailed, time-dependent

dose modelling; the use of transfer coefficients or equilibrium modelling has been adequate for purposes of theCommittee. Data compilations have been generalized toallow widespread use in both time and space. Althoughprojections were needed to obtain committed doses, there
has been little emphasis on prognostic modelling. In
general, data-based methods of assessment with more
direct and simpler dose estimation procedures haveprovided results of reliable accuracy and allowed scientists
throughout the world to understand and apply or adapt
these same methods. This historical viewpoint issignificant and important to understand the evaluations of
the Committee. In specific circumstances, more theoreticalor more detailed models might have been more
appropriately considered, but these have generally not beenused by the Committee, nor will they be described or usedin this Annex.
4.
The Committee previously summarized its dose
estimation procedures in Annex A, “Concepts andquantities in the assessment of human exposures”, of the
UNSCEAR 1977 Report [U7] and in Annex A, “Doseassessment models”, of the UNSCEAR 1982 Report [U6].
These reviews are extended in this Annex with consideration of dose estimation procedures used in all earlier
assessments of the Committee. The selection of models andthe values of the parameters have been adjusted, based onbest available estimates.
5.
The procedures and models developed and used by
the Committee are believed to be reasonably accurate ingeneral application. They are largely based on empiricalevaluations of available measurements. In the widest sense,
the estimates of the average doses to the global populationfrom radiation sources are certainly well within the wide

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

variations that are known to exist. It is clear, however, that
more regionally appropriate values of environmental conditions or of human habits apply in specific circumstances.

21

Thus, the calculational procedures described here should
be used in other applications only with caution, and sitespecific data should be used where appropriate.

I. GENERAL CONSIDERATIONS FOR DOSE ASSESSMENTS
9.
The above definition of effective dose replaces aprevious similar definition of effective dose equivalent, HE:

A. DOSIMETRIC QUANTITIES
1. Definitions
6. For radiation assessment purposes, a number of
specialized quantities are used. A historical review of thequantities used by the Committee was presented in the
UNSCEAR 1988 Report [U4]. The Committee uses thesystem of radiation quantities and units adopted in 1980 by
the International Commission on Radiation Units andMeasurements (ICRU) [I8, I12] and the revised terminology and definitions proposed in 1990 by the International
Commission on Radiological Protection (ICRP) [I1].
7. For assessments by the Committee, the fundamental
dosimetric quantity used is the absorbed dose, D, averaged
over a tissue or organ; its unit is joule per kilogram, which
is given the special name gray (Gy). The relationship of
this quantity to the risk of biological effect is described by
the weighted dose quantities. Values of weighting factors
have been recommended by ICRP for the various types andenergies of radiation incident on the body or emitted fromwithin the body and for selected tissues and organs [I1].
Equivalent dose, HT, is the averaged absorbed dose in tissueor organ T, modified by the radiation weighting factor, wR:
HT 

jw D
R

R

(1)

T,R

where DT,R is the mean absorbed dose in tissue or organ T
due to radiation R. The unit of equivalent dose is joule perkilogram, and it is given the special name sievert (Sv).
Values of wR are given in Table 1.
8. Effective dose, E, is the sum of the weighted equivalent doses in all the tissues and organs of the body. It is
calculated from the following expression, where wT is theweighting factor for tissue T:E 

jw jw D
T

T

R

R

T,R

(2)

Effective dose has the unit joule per kilogram, which is
given the name sievert (Sv). ICRP has selected values of wT
to assess health detriment arising from the irradiation ofvarious organs and tissues. The recommended values of wT
are given in Table 2. The values have been selected for areference population of equal numbers of both sexes and awide range of ages. They apply to workers, to the generalpublic and to either sex.

HE 

j w (1977) j w D
T

T

R

R

T,R

(3)

which was promulgated by ICRP in 1977 [I11]. The
difference between HE and E is in the values of the weighting factors, wT. In equation (3) this is noted by appending(1977) to the wT expression. Values of wT(1977) are also
indicated in Table 2. Normally, this now outdated concept
would not be used by the Committee, but some very
extensive calculations of external dose coefficients havebeen performed and reported as values of HE rather than E,
and there is no unambiguous way to convert from one
value to another without access to the original calculations.
It is presumed that eventually these calculations will be
repeated so that values of E can be used in those few
circumstances where it is not now possible. For highenergy gamma radiation the numerical values of E and HE
should be approximately the same. However, for lowenergy gamma radiation, bremsstrahlung, and electrons,
the dose to the skin is typically much higher than the doseto any other organ, and the skin was specifically excluded
from consideration in HE. To simulate the value of E wherecomplete recalculation of E from HE is not possible, thevalue of 0.01 Hskin has been added to HE. This practice ofadding a weighted component of skin dose to HE wassuggested by ICRP [I14] in 1978 and was first used by theCommittee in the UNSCEAR 1982 Report [U6] to calculate doses from fission noble gases released from nuclear
reactors.
10. The term exposure is often used in the general
sense of being exposed to a radiation source, inferringthat a dose is received, but it also has a more specificdefinition. Exposure is the total electrical charge of ionsof one sign produced in air by electrons liberated by x or
gamma rays per unit mass of irradiated air at NTP. The
unit of exposure is coulomb per kilogram. An old unit,
the roentgen, R, is still used, as noted, for example, inreporting after the Chernobyl accident. One roentgen isequal to 2.58 104 C kg1. In this sense, the term exposure applies to ionization of air by x or gamma rays, but
the more common usage is also prevalent. Anotherdosimetric quantity is the kerma, which is the initialenergy of charged particles liberated by uncharged
particles in a unit mass of material. The unit is joule perkilogram, given the name gray (Gy). Under the assumption that charged particle equilibrium exists within the

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

22

volume of material, the kerma and absorbed dose may be
assumed to be equivalent. This assumption is used by
the Committee in most circumstances in specifying
absorbed dose rates in air or tissue.
11. When radionuclides are released to the environment,
they persist until they are lost through radioactive decay,
causing radiation exposures into the future. To compare
doses delivered over different time periods, the Committee introduced the concept of the dose commitment. The
dose commitment, Hc,T or Ec, is defined as the time integralof the average individual dose rate (per caput dose rate)delivered as a result of a specific practice:
H c,T "m

H T dt

Ec 

or

0

"

m E(t)dt
(4)

The integral is taken over infinite time to account for
exposures occurring during all future time and may thus
involve the average individual dose rates over generations.
The dose commitment from one year of a practice isnumerically equal to the equilibrium dose rate, if thepractice continues indefinitely at constant rate. If theintegration is carried out only to a specified time, this isthen termed a truncated dose commitment.
12. When prolonged exposure to a single individual froma single intake of a radionuclide is being considered,committed dose quantities are used. The time distributionsof the absorbed dose rates vary with the radionuclides, theirform, mode of intake, and biokinetic behaviour. The
committed equivalent dose, HT (τ), is defined as the time
integral of the equivalent dose rate, where τ is the integration time in years:t0τm H (t)dt


ST 

jH N
i

T,i

i

(6)

where Ni is the number of individuals in population
subgroup i receiving mean organ equivalent dose HT,i . The
collective effective dose, S, is defined in a similar manner.
The population and the time period over which the dose isdetermined should be specified. The collective dose commitment may become rather uncertain if applied to very longtime periods in which future environmental conditions andthe populations affected cannot be reasonably anticipated.
2. Age groupings

0HT(τ) population size. The collective equivalent dose, ST, is theaverage equivalent dose in an exposed group of individuals
multiplied by the number of individuals in each group:

(5)

T

t

The value of τ is taken to be 50 years for adults and fromtime of intake to age 70 years for children. The committed
effective dose, E(τ), is the sum of the committed equivalentdoses to tissues and organs multiplied by the appropriate
tissue weighting factors, wT. In general, the Committee
considers doses to adults; doses to children are considered
only when such doses are significantly different. ICRP has
developed age-dependent models for the respiratory andgastrointestinal tract and for the systemic biokinetic
behaviour of radionuclides that are of importance in theenvironment. These models have been used to computevalues of committed effective dose per unit intake bymembers of the public by inhalation and ingestion. These
values are compiled in ICRP publications [I2, I3, I4, I5],
and general use of these values is made by the Committee.
13. Collective dose quantities have also been used by theCommittee. These are aggregate quantities of dose and

14. In many instances, the effective doses in populations
have been estimated by the Committee for the adultindividual. Data on concentrations of radionuclides intissues have not always been widely available for other age
groups. In some cases, the uncertainties have been as great
as the possible differences. For certain radionuclides andpathways, however, the differences may justify separate
dose estimates. This is particularly true for 131I. The
availability of dose per unit intake estimates for other age
groups means that calculated dose estimates can be derived
from measured concentrations in foods, and more extensive
reporting of age-specific results can be expected in the
future.
15. Earlier estimates of doses from fallout 131I were madefor infants, using the age of 6 months as representative of the01 year age group [U7, U8]. For releases of 131I from nuclear
reactors, parameters were given in the UNSCEAR 1977
Report [U7] for the ages 6 months, 4 years, 14 years andadult. For assessment of exposures from the Chernobyl
accident, dose estimates for 131I were made for 1-year-old
infants and adults [U4]. In the UNSCEAR 1993 Report [U3],food consumption amounts were indicated for infants,
children and adults. In that report age-weighted annual
intakes ofnaturallyoccurring radionuclides were then derived,
assuming the fractional distribution of adults, children andinfants in the population to be 0.65, 0.3, and 0.05,
respectively. An age-independent dose per unit intake (theadult value) was applied [U3]. Age-dependent dosecoefficients are now available from ICRP, and the number ofage groups considered could be expanded to six: 3 months(from 0 to 1 years), 1 year (from 1 year to 2 years), 5 years(>2 years to 7 years), 10 years (>7 years to 12 years), 15 years(>12 years to 17 years), and adult. For most purposes, theCommittee will consider the age categories of infants,
children, and adults and use the available dose coefficients
corresponding to 12 years, 812 years, and >17 years,
respectively, for these categories. The fractional distribution of
the population within these categories is that mentioned
above, namely, 0.05, 0.3, and 0.65 for infants, children andadults, respectively.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

B. ENVIRONMENTAL BEHAVIOUR OF
RADIONUCLIDES
1. Transfer processes
16. Radionuclides are generally released in trace
quantities to the environment. They are then physicallytransported in the air or water media in which they arelocated. The measurements of radionuclide transfers frompast releases have been used to study and infer large-scaleatmospheric and hydrological movements on the earth. The
fallout radionuclides 90Sr and 137Cs have been used to infermaterial removal or renewal times (residence times) inenvironmental regions. Tritium is a tracer for the worldhydrological cycle and 14C for the global carbon cycle. The
specific removal or transfer processes of the various
exposure pathways have been extensively studied.
17. Radioactive materials, either particles or gases, may betransported great distances by local and large-scale air
movements. The time periods that the materials remain
airborne depend on the latitude, time of year and height ofinjection into the atmosphere. The depletion processes include
gravitational settlement and dryimpaction, incorporation intorain drops and washout by falling precipitation. The physicaland chemical characteristics of the materials themselves, suchas particle size and chemical and physical forms, may
influence the removal rates.
18. The predominant features of large-scale mixing
processes and air movements in the atmosphere were
presented in the UNSCEAR 1982 Report [U6] inconnection with discussion of exposures from nuclear
explosions. They were used to describe the occurrence of
fallout. The measured deposition of 90Sr could, however, be
used as a starting point for the dose assessment, obviating
the need to evaluate the deposition from the uncertain
input amounts. With improved estimates recently availableof the input of fission radionuclides to the atmosphere from
nuclear tests, quantitative aspects of the general model canbe pursued with seasonal values of residence times
assigned to the various compartmental regions andlatitudinal deposition estimated. This exercise is discussed
in Annex C, “Exposures to the public from man-madesources of radiation”.
19. Releases of radionuclides from nuclear fuel cycle
installations occur at ground level or through stacks ofassumed representative heights of 30 or 100 m. The longterm, sector-averaged Gaussian plume model can be used
to calculate air concentration for limited distances
following airborne releases. Estimates can be obtained
directly from the model or from a simple analyticalexpression that gives a good fit to the model results. The
air concentration at one kilometre per unit release istypically 5 107 s m3 and decreases as a result of further
dispersion at a rate inversely proportional to distance,
expressed in kilometres, raised to the power 1.21.4.
Derivation of these quantities is discussed in Section I.B.3.

23

Integration to 50 or 100 km defines the local exposures.
Further integration to a distance of 2,000 km defines thecontinental or regional component of exposure. Most
particles from near-surface releases are deposited within
this distance. Only fine aerosols and gases may become
further dispersed in the troposphere.
20. Global modelling of atmospheric releases will be
described with respect to the specific radionuclides. Mixing
occurs first within the latitude band, then within thehemisphere. Gradual interhemispheric exchange occurs forgases such as 85Kr, for which removal processes areminimal. Tritium and 14C enter the global cycles of therespective elements.
21. Radioactive material released to the aquatic
environment is transported and dispersed by advective andturbulent processes occurring in the water body.
Interactions of radionuclides with suspended matter andsediments may remove radionuclides from the solution.
Methods for modelling hydrological transport have beendeveloped and applied, usually for specific categories ofwater bodies: lakes, rivers, estuaries, coastal seas andoceans.
22. UNSCEARhas needed hydrological transport estimatesto evaluate the exposures from releases of radionuclides from
nuclear fuel cycle installations. For fuel reprocessing plants,
use has been made of dispersion estimates surrounding theplants at Sellafield and La Hague. In the general case forreactor releases, the Committee made use of relationships
between water volumes, water usage and potential intake toestimate collective doses. The water uses considered included
drinking water, fish and seafood production and irrigation.
Some minor pathways might be involved in the local regions,such as immersion and exposure to shoreline contaminants.
Some general considerations with regard to aquatic models
and suggestions about which models to use have been
published, e.g. [S2]. Details of the procedures used by
UNSCEAR will be presented later in connection withingestion exposures.
2. Parameters for dose estimation
23. The basic parameters used in models to describe
environmental behaviour and transport of radionuclides
and to make dosimetric calculations are transfer
coefficients, Pij. These describe the relationships of integrated
concentrations or dose in successive environmental
compartments, e.g. movement from compartment i tocompartment j. The pathways of transfer of radionuclides
through the environment commonly evaluated in UNSCEAR
dose assessments are illustrated in Figure I along withdesignations of the transfer coefficients. As an example, P34 is
the time-integrated activity concentration of a radionuclide in
the body divided by the time-integrated concentration of thesame radionuclide in the diet. This methodology for deriving
relationships between measured quantities has been used by
the Committee since 1962.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

24

SourceP01

External irradiation P
1

15

AIR
P

12

2

SOIL

Inhalation P14
3
P23

DIET

Ingestion 4

BODY

P34

5
P

DOSE

45

External irradiation P25

Figure I. Terrestrial pathways of transfer of radionuclides and dose to humans.

24. For a particular environmental transfer pathway, theamount of radionuclide released to the environment
multiplied by the intervening transfer coefficients gives anestimate of the resulting effective dose. If measurement results
are available at any point in the chain, the calculation may
begin at that point. This minimizes the uncertainties that may
exist in determining transfer coefficients for earlier steps in
the transfer pathway. Thus, assessments of dose derived bytheCommittee have started with integrated concentrations ofradionuclides in air, deposition densities, measured
concentrations in foods or body burdens.
25. The measurements used to evaluate transfer
coefficients have been made over a number of years by
research and monitoring organizations in many locations.
The transfer coefficients derived for estimation of effective
doses from atmospheric nuclear testing were summarized
in the UNSCEAR 1993 Report [U3] for a long listing ofradionuclides for the pathways of external irradiation,
inhalation and ingestion. The many measurement results
acquired following the Chernobyl accident have shown that
more seasonal or locallycharacteristic conditions should betaken into account in evaluating exposures from specificsingle releases of radioactive materials. Several programmes
to compare results and validate models were instigatedfollowing the Chernobyl accident. Those activities arecontributing results useful for deriving specific values formany transfer coefficients, e.g. [I16, I17].
26. Tritium and 14C are modelled differently than otherradionuclides, since they are mobile in the environment
and are readily incorporated into living organisms. The
transfer of tritium and 14C is not modelled using transfer
parameters but by a specific-activity approach. For tritium,
it is assumed that the tritium to hydrogen atom ratio in thevarious environmental compartments is simply proportional to the ratio in moisture in air. For carbon, the 14C
activity per gram carbon in all compartments is assumed tobe the same as that in air.
27. The specific methods used by the Committee toestimate doses to humans caused by releases of radioactive
materials to the environment are described in the following
Chapters. The rationale for the selection of the various
parameters is presented, so that it will be clear whenalternative selections might be desirable for specific local
conditions. The methods are intended to be widely
applicable, and since they are largely empirically based,
they should provide realistic estimates of doses in mostgeneral circumstances of release of radionuclides.

3. Atmospheric dispersion
from a near-surface release
28. Radionuclide concentrations in the environment
downwind of an isolated source such as a nuclear reactor
are usually undetectable at distances greater than a few
kilometres. In such cases, the air concentrations needed as
the starting point for dose assessments to the public mustbe estimated using a mathematical model.
29. Average air concentrations close to a specific source are
traditionally calculated using the long-term sector-averaged
Gaussian plume model [I15]. In this model, the plume isassumed to spread uniformly across a sector subtended by anangle ∆θ (usually chosen to be 30o). Air concentrations at agiven distance downwind are calculated for each of six
atmospheric stability classes using average values of windspeed, inversion height and vertical dispersion parameter foreach class. The long-term mean concentration is found bysumming over classes, taking into account the frequency ofoccurrence of each class and the frequency with which thewind blows towards the site of interest. The model is able toaccount for reductions in air concentration due to wet and dry
deposition. A general discussion of the processes governing
atmospheric dispersion was presented in the UNSCEAR 1982
Report [U6].
30. The mathematical statement of the long-term sectoraveraged Gaussian plume model is as follows:

C a,j 2

B

1/2x )2 jfjQ D w

6i1

f i F( Fz,i, H , h i) exp ( 8 x u i1) D d,i(7)(u i Fz,i)where Ca,j is the long-term average air concentration(Bq m3) in sector j; fj is the frequency with which the windblows into sector j; Q is the release rate (Bq s1); x is thedownwind distance (m); ∆θ is the sector width (radians);
fi is the frequency of occurrence of stability class i;
F(σzi,H,hi) is the vertical shape function; σz,i is the verticaldispersion parameter for stability class i (m); H is theeffective release height (m); hi is the mixed layer height forstability class i (m); λ is the radioactive decay constant for
the radionuclide in question (s1); Dd,i is the depletion
factor for dry deposition; Dw is the depletion factor for wet

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

deposition; and ui is the average wind speed for stability
class i at the release height (m s1).
31. For ground-level concentrations, the vertical shapefunction is given byF( Fz,i, H , h i) j exp  (H 22nF h )
"

2

i

n"

(8)

2
z,i

which accounts for reflection from the ground and from anelevated inversion through the method of virtual sources.
The summation index n in equation (8) represents thenumber of reflections that the plume has undergone. The
summation converges slowly in some applications. To
simplify the calculation, Yamartino [Y1] proposed
approximations as follows: for σz,i / hi  0.63, truncate
equation (8) at n = 0, ± 1; for 0.63 < σz,i/h i  1.08, F(σz,i,H,hi) = (2π)½ σz,i/hi (1  k2)[1 + k2 +2k cos(πH/hi)] where k= exp[½ (π σz,i/hi)2]; for σz,i/hi > 1.08, F(σz,i, H,hi) = (2π)½σz,i/hi. These approximations result in minimal error inevaluation of equation (8).
32. Plume depletion due to dry deposition is normallytreated using the source depletion method, in which casethe depletion factor takes the form

be used for the meteorological and release parameters
appearing in the equations. In the absence of site-specific data,
the representative values listed in Table 4 give reasonableestimates of air concentrations. Values of λ, vd and Λ should
be chosen for the radionuclide of interest.
35. One aim of applying the above method is to derive
long-term average dilution factors, Ca/Q, for downwind
distances between 1 and 2,000 km from the source. The
results of the calculation are given in Table 5. A long-lived
radionuclide was assumed so that radiological decay could
be neglected. The parameter values in Table 4 were used
and the deposition velocity vd and washout coefficient Λ
were set to representative values of 0.002 m s1 and
0.0001 s1, respectively. Precipitation was assumed tooccur 500 hours per year, 80% of the time during class D
conditions and 20% during class C, at an average rate of
1.5 mm h1. The washout time ts was assumed to be equal
to the travel time tt between source and receptor for tt < 4hours and equal to tt/2 for tt >24 hours; in the range 4 < tt
< 24, ts was assumed to vary linearly between tt and tt/2.
36. The variation of air concentration with downwind
distance beyond 1 km can be approximated by thefollowing simple function, which was used in previous
UNSCEAR assessments:(11)
C (x)  D Q x na

D d,i  exp "i

x

m
0

exp



(9)
dx

where αi = (2/π)½ vd/ui and vd is the dry deposition velocity(m s1). The depletion factor for wet deposition is given by
Dw = exp (Λts), where Λ is the washout coefficient (s1)and ts (s) is the time over which precipitation occurs during
the travel of the plume from source to receptor.
33. A number of investigators [B13, P6, V1] havesuggested forms for the vertical dispersion parameter. The
following scheme of Smith [S1] and Hosker [H8] is used,since it is able to take account of the surface roughness,z0(m), of the site:

Fz  g(x) F(x, z 0)

1

2

H
2 Fz,i2 (x)Fz,i (x)25(10)

where g(x) = axb/(1+cxd) and F(x,z0) = ln[pxq[1+(rxs)1]]
when z0>0.1 m and F(x,z0) = ln[pxq(1+rxs)1] whenz00.1 m. The parameters a, b, c, and d depend on theatmospheric stability class, and the parameters p, q, r, ands depend on the surface roughness. Representative values
are given in Table 3.
34. Equations (7) to (10) provide a relativelysimple method
for calculating long-term average air concentrations due to a
specific source. Wherever possible, site-specific values should

where D1 is the dilution factor at 1 km (s m3) and x is thedownwind distance (km). Figure II shows that equation(11) gives a very good representation of the detailed resultsof Table 5 and can therefore be used to estimate airconcentrations in place of equations (7) to (10) if, forexample, site-specific data are not available. The best
approximation to the calculated results is obtained withvalues for D1 and n of 5.3 107 s m3 and 1.42, respectively.
The value for n is similar to the value of 1.5 used inprevious UNSCEAR assessments. The value for D1 is lower
by a factor of 6 than the value of 3 106 s m3 suggested in
the UNSCEAR 1982 Report [U6]; this value reflects
concentrations at a location toward which the wind blows
about 50% of the time, whereas the currentlyrecommended
value of 5 107 s m3 assumes a uniform wind rose at the
point of release.
37. The variability in calculated results has been
investigated by altering the parameter values used inequations (7) to (10). The meteorological parameters werevaried to cover the range of conditions that could occur
from time to time. The variability in deposition velocity
and washout coefficient reflects the values associated withdifferent radionuclides. Each parameter was varied inturn, holding all other parameters at the values given
above. Results are shown in Table 6 in terms of D1 and n,
the parameters required to implement equation (11). D1 isrelatively insensitive to changes in the values of theparameters except for wind speed and release height; n issensitive to these parameters, as well as to deposition

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

26
10 -6

value of 1.4 be used for this radionuclide. For calculationof radionuclide concentrations at a specific site, values of
D1 taken from Table 6 should be modified to reflect thefrequency with which the wind blows towards the locationof interest. For the purposes of calculating representative
population doses using the method presented in this Annex,
a uniform wind rose was assumed, with a frequency of 1/12,
or 0.083, averaging over 12 sectors.

DILUTION FACTOR (s m -3)

10 -7

10 -8

10 -9

10 -10n = 1.2
10 -11n = 1.4
n = 1.6

10 -12
1

10

100

1000

10000

DOWNWIND DISTANCE (km)Figure II. Dilution factor for estimation of air concentrations at distances from a surface release
determined from Gaussian plume model calculation(points) compared to power-function representationsof the form 5 10-7 r-n.

velocity and inversion height. Release heights can varyfrom low-level building vents to stacks of 100 m or more.
Building entrainment may reduce the effective releaseheight to some extent. The representative height of 30 m
has been retained as in earlier UNSCEAR assessments for
estimating collective doses following releases from nuclear
installations. Individual dose evaluations could depend
more critically on the release height assumption. As an
example, calculations indicate that for release from a stackof 150 m height, the collective dose in the local andregional area per unit release would be about 25% of that
estimated for a 30 m stack. The model is not very sensitive
to roughness length, washout coefficient, or the frequencyof stabilityclasses. Thus, adequate estimations can be madeusing the representative values suggested for theseparameters without the need for site-specific data.
Although wet deposition is very effective at removing
material from the plume, precipitation occurs less than
10% of the time and has little effect on long-term average
air concentrations. However, reliable values of the washout
coefficient are needed to calculate accurately the amount ofmaterial deposited on the ground and on vegetation by wet
deposition.
38. The results in Table 6 can be used to interpolate thevalues of D1 and n that most closely represent meteorological conditions at the site and the radionuclides ofinterest. Equation (11) can then be used to estimate airconcentrations at the downwind distance in question. For
noble gases, which do not deposit, a value of n equal toabout 1.2 should be used as long as other parameter values
remain near the representative values defined here. Tritium
should also be assigned a value of 1.2, since most tritium
deposited under dry conditions is quickly re-emitted to theatmosphere. Carbon-14 is efficiently deposited and partially returned to the atmosphere through plant and soil
respiration. On balance, it is recommended that the index

39. The long-term, sector-averaged Gaussian plumemodel has been extensively tested at local distances. Whenused with site-specific meteorological data, theuncertainties in its predictions are less than a factor of 2
within 10 km of the source and less than a factor of 4
between 10 and 100 km of the source [C10, H6, R4]. Useof the model is therefore adequate for local assessments.
Validation of the model on regional scales is more difficult.
Few point sources are strong enough or emit a uniqueenough contaminant to be detected unambiguously at
downwind distances greater than 100 km. Thus few datafrom routine releases can be used to test the model.
Regional-scale tracer studies have been carried out, but
only over short periods of time. These must be consideredcase studies that provide information only for themeteorological conditions prevailing at the time of therelease. They cannot be used to infer long-term average airconcentrations.
40. The problem of acid precipitation has driven thedevelopment of a number of models that simulate the longrange transport of air pollutants [J3]. These models aremuch more sophisticated than the Gaussian model
described above in their treatment of plume transport anddeposition and can track pollutants through space- andtime-varying meteorological conditions. They aremoderately successful in predicting the broad features of
the concentration field on regional scales. However, theyrequire considerable expertise, computer resources, andinput data to run and are therefore unsuited to the types ofassessments performed by the Committee. Comparisons of
their predictions with those of the Gaussian model would
help to establish the validity of the latter, but such studies
have not yet been done.
41. Although the accuracy of the predictions of theGaussian model beyond 100 km is difficult to quantify, anumber of factors suggest that the model overestimates true
concentrations at this range:
(a)

(b)

The model assumes that the plume travels in astraight line from source to receptor. In reality,
variations in wind direction will generally lead toquite complicated trajectories that increase the travel
time between source and receptor and provide theopportunity for enhanced mixing.
The model assumes that the stability class in effect at
the start of the release remains in effect until theplume reaches the receptor. In reality, a plumetravelling over hundreds or thousands of kilometres

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

(c)

(d)

will experience several diurnal cycles and a fullrange of atmospheric stabilities. A plume that has
undergone one or more unstable periods will bemixed through a deep vertical layer. It will remain
well mixed through subsequent night-time stable
periods and not be confined beneath a low capping
inversion, as is assumed in the model;
The model assumes that the plume is transportedwith the wind speed at the effective release height. In
reality, as the plume mixes to greater heights, theeffective transport velocity must be averaged overdeeper layers. Since wind speeds generally increaseaway from the ground, the result is extra dilution andlower concentrations;
The model assumes that the terrain over which theplume passes is flat, a supposition unlikely to holdover regional distances. Complex topography willtend to increase turbulence levels and deflect theplume trajectory, thereby reducing concentrations.

42. In contrast to the model features mentioned above,
the procedure for estimating plume depletion due to dry
deposition is not conservative. Material is assumed to belost uniformly over the entire depth of the plume when in
reality it is lost only at the deposition surface. Horst [H7]27

showed that this approach underestimates airborne
concentrations byan amount that increases with increasing
atmospheric stability, greater downwind distances, andlarger deposition velocities.
43. Thus, the accuracy of the Gaussian model at regional
scales is unknown, and uncertainties are large, but
probably within a factor of 10 for relatively simple
situations. The uncertainties would be somewhat smaller
for population doses since the concentration averaged overall distances and directions is probably better known thanthe concentration at a point.
44. In summary, the value of the dilution factor, D1, of 5
107 s m3 is assumed by the Committee to be representative for evaluating collective doses per unit release whensite-specific data are not available. The value is not very
sensitive to variations in meteorological or deposition parameters. The release height can be of greater influence and,if known to be different from the representative value of
30 m, should be taken into account. The index parameter,n, is more variable than the dilution factor with respect tometeorological and deposition conditions, but a value of 1.2
for noble gases and tritium and 1.4 for other radionuclides
should provide reasonable estimates of air concentrations.

II. EXTERNAL IRRADIATION
45. External irradiation from radionuclides naturally
present in the environment or released from man-madepractices or events is usually an important component of
the exposure of human populations. These exposures derive
primarily from gamma radiation arising from the decay ofthese radionuclides at locations outside the human body.
Secondarily, exposures to the skin from beta radiation may
be considered. The methods used by the Committee toestimate external exposures from the various sources arereviewed in this Chapter.

A. COSMIC RAYS
46. Cosmic rays originate in outer space; they consist
primarily of protons and alpha particles. Interactions in theupper layers of the earth’s atmosphere create secondary
components; the more important secondary particles from adose-assessment view are muons, neutrons, electrons,
positrons, and photons. Exposure to cosmic rays is strongly
dependent on altitude and weakly dependent on latitude. Doseassessments are based on both measurements and calculationsof the radiation transport to infer the dependence on altitude.
At lower levels of the atmosphere and at sea level, thedependence on the 11year solar cycle is small compared tothe uncertainty in the estimates and is currently ignored.

47. The method used by the Committee to assess doses
from the photon and directly ionizing component of cosmic
radiation at sea level has not changed substantially in many
years. In the UNSCEAR 1977 Report [U7] the basic value
was considered to be the ion-pair production rate, forwhich a value of 2.1 cm3 s1 was adopted. This value wasconverted to a dose rate of 32 nGy h1 and has been
assumed to be numerically equal to the effective dose rate
[U3, U4]. A mean shielding factor of 0.8 has been applied
to derive an indoor effective dose rate of 26 nSv h1. With
the further assumption that the average fraction of timespent indoors is 0.8 [U3, U4], the annual effective dosefrom the ionizing component of cosmic rays at sea level isjudged to be 240 µSv. Estimates of cosmic ray dose rates atelevations above sea level are obtained using a procedurepublished by Bouville and Lowder [B12]:E I(z)  E I(0) [0.21 e 1.649z  0.79 e 0.4528z](12)where E< 1(0) is the dose rate at sea level, 240 µSv a1, andz is the altitude in km. The dose rate from the photon andionizing component is known to vary with latitude, but thevariation is small. The dose rate is about 10% lower at thegeomagnetic equator than at high latitudes.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

28

48. For the neutron component of the cosmic radiation
exposure, the radiation field and the estimates of effective
dose have been more uncertain owing to a lack ofmeasurements. Recent measurements and calculations are
beginning to provide clarification. Because earlier
instrumentation had a low response to high-energy neutrons,
which are an important component of the spectrum, someincreases in the fluence rate and effective dose are being
suggested. Measurements made using a Bonner sphere
spectrometer [R3, S8] at the top of the Zugspitze mountain inGermany (altitude 2,963 m, atmospheric depth 718 g cm2)and associated calculations give a fluence rate of 0.126 ± 0.01cm2 s1 [S9]. Attenuation with altitude was described using
the function e0.00721p, where p (g cm2) is the atmosphericdepth. From this, a fluence rate at sea level (p = 1,033 g
cm2) of 0.013 ± 0.001 cm2 s1 can be derived. Measurements
also with Bonner sphere spectrometers gave a value of 0.0133± 0.001 cm2 s1 at about sea level for a geomagnetic latitudeof 53-N near Braunschweig in Germany [A6], and a value of
0.0123 cm2 s1 at sea level for a geomagnetic latitude of
45-N in Hampton, Virginia in the United States [G3]. The
effective dose (isotropic) corresponding to a fluence rate of
0.013 cm2 s1 obtained by applying a neutron fluence energy
distribution weighting factor of 200 pSv cm2 [S9] (equal to
720 nSv h1 per neutron cm2 s1) is 9 nSv h1. Birattari et al.
[B14], using an extended range remmeter, reported a value of
9 nSv h1 (±5%) in agreement.
49. The shape of the neutron energy spectrum athabitable altitudes is considered to be relatively invariant,and therefore the fluence to effective dose (isotropic)
conversion coefficient is expected to be generally valid. On
this basis, the annual effective dose rate from neutrons at
sea level would be estimated to be 80 µSv a1. This issubstantially larger than the value of 30 µSv a1 used in the
UNSCEAR 1993 Report [U3] and is still subject to great
uncertainty; the main factor in the increase in thecalculated dose is the inclusion of high-energy neutrons.
With the application of a shielding factor of 0.8 and an
occupancy factor of 0.8, the annual average effective doseat sea level is estimated to be 65 µSv at geographic
latitudes between about 40- and 50-.
50. For calculations of outdoor cosmic ray neutron dose
rates at other altitudes, the relation between height, hv, in
km, above sea level and atmospheric depth [R3] is, for p >230 g cm2,(13)
h  44.34  11.86 p 0.19

51. The fluence of neutrons, which arise from collisionsof high-energy protons within the upper atmosphere, isstronglyinfluenced by geomagnetic latitude. This variation
at habitable altitudes has not been satisfactorily quantified,
as measurements at different latitudes have not always
been comparable. Recent measurements at high altitudes
have shown a variation by a factor of about 4 [G3], withthe lower values near the equator. These results support thecalculations of Florek et al. [F3], who used the Los Alamos
Lahet Code System (LCS) to simulate neutron fluence asa function of latitude. Their results are expressed in terms
of kn, a latitude coefficient, as follows:E N (lat)  E N (90) k N(lat)(14)

with kφ ranging from 1.0 at 90- to 0.8 at 47-, 0.6 at 42-,
0.4 at 35- and 0.2 at the equator. The application of thisrelationship to available measurement results is discussed
in Annex B, “Exposures from natural radiation sources”.

B. NATURALLY OCCURRING
RADIONUCLIDES
1. Exposure processes
52. Many radionuclides occur naturally in terrestrial soils
and rocks and in building materials derived from them. Upon
decay, these radionuclides produce an external radiation field
to which all human beings are exposed. In terms of dose, theprincipal primordial (half-lives comparable to the age of theearth) radionuclides are 40K, 232Th, and 238U. Both 232Th and238
U head series of radionuclides that produce significant
human exposures. The two series are listed and discussed fullyin Annex B, “Exposures from natural radiation sources”.
53. The decay of naturally occurring radionuclides in soil
produces a gamma-beta radiation field in soil that also crosses
the soilair interface to produce exposures to humans. The
main factors that determine the exposure rate to a particular
individual are the concentrations of radionuclides in the soil,
the time spent outdoors, and the shielding by buildings.
However, as the materials of which most buildings are built
also contain radionuclides, the shielding by buildings of theoutdoor radiation field is often more than offset by thepresence of additional radionuclides in the building materials.
2. Methods for estimating exposuresv

Both altitude and latitude variations in the cosmic ray
neutron dose rate must be known to determine thepopulation-weighted average exposure of the world
population. Calculations of dose from cosmic rays toairline crews and passengers are based on measurements
and on detailed calculations using radiation-transport
codes tailored to follow the altitude and latitude of aparticular flight.

54. Two methods of evaluating external exposures fromnaturally occurring radionuclides have been used by theCommittee. The first is simply to summarize directly
measured external gamma dose rates in air outdoors andindoors, subtracting the dose rate due to cosmic rays. The
second is to calculate the external gamma dose rates in air
from measurements of the concentrations of the relevantradionuclides in soil. The two methods have provided
generally consistent estimates of exposure.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

55. Surveys with direct measurements of dose rate in air
from naturally occurring terrestrial radionuclides havebeen made in most inhabited regions of the world. In the
UNSCEAR 1993 Report [U3], data were included forcountries or regions in which three fifths of the world
population resides. Countryaverage dose rates ranged from24 to 160 nGy h1, with a population-weighted average of
57 nGy h1. The population-weighted average derived fromthis large sample was assumed to provide a representative
global value of outdoor external exposure.
56. Surveys to determine the concentrations of radionuclides in soil have also been made. These results can be
related to exposures by using estimates of the dose rates in
air per unit concentration of radionuclide in soil. The
Committee has relied on the calculations of Beck [B8] formany years. Extensive Monte Carlo calculations of kerma
in air and of organ dose for terrestrial gamma rays havebeen reported by Petoussi et al. [P4], Saito et al. [S5], andEckerman and Ryman [E7]. Results from three separate
calculations are included in Table 7; the values are quite
similar and can be considered equal. Uncertainty in theassumed average composition of soil could lead todifferences of greater magnitude [E7].
57. Absorbed dose rates in air indoors have also been
extensively measured. The values reported in the
UNSCEAR 1993 Report [U3] covered areas in which overa third of the world population lives. Country averages
ranged from 20 to 190 nGy h1, with a population-weighted
average of about 80 nGy h1. The population-weighted
average of the ratio of indoor to outdoor dose was 1.4.
Some of the outdoor measurements may have been
influenced by the presence of buildings nearby. The valueof the indoortooutdoor ratio is very sensitive to thestructural properties of buildings (materials and thickness).The building materials act as sources of radiation andalso as shields against outdoor radiation. In wooden andlightweight houses, the source effect is negligible, and thewalls are an inefficient shield against the outdoor sourcesof radiation, so that the absorbed dose rate in air could be
expected to be somewhat lower indoors than outdoors. In
contrast, in massive houses made of brick, concrete orstone, the gamma rays emitted outdoors are efficiently
absorbed by the walls, and the indoor absorbed dose ratedepends mainly on the activity concentrations of naturalradionuclides in the building materials. Under thesecircumstances, the indoor absorbed dose rate is generally
higher as a result of the change in source geometry, with
the indooroutdoor ratio of absorbed dose rates in air
between 1 and 2.
58. The Committee has used a coefficient of 0.7 Sv Gy1
to convert absorbed dose in air to effective dose equivalentand effective dose. This result was based on an analysis in
the UNSCEAR 1982 Report [U6], and more recent
calculations have confirmed the validity of this value for
adults. However, newer calculations [P5, S11] using Monte
Carlo radiation-transport codes indicate that higher values

29

should be used for infants and children. These values,
given in Table 8 for average energies of gamma rays, are
0.9 Sv Gy1 for infants and 0.8 Sv Gy1 for children.
59. In order to combine indoor and outdoor dose rates tocompute total doses, the Committee continues to use anindoor occupancy factor of 0.8, which implies that people
spend 20% of the time outdoors, on average, around theworld. The estimated 80% of time spent indoors is
considered likely to be low for industrialized countries intemperate climates and high for agricultural countries inwarm climates.

C. RADIONUCLIDES IN AIR:
CLOUD SHINE AND IMMERSION EXPOSURE
1. Exposure processes
60. Following the release of radionuclides to theatmosphere and before their deposition onto the ground,human beings may receive external exposure. Twosituations are usuallydistinguished: external exposure fromthe cloud passing overhead (referred to as “cloud shine”)
and external exposure from radionuclides in air
surrounding the human body (referred to as “immersion”).
The radiation dose from immersion is nearly always muchlarger than that from cloud shine. The dose fromimmersion can be readily calculated from the measured,
integrated concentrations of radionuclides in air. The dosefrom cloud shine is rarely calculated; its importance would
be significant only if other exposure pathways were absent.
One such example would be for persons underneath anelevated, passing plume.
61. Effective doses from immersion are typicallycalculated for gamma-emitting radionuclides, but beta andeven alpha particles can also produce external doses to theskin. Some radionuclides, notably 85Kr, which emits a weakbeta particle, produce nearly all of their dose via thepathway of immersion.
2. Methods for estimating exposures
62. Because of their relative insignificance, theCommittee has seldom considered external exposures fromcloud shine or immersion. Exceptions were made for the
Chernobyl accident and for the release of noble gases fromreactor operations. Since initial estimates of such exposures
were made, tissue-weighting factors and terminology todescribe equivalent and effective doses have changed [I1],and newer calculations of dose rates from immersion havebeen published [E7]. The net changes in the calculated
numbers appear to be small.
(a) Atmospheric nuclear testing
63. Although the potential pathways of cloud shine andimmersion were considered in the first report of theANNEX A: DOSE ASSESSMENT METHODOLOGIES

30

Committee, the UNSCEAR 1958 Report [U13], the doses
from these pathways for radionuclides released from
explosions of nuclear weapons have not been evaluated.
The conclusion was reached that, except at the immediatesite of the explosion, external irradiation from airborne
material is negligible in comparison with external
irradiation from fission products deposited on the ground.
As much of the material from nuclear explosions was
injected into the stratosphere or high troposphere, most of
the short-lived radionuclides potentially responsible for the
majority of dose from cloud shine or immersion would
have decayed before reaching the earth’s surface.

available. In others, data for only one or a few radionuclides were available. In the latter case concentrations ofother radionuclides were inferred from ratios measured in
nearby countries. In some cases, no measured air concentrations were available, so the integrated air concentration of 137Cs was inferred from its ground-deposition
density and a nominal quotient of ground deposition to
integrated air concentration of 1,000 m d1 [U4]; theintegrated air concentrations of other radionuclides werethen inferred from the ratios to 137Cs measured at otherlocations.
(c) Nuclear installations(b) The Chernobyl accident
64. Doses from “external irradiation during cloud
passage” were calculated for the releases of radionuclides
from the Chernobyl accident [U4]. Although exposure rates
could in theory be measured directly, in practice it isgenerallyimpossible to distinguish this smaller component
from radiation arising from material deposited on the
ground. Doses can, however, easily be calculated frommeasured air concentrations or inferred from measured
deposition densities.
65. The cloud-gamma dose for radionuclide i is evaluatedfrom the formulaE i  C a,i d c,i (1  F 0)  C a,i d c,i F 0 F s(15)where E is the effective dose (Sv) from external radiation
during cloud passage; Ca* is the integrated concentration in
outdoor air (Bq d m3); dc is the effective dose coefficient
per unit integrated air concentration (Sv per Bq d m3); F0
is the indoor occupancy factor (the fractional time spentindoors); and Fs is the building shielding factor (the ratio
of indoor to outdoor dose rate).
66. The first term in equation (15) is the component
received while the individual is outdoors, and the second
term is the component received indoors. At the time of theChernobyl assessment, values from Kocher [K7] wereused; these values were for HE + 0.01Hskin rather than E.
The values used then and the newer recommended valuesof HE + 0.01Hskin from Eckerman and Ryman [E7] arelisted in Table 9.
67. For the Chernobyl assessment, an indoor occupancy
factor of 0.8 and a building shielding factor of 0.2 were used
for all countries. The values of these factors had been usedpreviously by the Committee [U6, U7]. It was noted, however, that measurements as well as calculations of the shielding factor afforded by buildings showed a large variation,
depending on the type of building [C8, M6, S6, U4].
68. To make the above calculation, it is necessary toknow the integrated concentration in air of the many shortlived radionuclides. In some countries, complete data were

69. During the operation of nuclear reactors, several
fission noble gases are released, as is the activation
radionuclide 41Ar. Among the more prominent fission
noble gases are 133Xe from pressurized water reactors and85
Kr, 87Kr, 88Kr, 133Xe, 135mXe, 135Xe and 138Xe from boilingwater reactors [U6]. Much of the dose from these (andother) radionuclides is delivered by the pathway of cloudshine and immersion. Later reports [U3, U4] of theCommittee refer to the models developed in the UNSCEAR
1982 Report [U6]. Thus, while the absolute amounts and therelative mixture of radionuclides have changed, the doseassessment methods have not. As most of the fission-productnoble gases and the activation gas are short-lived, attention
has been focused on exposures to nearby residents.
70. When the radionuclide is uniformly distributed in theatmosphere or the photon energy is sufficiently low that
this is a reasonable approximation over the volume of aplume, then the simplest calculational method is the semiinfinite cloud model. This method assumes that theradiation from the cloud is in electronic equilibrium, sothat the energy absorbed by a given volume element equals
that emitted by the same element. For a point at groundlevel, only half the space contributes to the dose, so that theenergy absorbed is divided by two. The absorbed dose rate
in air is then given byD a  0.5

k

Da

jFE
n

Ca

i1

i

(16)
i

where D< a is the absorbed dose rate (Gy h1); Ca is theaverage activity concentration of the radionuclide in thecloud (Bq m3); ρa is the mass density of air (kg m3); Fi is
the fraction of photons of initial energy Ei (MeV) emitted
per disintegration; and k is a conversion coefficient fromenergy deposition per unit mass and unit time to absorbed
dose rate equal to 5.76 1010 Gy h1 (MeV kg1)1. A
modified version of this model, where Fi and Ei pertain tobeta emissions, is used for beta irradiation of the skin.
71. If the distribution of the activity concentration in theplume is sufficiently non-uniform to invalidate the above
approach, then a finite cloud model must be used. Such a
condition arises near the source, when persons are not inANNEX A: DOSE ASSESSMENT METHODOLOGIES

the cloud but receive dose from an overhead plume. In thismodel, the cloud is simulated by a number of small-volume
sources, and integration is performed over these sources.
The calculation proceeds by finding the photon flux
density, summing over all the decay energies for theradionuclide of interest and then converting to absorbed
dose. The basic expression for the photon fluence due tothe fraction Fi of photons of energy Ei emitted per
disintegration is [N2]

Ni 

m

X V F i B en (E i,µ i x) e

µi x

4 Bx 2

V

dV

(17)

where φi is the photon fluence; XV is the concentration of
the atoms of each radionuclide in volume element dV; µ i isthe linear attenuation coefficient, x is the distance from thevolume element dV; and Ben(Ei, µ ix) is the energy absorption build-up factor at a distance x for a radiation of initialenergy Ei, having an attenuation coefficient µ i. This
integral is evaluated numerically.
72. In general terms, the Committee has considered
0.7 Sv Gy1 to be the most appropriate average value of thequotient of effective dose rate to absorbed dose rate in air
for males and females for environmental exposures togamma rays. However, when the absorbed dose in air is theresult of a calculation such as is described in this Section,then there are sufficient data on the photon energy
spectrum to use more precise conversions. These
conversion coefficients have been derived for infants,
children, and adults by Saito et al. [S5, S11], based on asemi-infinite cloud model. These age-dependent results
have not been used by the Committee, but the energydependent variations for the adult have been incorporated
into the radionuclide-specific results [E7].
73. Based on the types of calculations indicated above,
the Committee has estimated values of the collective
effective doses from immersion exposure per unit releaseof fission noble gases and the activation gas 41Ar [U6].
These calculations are updated in Table 10. On theassumption of a semi-infinite cloud and uniform
concentrations over the mean paths of gamma rays in air,the effective dose rates to the adult per unit concentrationof the radionuclide in air, di, have been calculated [E7].
The collective dose over the local and regional areas isevaluated as follows:Si m C (x) d N 2 B x dx
a,i

(18)

i

where di is the dose factor for radionuclide i, N is thenumber of inhabitants per unit area, and x is the downwind
distance. The concentration of radionuclide i at distance x,
Ca,i(x), can be determined from equations (710). For
short-lived radionuclides, radioactive decay during thedispersal time must be taken into account. In this case theconcentration is

C a,i(x)  C a,j e 8x/u

31

(19)

where Ca,j was defined in equation (7), λ is the radioactive
decay constant (s1), and u is the wind speed (m s1) for agiven stability class. Since noble gases do not deposit, thewet and dry depletion factors Dw and Dd in equation (7) are
set equal to 1 in these calculations.
74. Analytical evaluation of the integral, equation (18),with the expression of equations (710) is not possible, so
a numerical integration is required. The results are given
in Table 10. The radionuclide releases apply to the modelsite with the meteorological conditions given in Table 4.
The release height was 30 m and the population densities
were taken to be 400 inhabitants km2 in the local area
(150 km) and 20 inhabitants km2 in the regional area(502,000 km). A similar method could be used to obtain
the immersion dose from radon released from mill tailings,
but the result is of much less significance than that due toinhalation.
75. The composition of noble gas releases fromreactors is variable, depending on the reactor type anddischarge delay features. If the composition is notknown specifically, representative compositions may be
assumed, such as used previously by the Committee andas listed in Table 11. For PWRs, the long-lived noblegas 133Xe predominates with secondary release of 135Xe.
For BWRs, the composition includes several short-lived
components. For GCRs, the noble gas release is assumed
to comprise wholly 41Ar. The dose factors derived inTable 11 to be applied in the general case to noble gas
releases are 0.11 man Sv PBq1 for PWRs, 0.43 man Sv
PBq1 for BWRs, and from Table 10 (41Ar) 0.90 man Sv
PBq1 for GCRs. Because of changes in the parametersand calculational procedure, these values are slightly
different from those previously derived [J1, U6].
76. For discharges from fuel reprocessing plants, the only
radionuclide of interest in terms of cloud dose is 85Kr [U6].
The Committee assessed the dose resulting from dischargesof 85Kr from the Windscale plant (Sellafield) between 1975
and 1979 using the methodologyprovided for the European
Community [N2]. The average annual release of 85Kr was
35 PBq and the resulting local and regional collective
absorbed dose commitments were as follows: gonads,
0.058; breast, 0.078; red bone marrow, 0.095; lungs, 0.074;
thyroid, 0.065; bone lining cells, 0.095; liver, 0.074; skin,
19; and remainder tissues, 0.078 man Gy. The collective
effective dose equivalent commitment was estimated to be
0.074 man Sv from the cloud gamma irradiation using theconversion coefficients of Poston and Snyder [P3]. A
further contribution from the beta irradiation to the skin is
0.19 man Sv, for which a skin-weighting factor of 0.01 is
applied. Thus, the normalized collective effective dose
commitment, HE + 0.01 Hskin , for this site is 0.0075
man Sv PBq1.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

32(d) Globally dispersed 85Kr
77. The model used to calculate the global collective dose
commitment from 85Kr released at fuel reprocessing plants
is given in the UNSCEAR 1982 Report [U6]. A twocompartment model similar to that proposed by Kelly et al.
[K3] is used in which the released krypton is assumed to beinstantaneously dispersed throughout the troposphere of
the northern hemisphere, which is assumed to have aheight of 10 km and a mass of 1.9 1018 kg (1 m3 of aircorresponds to 1.2 kg). Exchanges take place between the
troposphere of the two hemispheres with a half-time ofabout two years. Within a few years the 85Kr becomes
uniformly dispersed, and the sole removal mechanism isradioactive decay.
78. The whole-body absorbed dose commitment per unit
time integral of air concentration of 85Kr was estimated to
be 4.3 109 Gy (Bq a kg1)1 [N2], and the dose
commitment to the skin from the beta irradiation was
5.4 107 Gy (Bq a kg1)1. These values were restated in the
UNSCEAR 1988 Report [U4] to correspond to a collective
effective dose equivalent commitment from 85Kr of 0.17
man Sv PBq1, assuming a world population of 4 109. Thisvalue was then scaled to a value of 0.2 man Sv PBq1 forthe world population of 4.6 109 during the 19851989
period. Newer calculations [E7] indicate a value for
effective dose equivalent, HE, of 4.51 109 Sv (Bq a kg1)1
and for skin, Hskin, of 5.00 107 Sv (Bq a kg1)1, or 7.92
nSv (Bq a m3)1 (HE + 0.01 Hskin). With this slight change
and for a world population of 6 109, the normalized
effective dose commitment becomes 0.22 man Sv PBq1.

varied, these results were consistent with the loss of half of thematerial with a half-time of seven days and the other halfbeing firmly fixed on urban surfaces. This urban runoff effect
was reflected in the Chernobyl assessment in the UNSCEAR
1988 Report [U4] by applying these coefficients to that portionof a country’s population considered to be urban. Such aneffect was not considered in the Committee’s assessment ofdose from nuclear weapons fallout.

2. Methods for estimating exposures
81. The Committee has traditionallyused two approaches
to estimate the external doses that result from thedeposition of radionuclides on soil surfaces: direct
measurements and calculations based on radionuclide
deposition densities, which are the same procedures as
used to evaluate exposures from naturally occurring
radionuclides. As the calculational approach is more easily
applied and as it is not always possible to measure very low
dose rates, it is results of this approach that are moregenerally available.
(a) Atmospheric nuclear testing
82. The evaluation of radiation doses from fallout ofradionuclides onto the earth's surface following the testingof nuclear weapons in the atmosphere was one of theearliest problems to be addressed by the Committee andone that has been regularly considered. The general
method of assessing radiation doses from fallout from
nuclear tests is indicated in Figure I. Within this model theexternal effective dose commitment, Ec, for a specificradionuclide released in an atmospheric test is

D. RADIONUCLIDES DEPOSITED ON SOIL
1. Exposure processes
79. Radionuclides released to the atmosphere undergo
decay in transit or are deposited on the earth’s surface by
wet or dry deposition within relatively short periods. Therefollows a generally longer period in which the radionuclides on the terrestrial surface will eventuallydecay andproduce external radiation exposure and dose to the
population living in the areas. Radionuclides are initiallydeposited on the upper surface of the soil, but they quickly
weather into the first centimetre of soil, especially if they
are deposited via rainfall. This weathering effect and also
the fact that the soil surface is not a smooth plane (soilroughness) reduce the radiation field at the generally usedreference height of 1 m above the soil surface. Other
mechanisms, such as plowing and countermeasures, canreduce the exposure rate, but such processes have not been
considered in assessments of the Committee.
80. Followingthedeposition ofradioactive material from the
Chernobyl accident, several groups observed that the measured
external gamma exposure rate decreased more rapidly overurban surfaces than over grass surfaces [J2, K6, S7]. AlthoughE c  P 01 P 12 P 25 A 0  P 25 F

(20)

where A0 is the amount released, P01 is the integrated
concentration of a radionuclide in air at a specified location
divided by the amount released, P12 is the quotient of thedeposition densityand the integrated air concentration, andP25 is the quotient of the effective dose commitment and thedeposition density. The second part of the equation
represents a more direct method of evaluation, namely
beginning with the measured deposition density F (alsoequal to P01P12A0) and multiplying this by the transfer
coefficient P25.
83. The P25 transfer coefficients for external irradiation
have been calculated by multiplying the dose-rate
conversion coefficients for radionuclides deposited on the
ground, derived from Beck [B9], by the mean lifetime of
the radionuclide and by an average factor accounting for
airtotissue dose conversion, indoor occupancy inbuildings (80% assumed) with a shielding factor of 0.2.
The latter factor is 0.7 Sv Gy1 (effective dose rate in the
body per unit absorbed dose rate in air) times 0.36 (0.2
outdoor occupancy plus 0.8 indoor occupancy times 0.2

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

building shielding). For short-lived radionuclides (allexcept 137Cs for fallout from nuclear testing) the dose-rate
conversion coefficient applying to a plane source has beenused. For 137Cs, the dose-rate conversion coefficient applying to an exponential concentration profile in the groundof mean depth 3 cm is used. The indoor occupancy, as wellas the shielding factor, can vary a great deal among
different populations and is a source of uncertainty in thecalculations of external dose. Also, the different behaviourof radionuclides deposited in urban and rural environments
has not been taken into account for estimates of dose from
nuclear weapons fallout. This difference was, however,considered for the assessment of doses from the Chernobyl
accident (see below).
84. The P25 transfer coefficients that are used to estimate
external doses from deposited radionuclides from fallout
from nuclear testing are presented in Table 12. Transfer
coefficients for many other radionuclides can be derived
from the basic data of Beck [B9]. In earlier assessments theCommittee assumed a plane source to be appropriate forshort-lived radionuclides, however, to account for groundroughness, it is more realistic to assume an exponentiallydistributed source with a relaxation depth of 0.1, 1, and 3
cm for radionuclides of half-lives <30 days, 30100 days,
and >100 days, respectively. This change reduces the doses
by about 15%50%, but it does not have a significant
impact on the calculated total dose from nuclear weapons
fallout, which is dominated by the dose from 137Cs.
10 1

ANNUAL EFFECTIVE DOSE (nSv)

137Cs

10 0

125Sb

10 -1

10 -2
54Mn
95Zr

10 -3
1

2

144

Ce

3
4
5
6
7
8
YEARS FOLLOWING DEPOSITION

106

Ru

9

10

Figure III. External exposure following unit deposition(1 Bq m -2) of radionuclides.

85. An indication of annual contributions to doses fromexternal exposure following a single deposition event may
be of interest, although the analysis is simple, as it depends
only on the radioactive decay of the radionuclides. The
time course of contributions to dose from unit deposition
density of the radionuclides is illustrated in Figure III, andthe annual average doses are listed in Table 13. Several
short-lived radionuclides (131I, 140Ba, 141Ce and 103Ru) makeno contributions to external exposure beyond the first year
following deposition. The values in Table 13 have been

33

calculated from decayed monthly deposition density,
averaged over the year and multiplied by the absorbed dose
rate in air per unit deposition density (Table 12, column 3)
and by the shielding/occupancy factor of 0.36 and theconversion factor 0.7 Sv Gy1. The sum of the annual
contributions to dose is equal to the dose commitment.
(b) The Chernobyl accident
86. The methods used to calculate external doses causedby the Chernobyl accident were basically those applied toestimate the external doses from radionuclides produced in
atmospheric nuclear testing, although several modifications were introduced to account for the shorter term of therelease, urban-rural differences, and an improved
assessment of the movement of radionuclides into soil. The
results of calculations of doses from the Chernobyl accident
were presented in the UNSCEAR 1988 Report [U4].
During the first month after deposition, a number of shortlived emitters, including 132Te, 132I, 131I, 140Ba, 140La, and136
Cs, were important components of the total external
gamma exposure rate (or dose rate in air). For several
months, 103Ru and 106Ru made contributions, but since then
only 134Cs and 137Cs have been of significance. Exposure
from 137Cs remains significant for several years and must
be projected into the future.
(i) First month
87. The outdoor exposure X1 (C kg1) during the firstmonth was assessed by four different methods, with the
choice dependent on the data available. If continuous or
daily data were provided, the exposure rates were
integrated. If incomplete data were provided, an attempt
was made to fit a power function of the form atb to the data,
where t is time (days) and a and b are constants to be
determined. X1 is then the integral of this function fromarrival day 1 to day 30.
88. If measurements of external gamma exposure rate
were not available, two approaches were used. If data on
the ground deposition of the radionuclides were provided,
the exposure rate from each radionuclide was computed
using the coefficients published by Beck [B9] for arelaxation depth of 1 mm to account for surface roughness.
In several cases only data on the deposition of 137Cs wereavailable, and X1 was evaluated on the basis of therelationship of the exposure to 137Cs deposition density as
measured at a specific location, e.g. Neuherberg, Germany[G2].
89. The effective dose during the first month, Ee1, (Sv)was calculated from X1 by:
E e1  A X 1 (1  F 0)  A X 1 F 0 F s(21)

where A is the conversion coefficient (23.6 Sv per C kg1,
i.e. 33.7 Gy per C kg1 × 0.7 Sv Gy1), F0 is the indoor
occupancy factor, and Fs is the building shielding factor.
The last two values were taken as 0.8 and 0.2, respectively.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES34(ii) One month to one year
90. The calculation of external gamma dose beyond one
month was based on the measured total deposition of 134Cs
and 137Cs and, although less important, 103Ru, 106Ru, and
131
I. The conversion coefficients for long-term deposition todose rate depend on the penetration of these radionuclides
into soil. Change with time is accounted for by using
coefficients appropriate for a relaxation depth of 1 cm
during the first year and 3 cm thereafter. Also, the effect ofmore rapid removal of radionuclides from urban surfaces
was considered.
91. The equation for the calculation of external gamma
effective dose, Ee2, (Sv) for the time period between one
month and one year for radionuclide i is as follows:E e2,i Fi

8id e2,i (e8i m /12

 e 8i m )distance from the release point. With use of an effective
deposition velocity that accounts for both wet and dry
deposition, the deposition densities of radionuclides maybe
estimated. This method is appropriate for routine
continuous and near-surface releases from sources such as
nuclear installations. The local area of exposure is taken to
be 150 km surrounding the point of release, and theregional area extends to 2,000 km.
94. In the dispersion estimation method, equation (11),an average dilution factor is assigned at 1 km, namely
5 107 Bq m3 per Bq s1 released, and further dispersion
reduces the radionuclide concentration in air in inverse
proportionality to the 1.4 power of the distance. The airconcentration may be related to the deposition density bymultiplying by the effective deposition velocity. The
general formula for application of the transfer factor
method is(22)

1  F 0 (1  F s ) 1  F p (1  F u )Si50

 5 107 v

g P25

N1

x 1.4 2 π x dx

1

(24)

2000

where Fi is the deposition density (Bq m2); de2,i is thedeposition density to effective dose conversion coefficientduring the period between one month and one year
(relaxation depth of 1 cm) (Sv per Bq m2); λi is theradioactivity decay constant (a1); m is a constant equal to
one year; Fp is the urban fraction of a country’s population;
Fu is the fraction of the deposition that remains fixed on
urban surfaces (assumed to be equal to 0.5); and F0 and Fs
are as defined previously. Effective dose equivalent
conversion coefficients are listed in Table 14.
(iii) Periods beyond one year
92. External effective dose, Ee3,i, (Sv) for periods beyond
one year were evaluated according to the equation
E e3,i Fi

8id e3,i e

8im

(23)

1  F 0 (1  F s ) 1  F p (1  F u )where de3,i is the deposition density to effective conversion
coefficient for periods greater than one year. Thiscoefficient is based on a relaxation depth of 3 cm. Valuesof this coefficient are also listed in Table 14.
(c) Nuclear installations
93. Releases from nuclear installations of radionuclides
that contribute to external exposures are, in general, too
low to be measured in air or deposition at distances beyond
the installation site and point of release. As was discussed
in Section I.B.3, long-term average dispersion of radionuclides in air may be estimated using a formulation thatcombines a dilution factor at 1 km and a power function ofx 1.4 2 π x dx

 N2
50

where Si is the collective effective dose per unit release ofradionuclide i (man Sv Bq1); vg is the effective deposition
velocity (m s1); P25 is the transfer factor from deposition
density to dose (Sv per Bq m2); N1 is the population density
in the local area (inhabitants km2); N2 is the population
density in the regional area (inhabitants km2); and x is the
distance from the point of release (km). The parameter x1.4
should actually be expressed as (x/1 km)1.4 to rectify theunits. The quantity in brackets has the unit number of
persons. The population densities applied are those assumed
for model reactor site: N1 = 400 inhabitants km2 and N2 = 20
inhabitants km2. The value of the effective deposition
velocity is taken to be 0.002 m s1 for annual averagedeposition, which is the value for dry deposition alone. In
reality, more material is deposited under wet conditions thanunder dry, and an effective deposition velocity for point
sources that includes both wet and dry contributions would
range from 0.005 to 0.013 m s1, depending on downwind
distance. However, use of a larger value in conjunction withthe power law expression for the air concentration (equation
11) results in a greater estimated activity amount deposited in
the local and regional areas than was released to theatmosphere. The most probable explanation for this is that
equation 11 overestimates air concentrations, for the reason
given in paragraph 41. Although larger (more negative)
values of the exponent in the power function could be selected
to offset a higher effective deposition velocity, the values of
1.4 and 0.002 m s1 preserve the mass balance to distances of
2,000 km and ensure that doses from airborne and deposited
activity are not underestimated. Therefore, these values areused on local and regional scales for purposes of estimating
average deposition.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

95. Estimates of normalized collective effective doses
from external exposure from radionuclides released as
particles in airborne effluents from reactors are listed inTable 15. The transfer coefficients P25 from deposition
density to effective dose were derived from the basic dataof Beck [B9]. The collective effective doses per unit release
were then estimated using equation (24). Other fission andactivation products could be added to Table 15 by applying
this method.
96. A representative composition of radionuclides inparticulates released in airborne effluents is not easily
established because of the large number and varying
amounts of radionuclides that may be involved. An earlier
approach of the Committee [U6] was to assume equal
activity distribution across 18 radionuclides that were35

commonly reported to be present. A slight variation would
be to recognize two groups of radionuclides, one of more
dominant contributors to the total activity release and asecondary group. Consistent with reported data [U5, U6] isto assume 90% of the activity of release present as 54Mn,
58
Co, 60Co, 89Sr , 134Cs, 137Cs, and 140Ba (including 140La).
The radionuclides of the second group, contributing 10%of the activity total, are 51Cr, 59Fe, 65Zn, 90Sr, 90Y, 95Zr
(including 95Nb), 124Sb, 136Cs, 141Ce, and 144Ce. Assuming
equal contributions to the activity release within eachgroup, weighted average values of the local and regional
collective effective doses per unit total (representative)
release of particulates are obtained. These results areincluded in Table 15. Adjusted weightings could be madein specific circumstances, if the exact composition of therelease is known.

III. INHALATION EXPOSURE
97. There are two main processes that contribute tointernal exposure, the general term used to describe
exposures that involve the intake of radionuclides into thebody as opposed to external exposure, which is consideredin Chapter II above. The two processes are inhalation ofcontaminated air and ingestion of contaminated foodstuffs.
For inhalation, if the time dependent concentration of aradionuclide in air is known, it is a straightforward matterto calculate committed dose by multiplying by a breathing
rate and by a dose-conversion coefficient. The ingestion
pathway involves additional steps of transfer to plants andanimals, from which are derived the foods consumed byhumans. For convenience, inhalation exposures are
considered in this Chapter and ingestion exposure in thefollowing Chapter.

[I4] are indicated in Table 16. An indication is also given
in Table 16 of the fraction of the population in each of thesix age categories and the age-weighted average breathing
rate. The age-weighted value corresponds to 19 m3 d1.
Considering the uncertainty of the age distribution of the
population and the differences between countries, arounded value of the nominal breathing rate of 20 m3 d1would seem to be appropriate for use in most applications.
In assessments of the Committee, the population groups
specified as infants, children, and adults are assumed tocorrespond with the ICRP age categories of 12 years,
812 years, and >17 years, respectively.

A. NATURAL RADIONUCLIDES
98. Many of the Committee’s past calculations of
inhalation doses were performed using a nominal breathing
rate of 20 m3 d1, or 7,300 m3 a1. This generally reflects
the concern of the Committee with the collective dose,
which is substantially determined by the intake of theadults in the population. For calculating inhalation doses
from the Chernobyl accident, inhalation rates of 22 m3 d1
for adults and 3.8 m3 d1 for infants were used [U4]. The
latter values are the same as those used for naturally
occurring radionuclides and are derived from the same
source [I7].
99. The Committee has generally used the dosecoefficients published by ICRP for its evaluations. Initiallysuch values were available only for adult workers, but
starting in 1989 age-dependent values have been madeavailable for members of the general public. The latest
compilation of values for both ingestion and inhalation isprovided in [I5]. The breathing rates now used by ICRP

100. Naturally occurring radionuclides are present in theatmosphere owing to their production by cosmic ray
interactions, the emanation of gases from soil or building
materials and the resuspension of soil particles from theground surface. The main cosmogenic radionuclides, 3H
and 14C, are fairly uniformly dispersed in the atmosphere.
Inhalation exposures from these radionuclides are,however, almost completely negligible compared with theingestion exposures.
101. Soil-derived radionuclides are present in air invariable amounts, depending on local soil, wind, andmoisture conditions. In earlier assessments by UNSCEAR[U6, U7], a dust loading of 50 µg m3 was assumed andapplied to typical concentrations of natural radionuclides
in soil. Some portion of the solid matter in air may notcome from the soil, however, but from organic matter,building dusts, smoke, and fly ash from coal burning.

36

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

102. A very important contribution to inhalation exposure
is made by radon and its decay products. The gas emanatesfrom soil and can enter and attain high concentrations inindoor spaces. Because this exposure component dominatesthat from all other pathways, it is important that thedosimetry for radon be well established.
103. The ICRP has not provided values of the doses perunit intake for 222Rn and 220Rn and their decay products
from application of the respiratory tract model [I4, I5], andthe dosimetry for these mixtures is very complex. Becauselung cancer has been observed and studied extensively inminers exposed to 222Rn, the ICRP [I13] has adopted aconversion convention for radon exposures that is based onequality of detriments from epidemiologicaldeterminations. The detriment per unit effective dose formembers of the public is 7.3 105 per mSv, and the
detriment (to miners) per unit exposure to 222Rn progeny is
8.0 105 per (mJ h m3). Thus, an exposure to 222Rn
progeny of 1 mJ h m3 is equivalent to an effective dose of
1.10 mSv. As 1 mJ h m3 is equal to 1.80 105 Bq h m3 of222
Rn in equilibrium with its short-lived progeny, a dosecoefficient of 6.1 nSv per (Bq h m3) can be derived andapplied to equivalent equilibrium concentrations (theactivity concentration of radon, in equilibrium with its
short-lived progeny, which would have the same potentialalpha energyconcentration as the existing non-equilibrium
mixture). The dosimetric evaluations give dose coefficientsin the range 615 nSv (Bq h m3)1. The value previously
used by the Committee in earlier evaluations [U3, U4],
9 nSv (Bq h m3)1, is within this range and would seem to
be still appropriate for use in dose evaluations. An
epidemiologically based conversion convention is notavailable for 220Rn. However, by analogy with the risk
determined for 222Rn and by comparing the dosecoefficients for 220Rn and 222Rn calculated on a dosimetricbasis [I18], a dose-conversion convention of 40 nSv per(Bq h m 3) equilibrium equivalent concentration of 220Rn
can be derived; this value is intended to include the dose toorgans other than lung due to the transfer of 212Pb from thelung. The half-life of 212Pb is sufficiently long (10.64 h) for
this effect to be significant, whereas none of the short-lived
progeny of 222Rn is sufficiently long-lived to merit similar
consideration.

B. RADIONUCLIDES RELEASED TO THE
ATMOSPHERE
104. In its various assessments, the Committee has used
the best available estimates of dose per unit intake ofradionuclides by inhalation; whenever possible, the values
provided by ICRP have been used. The ICRP values havebeen updated [I4, I5] based on a revision to the ICRP
model of the respiratory tract [I6], and age-dependent
values for the general public are now provided. The values
for radionuclides used by the Committee in its assessments
are given in Table 17.

1. Exposure processes
105. Inhalation of radionuclides in air can result from ashort-term or continuous release processes. Inhalation israrely the primary pathway of exposure if radionuclides are
released to the atmosphere, but there are some notable
exceptions. The importance of radon and its decayproducts
was mentioned in the preceding Section. Anotherexception involves radionuclides of extremely low
biological availability. Such radionuclides pass readily
through the gut following ingestion intake, but they can bedeposited in the lungs following inhalation intake and beretained for long times. The most notable example of sucha radionuclide is 239,240Pu.
2. Methods for estimating exposures(a) Atmospheric nuclear testing
106. According to the general model developed by theCommittee to describe environmental transport processes,
the equation for committed effective dose, Ec, (Sv) via
inhalation isE c  P 01 P 14 P 45 A 0  P 245 F(25)where P01A0 is the integrated air concentration (Bq a m3),
P14 is the breathing rate (m 3 a 1) and P45 is the doseconversion coefficient (Sv Bq1) for inhalation. To
determine the integrated air concentration, measurements
must be made for the entire time that radionuclides remain
in air. Since this is not always achieved in practice, thesecond part of the equation is the more common approach,in which the integrated air concentration is estimated fromthe deposition density, F. In this case, the transfer
coefficient for the inhalation pathway is determined as P245
= P14P45/P12.
107. The average value of P12, which is also the effective
deposition velocity, varies with the precipitation rate atdifferent locations and also with the chemical and physicalnature of the radionuclide considered. The average valueof P12 for particulate material deposited following
atmospheric nuclear testing has been estimated to be
1.76 cm s1, or 5.56 105 m a1 [B2]. Although this value isbased on observations in New York City over several years,
measurements in the United Kingdom [C7] and Sweden[B10, D5] are in reasonable agreement after normalization
to the same annual precipitation. Furthermore, since theannual rainfall in New York City is fairly close to the
population-weighted average for the whole world, the New
York value is considered adequate for global averagecalculations.
108. Values of the transfer coefficient, P245, for theinhalation pathway are listed in Table 18. These update thelisting in the UNSCEAR 1993 Report (Table 8, page 127[U3]). The values are for the adult with a breathing rate ofANNEX A: DOSE ASSESSMENT METHODOLOGIES

7,300 m3 a1 and P45 values from Table 17. These transfer
coefficients are applicable to the release and deposition
conditions of radionuclides in fallout from nuclear tests.
(b) The Chernobyl accident
109. For the Chernobyl accident assessment, a somewhat
modified approach was used to account for a filtration
effect that reduces the concentrations of radionuclides inindoor air [U4]. The calculation of the inhalationcommitted effective dose, Eh,i (Sv) for radionuclide i was asfollows:
E h,i  C a,i B d h,i (1  F 0 )  C a,i B d h,i F 0 F r(26)

where C*a,i is the integrated activity concentration ofradionuclide i in outdoor air, B is the breathing rate, dh,i isthe committed dose per unit intake from inhalation, F0 is
the indoor occupancy factor and Fr is the ratio of indoor to
outdoor air concentration. The latter parameter was
assigned a value of 0.3 for all countries [C9, R2, U4].
110. If the integrated concentration in air is known, thenthe calculation is very simple as indicated above.
Furthermore, if an average concentration over a one-year
period is known, then the calculation is also quite
straightforward. It is, however, rather rare thatmeasurements of integrated activity in air are available
following accidental releases, especially over a short periodof time. In that case the integrated concentration in air isusually estimated on the basis of the deposition density fora particular radionuclide and the effective deposition
velocity, as mentioned above. The deposition density
divided by the deposition velocity gives the integrated
concentration in air.
111. If the relative amounts of the radionuclides released
at the time are known and if these releases are concurrent,
then the measurement of the deposition density for only
one radionuclide in the mixture can be consideredsufficient to define the deposition densities of all radio-

37

nuclides at the time of deposition, if the deposition
velocities of the radioelements do not differ significantly.
In fact, measurements of the deposition density of a longlived radionuclide can be made many years after thedeposition occurred and used to define the original
deposition densities of all radionuclides, provided that thesoil is undisturbed and the sampling is deep enough toencompass all of the original deposition.
112. Other methods can be used to define the deposition
densities and the integrated air concentrations ofradionuclides. Although subject to more error and in needof more sophisticated interpretation, measurements of theexternal gamma-dose rate in air, of concentrations ofradionuclides in foodstuffs, and even of radionuclides inpeople can be used to estimate the original deposition
densities and integrated air concentrations.
(c) Nuclear installations
113. Estimates of inhalation exposure from releases ofradionuclides from nuclear installations may be madeusing the dispersion model presented in Section I.B.3 andthe transfer coefficients P245. The results of this calculation
are listed in Table 19. These estimates apply to longer-term
releases, as the meteorological conditions for the representative site have been averaged over an annual period. The
deposition velocity appropriate for near-surface releases of
0.002 m s1 has been used. This is determined mainly by
dry deposition, since precipitation can be expected to occur
only during a small fraction of the time of plume passage.
114. As discussed above with regard to external exposure(paragraph 96), a representative composition of radionuclides
in particulates released in airborne effluents from reactors may
be assumed. A weighted average of the collective dose frominhalation exposure per unit release of particulates may thenbe derived for general application. The values pertaining to
the local and regional areas are included in Table 19. The
transuranium radionuclides are not normally reported inroutine releases from reactors, however for reference
purposes, the values are included in Table 19.

IV. INGESTION EXPOSURE
115. Ingestion exposure occurs when radionuclides in theenvironment enter food chains. This component and thatof external exposure are usually the significant andcontinuing sources of exposure following releases ofradionuclides to the environment. Radionuclides released tothe atmosphere may deposit onto both terrestrial and aquatic
surfaces, for which different calculational methods are
required. The terrestrial and aquatic food pathways are
considered in separate Sections of this Chapter.

116. Ingestion exposures have been evaluated by
UNSCEAR for natural radionuclides present in theenvironment and for several cases of radionuclide releaseto the environment, including atmospheric testing, releases
from nuclear fuel cycle installations and the Chernobyl
accident. For the most part, annual average values havebeen considered with the aim of evaluating committed
exposures. This is adequate for longer-term or continuous
releases. Short-term releases at particular times, such as

38

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

was the case for the Chernobyl accident, require taking into
account some seasonal variations.

A. NATURAL RADIONUCLIDES
117. In the general case, doses from the ingestion ofnatural radionuclides in foods and drinking water havebeen estimated from measured concentrations of theradionuclides in body tissues or organs. For 40K, metabolicbalance maintains body levels irrespective of intake
amounts. For uranium- and thorium- series radionuclides,
however, this is not the case, and the concentrations infoods, water and total diet have been useful for determining
geographic variations in the body burdens.
118. Beginning with the UNSCEAR 1993 Report [U3],representative dietaryintakes of natural radionuclides werecompiled; these could be used with age-dependent
estimates of dose per unit intake to extend the limited dataon tissue concentrations and to obtain more broadly baseddose estimates.
119. Estimates of dose per unit intake of radionuclides areprovided by the ICRP [I5]. These are the committed
effective doses to age 70 years, based on recent metabolic
data and models. The values used in UNSCEAR
assessments are summarized in Table 20. The age
categories are infants (12 years), children (>7 years to 12
years), and adult (>17 years). Values for age categories
from 0 to 1 year, >2 years to 7 years, and >12 years to 17
years are also provided by ICRP [I5].

B. RADIONUCLIDES RELEASED TO THETERRESTRIAL ENVIRONMENT
120. An extensive database of deposition and dietmeasurements from the years when there was atmospherictesting has allowed empirical relationships to be derived toevaluate transfer coefficients for radionuclides released inthis practice. Empirical models describing the time courseof annual transfers from deposition to diet and from diet tothe body have been the basis of the Committee’s
evaluations of doses from 90Sr and 137Cs, and this method
was also applied to transuranic radionuclides. Fewer data
have been available from which to derive ingestion
pathway transfer coefficients for 131I, 140Ba, 89Sr and 55Fe.
1. Transfer processes
121. Plants are the primary recipients of radioactive
contamination to the food chain following atmosphericreleases of radionuclides. Vegetation may be subject todirect and indirect contamination. The directcontamination of terrestrial vegetation refers to thedeposition of radioactive materials from the atmosphere
onto the above-ground parts of plants. Indirectcontamination refers to the sorption of radionuclides fromthe soil by the root system of plants. Secondary recipientsof food chain contamination are animals that consume
plants or other animals. Both plant and animal products
enter the diet of humans.
(a) Direct deposition on plants
122. Direct deposition on plants may play an important rolein the contamination of plant products for someradionuclides, including those characterized by low root
uptake and short-lived radionuclides, especially 131I, that cantransfer relatively rapidly through the food chain. The directcontamination of plants may be of two types: primary, which
involves direct transfer from the source via the atmosphere to
the plants, and secondary, by which activity already depositedon the ground may be resuspended, e.g. by the wind, and thustransferred to the plants. The resuspension process is not
usually a substantial factor, except for radionuclides with verysmall uptake through the roots. Primary direct deposition
involves three processes: deposition, interception andretention. Direct contamination of the plants depends on thedevelopment stage of the plants at the time of contamination.
This, in turn, depends on the season of the year when thecontamination occurs.
123. Radionuclides in the atmosphere may be deposited aseither dry or as wet deposition. Dry deposition occurs
continuously, while wet deposition occurs when rain orsome other form of precipitation intervenes. Dry deposition
is usually described by applying the deposition velocity, vg= F/C [C1], where F is the fallout rate of the depositingradionuclide to a unit area of land (Bq m2 s1), and C isthe concentration in ground-level air over the area of landconsidered (Bq m3). The unit of vg is thus m s1. The
deposition velocity varies with the aerodynamic diameterof the particles deposited. Particles with a diameter
between 0.1 and 1 µm have a deposition velocity of about
0.02 cm s1; those between 1 and 10 µm have values
ranging from 0.02 to about 5 cm s1 [H1]. This magnitude
also varies with the type of surface and with the chemicaland physical characteristics of the radioelements involved.
124. Wet deposition occurs during precipitation. The
wash-out ratio, W, is defined as the ratio of theradionuclide concentrations in precipitation (Bq l1) and inground-level air (Bq m3) [E3]. Experience from global
fallout studies has shown that around 90% of the total
deposition of 90Sr and 137Cs occurs as wet deposition. In an
accident, most of the deposition usually takes place within a
few days. The Chernobyl accident demonstrated that highrainfall during the cloud passage results in deposition rates anorder of magnitude higher than those observed for dry
conditions [E5].
125. Interception is the fractional deposition of radionuclides on the plant surfaces. It depends on both thephysical characteristics of the deposit and the growth formof the plants. The subsequent fate of the deposit, i.e. theANNEX A: DOSE ASSESSMENT METHODOLOGIES

retention, is influenced by these factors and by the rate at
which the material is removed by precipitation and otherprocesses, called weathering or field loss.
126. The fraction of material intercepted by the crop
canopy was studied by Chamberlain [C2], who derived anempirical parameter dependent on the physico-chemical
properties of the deposit, the manner of deposition, themorphology of the crop and the meteorological conditions.
The quotient of the fraction retained and the dry weightbiomass usually falls within the range 0.24 m2 kg1 [C3].
The normalized specific activity is defined as theconcentration in the crop (Bq kg1 dry weight) divided by
the deposition density rate (Bq d1 m2) [C2]. The
normalized specific activity is thus a rate factor with theunit m2 kg1 d. Values between 20 and 40 m2 kg1 d havebeen observed for 137Cs and 90Sr for herbage in good
growing conditions [E4]. Chamberlain found that wintergrass had normalized specific-activity values 23 times
higher than summer grass.
127. The weathering or field loss is expressed by M/M0 =et/τ, where M0 and M are the quantities retained on thecrop initially and after time t and τ is an empiricalconstant. During the growing season, τ is about two weeks;in the winter period, it increases to about eight weeks.
When there is rain, the field half-time may be short.
128. Resuspension of radionuclides on the soil surface may
result in secondary direct contamination of the crops. The
resuspension factor RF is defined as the radionuclide
concentration in air (Bq m3) divided by the groundcontamination (Bq m2). The resuspension factor thus has
the unit m1. The resuspension factor measured at locationsin Denmark for 1003,000 days after the Chernobyl
accident decreased according to a power function of thetime, t, in days (RF = 9.3 106 t 1.17) [A3].
129. It appears that resuspended 137Cs is less available to theplant than primarily deposited amounts [A1], i.e. the transfer
factor for primary direct contamination is higher than that forsecondarydirect contamination. There may be two reasons forthe lower availability of resuspended particles compared withdirectly deposited fallout. First, a higher field loss can be
expected for resuspended particles than for global fallout.
Secondly, 137Cs adheres to minerals, especially clay, allowing
the radiocaesium to be less available for absorption by thecrops and thus for translocation to the grain.
130. A special case of secondary direct contamination ofcrops is rain splash, which may occur during heavy showers,
when the recoil from rain drops carries contaminated soil tothe surface of the vegetation. Secondary contamination is
expected to be less efficient with respect to translocation to theplants than the initial, direct contamination route.
131. Seasonal variation in direct contamination is ofparticular importance for cereals. This feature was firststudied by Middleton [M2]. It appears that the two

39

important factors influencing contamination of grain are
the initial retention and the translocation from thevegetative part of the seeds. Initial retention is largely
independent of the radionuclide, whereas translocation
depends strongly on the radioelement and its solubility.
132. Time of year was observed to affect the transfer factorof 137Cs to grain at the time of the Chernobyl accident [U4].
Transfer factors were higher in southern Europe, where thecrops were more developed when the deposition from
Chernobyl occurred than in northern Europe, where thegrowth of crops had not yet begun. Seasonality also
affected total diet intakes.
(b) Root uptake
133. In the first period after a radioactive contamination
event, direct deposition on plant surfaces is the dominant
pathway, but in the long term, the contamination of thehuman diet will depend on absorption through the roots ofplants. The extent to which plant roots absorb radionuclides from the soil depends not only on their physiology
but also on processes in the soil.
134. The uptake of radionuclides by plants from the soil isnormally described by the transfer factor Bv, the ratio ofradionuclide concentrations in vegetation and soil (Bq kg1
dry weight plant to Bq kg1 dry weight soil). Observed
values of Bv vary widely, mainly as a result of different soil
and vegetation types and environmental conditions. In
addition, management practices such as ploughing, liming,
fertilization and irrigation greatlyaffect uptake. Variability
can also result if uptake into the whole plant is comparedwith uptake into parts of the plant, such as grain. The
transfer factor Bv is not constant in time. Decreases occur
as radionuclides in soil become less available to plants
through changes in physical or chemical forms or inmoving below the rooting zone. In some cases, the rate of
uptake increases in time, when physical weathering ortransformation of the chemical form takes place or when
the radionuclide reaches an optimum depth for root uptake.
Databases for root uptake transfer parameters have been
published [I9, N7].
135. The main soil characteristics affecting the transfer ofradionuclides from soil to plants through root uptake are:clay and organic matter content, pH and cation exchange
capacity. These soil characteristics interact causing
variability in the transfer in different circumstances, so thatgeneralizations are not always valid. A high clay content
in the soil provides binding for caesium and reduces root
uptake. A high organic matter content often enhances theroot uptake of caesium but may also have the opposite
effect; an excess of potassium dilutes caesium ions, which
decreases uptake, but may also cause the desorption of
fixed caesium, which increases uptake.
136. The root uptake of 137Cs usually decreases with time,in the beginning quite rapidly, later more slowly. The

40

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

decrease is seen particularly in clay soils and is due to thefixation of caesium by clay minerals such as illite andvermiculite. In organic soils the decrease is mainly due toredistribution of caesium within and transport out of therooting zone. Strontium is less firmly fixed to the soil
matrix and is thus more available for root uptake thancaesium. The higher mobility of 90Sr also means that thisradionuclide migrates faster than 137Cs through the soilcolumn. Nevertheless, root uptake of 90Sr generallyremains
significantly greater than that of 137Cs over periods ofseveral years, and for terrestrially produced foods agenerally increasing 90Sr/137Cs ratio will occur with timeafter deposition.

(d) Losses in food preparation

137. Under special circumstances the root uptake,especially of 90Sr, may increase with time after contamination. This has been observed, for example, in the near zone
around the Chernobyl reactor where some of the 90Sr wasimbedded in uranium fuel particles. Weathering
throughout the years has dissolved these fuel particles,
making the 90Sr available for delayed root uptake by plants.

142. Food-processing retention factors, i.e. the fractionalamount of the radionuclide remaining in the food after
processing, are quite variable, depending on the food and theprocessing procedure. Drying foods increases theconcentrations in the dried products, typically by a factor of 5compared with the fresh foods. Boiling meat considerably
reduces the radionuclide content. It should, however, be kept
in mind that some of the water used for the boiling may beconsumed as soup or sauce. In dairy products, radionuclides
are retained less in cream, thus affecting the levels in various
milk products. Radionuclide contents in vegetables and fruits
are also significantly affected by washing, peeling, andcooking. In particular, the reduction of 137Cs by various
treatments is significant. If crops have been contaminatedonly by direct deposition, the effect of washing and peelingwill be even higher, because the contamination in that case isconfined to the outer parts of the crops. Some translocation
may eventually take place.

(c) Animal pathways
138. Several important pathways for the transfer of radionuclides to the diet of humans involve animal food chains,
including milk and eggs from living animals and meat or
flesh from animals and fish. Depending on the radionuclide and the metabolism in the organism, theconcentrations may be enhanced or reduced compared withthe earlier steps of the food chain. Some parts of the animal
are not consumed, e.g. bones, shells, skin and feathers, andthis prevents the transfer from animal products of boneseeking radionuclides such as 90Sr and plutonium. Bone tissue
might, however, re-enter the food chain as bonemeal invarious fodder products, and it might also appear in fertilizers.
139. The main animal pathway to humans of theradiologicallyimportant radionuclides such as 90Sr, 131I and137
Cs is milk consumption. All three radionuclides arereadily transferred from animal fodder to the milk. Other
radionuclides such as the transuranic elements are absent
or secreted to only a very small extent in milk. Caesium istransferred with its chemical congener potassium to thesoft tissues of animals, particularly muscle. Strontium ispreferentially transferred to bone, like its congener
calcium.
140. Fish and shellfish receive radionuclides both directlyfrom the water and from their food. Some radionuclides
that are of no concern in the terrestrial animal food chains
may be concentrated in aquatic animals. This is the case,for example, for plutonium, which is concentrated incrustaceans, and for polonium in fish and seafood. A
substantial part of the marine fish catch is used for making
fishmeal, which is used as fodder for pigs and poultry andfor fish produced in fish farms. In this way, marine
pathways may interact with terrestrial and freshwater
animal food chains.

141. Knowledge of the effects of processing and culinary
preparation on radionuclide contents in foods is needed
when assessing the radiation dose to humans from theingestion of contaminated foodstuffs. Appropriate
allowances might be made for the reductions brought about
by food processing to ensure that doses are notsystematically overestimated [N1]. However, theCommittee has not specifically considered this for its
calculations. In some cases, losses via food processing are
considered implicitly, if the assessment is based upon
nuclide content in people.

143. The process of milling cereal grains apportions theradionuclide content of the whole grain to significantly
lower radionuclide concentrations in the flour andcorrespondingly higher concentrations in the bran. The
intakes of 137Cs and particularly 90Sr are thus higher for
consumers of wholemeal bread than for consumers of whitebread. The concentrations of 90Sr and 137Cs in white breadare 20% and 40% of the concentrations in the wheat,
respectively. In rye bread, the percentage is 75% of that in
the grain for both radionuclides. There is essentially notransfer of 90Sr or 137Cs to alcohol from grain or potatoes
nor to sugar made from beets.
144. Conversion of foods, e.g. milk to cheese, may alsochange the radionuclide concentrations. The concentrationof 90Sr in cheese is thus typically 510 times higher than inmilk, while the concentration of 137Cs in normal cheese is
only about 70% of that in milk. Butter contains essentiallynone of the 90Sr and 137Cs present in the milk.
145. Assessments by UNSCEAR have not specificallyaccounted for losses in food preparation. Rather, it has
been assumed that dietary intake estimates reflect actual
amounts in prepared and consumed foods. When it appears
that this is not the case, the intake estimates will need to be
adjusted.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES(e) Behaviour of tritium and carbon-14
146. The radionuclides tritium and 14C require special
consideration because of their high mobility in theenvironment and the fundamental nature of hydrogen andcarbon cycles in the biosphere.
147. Carbon is highly mobile and is distributed throughout
the environment. A small fractionation effect reduces environmental concentrations of 14C byabout 5% relative to stable 12C,
but this difference is usually disregarded in the models.
Carbon-14 released into the environment via the atmosphere
enters the carbon cycle and becomes dispersed in theatmosphere, terrestrial biosphere, and more gradually into theocean, ocean sediment and sedimentary rocks. Much of the
carbon in plants has a short residence time, although carbon isheld longer in woody plant parts and is released only on
decomposition. Turnover time of carbon in humans isgenerally of the order of a few days or weeks. The mostimportant form of carbon from the point of view of dose isCO2, since this is the form in which carbon becomes bound inplants and ingestion contributes 99% of the dose from 14C. The
remaining fraction of dose comes from inhalation of 14C in air.
148. Tritium released to the atmosphere occurs in two
forms: tritiated hydrogen (HT) and tritiated water vapour(HTO). HTO is subject to the same wet and dry deposition
processes as other nuclides, but it can also diffuse into thesoil pore space and the leaf stomates [B7, G1]. If the HTO
gradient is reversed, however, (for example, if a wind shiftblows the plume away), tritium will rapidly be lost fromthe soil and plants to the atmosphere by evaporation andtranspiration, generating a secondaryairborne HTO plume.
HT can diffuse into the soil and be converted to HTO by an
enzyme-mediated reaction [D2, T2]. Tritium not returnedto the atmosphere by evaporation moves through the soil
primarily by the mass flow of liquid water.
149. Like other radionuclides, tritium enters plants via
root uptake. Under steady conditions, the concentration in
the plant lies between the concentrations in the soil and theair, with a magnitude that depends on atmospherichumidity and the air/leaf temperature difference [M4].
Some of the tritium that enters plants can be incorporated
into organic compounds to form organically bound tritium(OBT) [D3]. Tritium bonded to carbon forms nonexchangeable OBT, which has a much longer retention
time in plants and animals than HTO and so can contributesignificantly to the total dose. Organically bound tritium
makes up only a small percentage of the total tritium
activity in most plants, but up to 90% in grains, whichhave a high organic content.
150. Tritium is taken into the bodies of animals (including
humans) bythe normal mechanisms, and HTO equilibrates
with body fluids within minutes. For the most part, theretention time of tritium in the body is about 10 days,
although for the organically bound form it increases toabout 40 days [T3].

41

2. Food and water consumption
151. The consumption of foods and water by individuals
varies widelyaround the world, depending on climate, foodavailability and cultural dietary preferences. Locally
produced or gathered foods are now usually greatly
supplemented by foods imported from other regions orcountries. Moreover, it is difficult to obtain accurate
estimates of food consumption: there are considerable
individual variations, and many foods are of a seasonal
nature. Average rates in countries may be indicated by foodbalance analysis, taking into account local production,
imports, and exports [F1]. These will be overestimates,
however, if losses from wastage or preparation are nottaken into account.
152. When UNSCEAR has needed dietary intake
information, it has used values reported from a few
countries. For example, the analysis of fallout 90Sr transfer
to humans was based on measurements in Argentina,
Denmark, and New York City. For lack of more extensive
data, these results were averaged and assumed to begenerally applicable. Milk consumption has been reported
for many other locations. For general assessment purposes,
the Committee has used an average dietary intake of
500 kg a1.
153. For the analysis of exposures following the Chernobyl
accident, the Committee compiled consumption data for all
countries reporting first-year measurements. These values
were as assessed by scientists of the various countries or,secondarily, derived from food balance considerations. The
consumption rates, as given in the UNSCEAR 1988 Report[U4], are listed in Table 21. This listing is relatively
extensive, allowing regionally relevant estimates to be
derived. Some variations within geographic regions are
fairly wide. Some of these differences might be explainedby local habits. Other differences may result frominconsistencies in the definitions of the food categories,
especially for leafy and other vegetables. Populationweighted average values from this listing [U4] are given in
Table 21. These may be taken to be reasonable
representative for verybroad geographic regions. Changing
dietary habits, however, require such food consumption
data to be periodically updated.
154. Although many regional differences in consumption
can be noted, the data seem to separate only very broadly
into western and eastern countries. The western dietcontains greater amounts of dairy products and meat.
These foods are replaced by grain products, vegetables, andfish in Asian countries. The average value of consumption
for the world would not apply to any individual and could
only be used in some generic dose assessments. The
consumption rates of children are less well known. In the
UNSCEAR 1993 Report [U3], milk consumption was
assumed to be 120 kg a1 for infants and 110 kg a1 forchildren. Other foods were assumed to be consumed at therate of two thirds (children) or one third (infants) of theANNEX A: DOSE ASSESSMENT METHODOLOGIES

42

adult values [C5]. This gives consistent and reasonablevalues to be used in dose assessments (see Table 13 ofAnnex B, “Exposures from natural radiation sources”).
155. Drinking water intake has been estimated forreference individuals. For both water and beverages, theestimates are 500 l a1 for adults, 350 l a1 for children and
150 l a1 for infants [I7]. Since the water balance is affectedby ambient temperatures, regional estimates of thesequantities should be established, if possible.
156. The consumption of foods from semi-natural andnatural ecosystems, such as mushrooms and game, varies
widely and is, in general, poorly known. Although thesefoods may comprise only a few percent (510 kg a1) of anindividual’s total annual dietary intake, such intake could
be important for some radionuclides in certain times andplaces such as the arctic food chain (lichen-caribou/
reindeer-human) for both natural and fallout radionuclides
and for consumers of game and forest mushrooms andberries for 137Cs following the Chernobyl accident. Usually
only a very small portion of a country’s population will besignificantly affected, so collective dose estimates are little
modified. For further analyses of these situations, betterdata on the consumption of these foods are needed.

3. Methods for estimating exposures(a) Atmospheric nuclear testing

deposition density rate in the year i, Fi, in the previous year,Fi1, and in all previous years, reduced by exponential decay.
The exponential decay with decay constant λ reflects both
radioactive decay and environmental loss of the radionuclide.
The coefficients bi and the parameter λ are determined byregression analysis of measured deposition and diet data.
159. The transfer coefficient from deposition to diet isgiven by
P 23 

"m
0"

C (t) dt /m0F (t) dt

j C / j F (28)
"

or

i1

"

i

i1



i

From the above model, the transfer coefficient can be
expressed as
P 23  b 1  b 2 b 3 e &m / (1  e &m)

(29)

where bi are the transfer components per unit annual
deposition: b1 is the transfer in the first year, primarily
from direct deposition; b2 is the transfer in the second year
from lagged use of stored foods and uptake from thesurface deposit; and b3 is the transfer via root uptake fromthe accumulated deposit. The units of P23 and bi are
Bq a kg1 per Bq m2. In the exponential term, the unit for
λ is a1 and m is a constant equal to one year. The values ofthe parameters used are given in Table 22.

157. To make reliable assessments of doses through theingestion pathway of radionuclides released in atmosphericnuclear testing, extensive empirical data were compiled on theconcentrations of the relevant radionuclides in different typesof food and the diets of different population groups. The data
were analysed in previous reports of the Committee,
especially for 90Sr and 137Cs, which together with 14C, are themain contributors to the ingestion dose commitments fromthis practice [U6, U7]. To evaluate the transfer coefficients,
regression analyses were applied to models relating measured
radionuclide concentrations in diet to the annual deposition
density rates and the measured concentrations in relevantorgans. Information on diet and deposition levels of otherradionuclides are incomplete, so the P23 coefficients estimatedfor such radionuclides are less reliable than those available for90
Sr and 137Cs.

160. Results of regression fitting of this fallout model tomonitoring data were presented in previous UNSCEAR
Reports [U6, U7, U8]. Relatively minor adjustments inparameter values were needed in the fits to extended
monitoring data, indicating, in particular, that the
projections of long-term transfers are confirmed.

158. The empirical model used to relate the deposition
density of a radionuclide, specifically 90Sr or 137Cs, to theintegrated concentration in components of the diet or in
total diet is the following

162. If P23 is multiplied by the individual annual
consumption of food (kg a1), the transfer coefficient P24,
which relates to the intake of the radionuclide, is obtained.
The transfer coefficients P45 then relate the intake amount to
the dose (Sv Bq1). This is a committed dose that accounts forlonger-term retention of the radionuclide in the body.

C i  b 1 F i  b 2 F i1  b 3

je F
"

n1

&n 

in

(27)

where Ci is the concentration of the radionuclide in a foodcomponent or in the total diet in the year i due to the161. Adequate representations of transfers to the total dietor to separate components of the diet are obtained for
relatively uniform deposition during the year, as occurred
for fallout from atmospheric weapons testing. For
deposition occurring within a much shorter time period,
such as following the Chernobyl accident, the transfer isdependent on the particular agricultural conditions at thetime of deposition and on short-term restrictions on certain
foods in the diet that may have been imposed.

163. The Committee’s earlier evaluations of transfer
coefficients related the integrated concentration of theradionuclide in the body to the dose [U6]. For 90Sr, theempirical relationship was as follows:ANNEX A: DOSE ASSESSMENT METHODOLOGIESC b,i  c C d,i  g

je
"

m0

&bm

C d,im(30)

where Cb,i and Cd,i are the concentrations of 90Sr in boneand diet in the year i and the parameters c and g may berelated to short- and longer-term components of 90Sr
retention in bone. The exponential term accounts forradioactive decay and removal from the body. Averagevalues derived for the parameters are listed in Table 22.
This formulation is useful for determining the annualcomponents of dose from a specific deposition occurrence.
164. The results of transfer coefficient evaluations for anumber of radionuclides are listed in Table 23. For 90Sr and137
Cs, the values are the same as those previously derived[U3]. It should be stressed that the transfer coefficients P23,
P234 and P2345 are all calculated for an even distribution of
the deposition throughout the year, as was the case for
global fallout from atmospheric nuclear testing. If thedeposition occurs during the winter season, the transfer
coefficients are lower, and for a summer deposition they
are higher than the value for the even distribution.
165. The transuranic radionuclides considered by theCommittee in dose evaluation from atmospheric testing
were 238Pu, 239Pu, 240Pu and 241Pu together with its decay
product 241Am. The empirical model described above,equation (27), has been used to relate the deposition
amount to the integrated levels in diet. The lag term,
however, was not included (b2 = 0).
166. As the number of measurements of the annual
ingestion intake, Iig, of plutonium radionuclides were very
few and covered only 11 years, the determination of λ isvery uncertain; large variations in the value of λ result in
only small variations in the value of Iig. Taking λ to be verysmall, Bennett [B2] found the average solutions for b1 andb3 to be 3.3 102 Bq per Bq m2 and 3.5 104 Bq per Bq m2,
respectively, for 239,240Pu. The estimation of P234 depends on
the real value of λ. It could be as low as 5 102 Bq per
Bq m2 if the availability of plutonium decreases with amean residence time of 50 years (λ = 0.02 a1) and as highas about 10 Bq per Bq m2 for 239Pu and 3 Bq per Bq m2
for 240Pu, if the availability of plutonium decreased only asa result of radioactive decay (λ = 3 105 a1 and 1 104 a1).
Aarkrog [A4] estimated the transfer of 239,240Pu to bread, an
important component of diet, to be 2 102 Bq per Bq m2.
Until additional information becomes available, thegeometric mean of the extremes for transfer to total dietcan be assumed for the transfer coefficient P234, namely,
0.7 Bq per Bq m2. This result corresponds to a mean
residence time of 239,240Pu in soil of about 100 years, thevalue that was also adopted in Annex C of the UNSCEAR
1982 Report [U6] for the mean residence time in soil oflong-lived natural radionuclides released from industrial
plants.

43167. For 238Pu, the above estimate of P234 using the 50year
residence time in soil is appropriate, considering thesimilar radioactive half-life of this isotope. Given the shorthalf-life of 241Pu (14.4 a), the value of P234 is dominated by
the rate effect and is taken to be equal to 4 102 Bq per
Bq m2. In the case of 241Am, the formulation iscomplicated by the need to take the decay of 241Pu into
account. Using the equivalent of equation (29) and taking
λ s to be very small and b1 to have the same value as thatobtained for 239,240Pu, Bennett [B11] estimated b3 to be equal
to 8 104 Bq per Bq m2. This value is very uncertain, as
only one measurement of the annual dietary intake of241
Am has been reported, but it points to the possibility thatamericium contained in the soil may be slightly moreavailable to plants than plutonium. The value of P234 can beroughly assessed to range from 6 102 Bq per Bq m2 for aresidence time of 241Am in soil of 50 years to 0.7 Bq per
Bq m2 if the availability of 241Am decreases only byradioactive decay. The geometric mean of this range is
0.2 Bq per Bq m2.
168. The estimated values of the transfer coefficients for thetransuranic radionuclides are listed in Table 23. Theseestimates are about 20 times higher than those usedpreviously by the Committee because of the higher values ofthe dose factors that have since been recommended by ICRP.
The total dose from plutonium will, however, not be
influenced by this change because the dominating pathwayfor
plutonium is inhalation, and here the dose factors are reduced
by a factor of 4 (for class Y = type S). The transfer
coefficients for 241Am are also listed in Table 23.

169. Curium is chemically very similar to americium, and it
may be assumed that the transfer coefficients for the various
curium isotopes can be calculated as for 241Am, taking thehalf-lives of the curium isotopes into consideration.
Curium-244, which has a half-life of 18.1 years, can thus beassumed to have a P234 coefficient equal to 0.04 Bq per Bq m2
and with the dose factor 1.2 107 Sv Bq1, P2345 is estimated tobe 5 nSv per Bq m2.
170. Less complete data are available for deriving transfer
coefficients for 131I, 140Ba and 55Fe. Radioiodine can betransferred rather quickly via the pasture–cow–milk chain tohumans. Hence, although 131I is a short-lived radionuclide(half-life: 8 days), it may contribute significantly to the dosein the first weeks after a release. P234 for 131I was calculated tobe 0.07 Bq per Bq m2 from a P23 coefficient for milk of0.63 mBq a l1 per Bq m2 and an average milk consumption
rate of 0.3 l d1. The dose factor for 131I ingestion by adults is
2.2 108 Sv Bq1, so P25 becomes 1.5 nSv per Bq m2 for 131I.
For the age group 01 year, daily milk consumption is 0.9 l
and the dose factor is 1.8 107 Sv Bq1, so in this case P25
becomes 37 nSv per Bq m2. A weighted average of P25 for all
age groups for ingestion is 4.3 nSv per Bq m2 for 131I.
Similar considerations were applied for 140Ba. The estimates
are included in Table 23.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

44

171. The transfer coefficient P234 for 55Fe was estimated in
the UNSCEAR 1993 Report [U3] to be 10 Bq per Bq m2.
For adults, P45 is 3.3 1010 Sv Bq1 for 55Fe, and P2345 for theingestion of 55Fe becomes 3 nSv per Bq m2. It should be
noted that P234 for 55Fe also includes a contribution fromconsumption of fish, which are known to concentrate 55Fe
from seawater [I10]. Hence P234 for 55Fe in the terrestrial
environment is overestimated. Values of 6 Bq per Bq m2
for P234 and 2 nSv per Bq m2 for P2345, assumed applicablefor terrestrial pathways, have been inserted in Table 23.
172. Although not previously considered in exposure
evaluations from ingestion, some limited data may be used
to derive rough estimates of transfer coefficients for 54Mn
and 144Ce. From measurements of 54Mn in grain in
19621966 in localized areas in the northern hemisphere[A5], the transfer coefficient from deposition density toconcentration in grain was estimated to be 0.025 Bq a
kg1 per Bq m2. Assuming an annual consumption ofgrain products of 80 kg and that all 54Mn in the diet comesfrom grain products, P234 becomes 2 Bq per Bq m2. For
adults, P45 is 7.1 1010 Sv Bq1 for 54Mn, and P2345 foringestion of 54Mn becomes 1.4 nSv per Bq m2.
173. Cerium is relatively unavailable to plants. Assumingthat a first-term component similar to that of plutonium
applies and that there are no other terms because of the shorthalf-life of 144Ce, the transfer coefficient P234 would have thevalue 0.1 Bq per Bq m2. The further transfer coefficients havebeen added in Table 23.
 









5T


%U


+

 










(G


5T





$C



 

















 

Figure IV. Ingestion exposure following unit deposition(1 Bq m -2) of radionuclides.

174. The above analyses have been made to derive dose
commitments per unit deposition density, but annual valuesof contributions to dose from a single deposition event may
be useful and can also be provided. These values are given
in Table 24. Because of the short half-lives of 131I and 140Ba,
the dose is delivered within a few weeks of deposition andfor 89Sr, within a few months. All of the dose is delivered
within one year of the deposition. There is no model for thetransfer of 55Fe to diet, but an approximation may be madeon the pattern of transfer of longer-lived radionuclides todiet, namely significant transfer within the first two years
following deposition and residual transfer over theremaining mean life of the radionuclide (see footnote toTable 24). The empirical models for 90Sr and 137Cs provide
the time course of transfer to dose for annual periods
following deposition of these radionuclides. The annual
contributions to dose from ingestion for a period of 10
years following deposition are illustrated in Figure IV. The
contributions beyond 10 years, which are significant only
for 90Sr and 137Cs, are given in Table 24. The total of allannual contributions is equal to the dose commitment.
(b) Nuclear installations
175. Radionuclides released to the atmosphere from
nuclear installations may contribute to exposures fromingestion in the local and regional areas surrounding thesite. The concentrations of the radionuclides in theenvironment and the doses are too low to be measured, butthey can be estimated with calculational methods.
176. The dispersion estimation method described in Section
I.B.3 and applied to the external exposure pathway is also
applicable to ingestion exposure, substituting in equation (24)the P25 transfer coefficients for ingestion intake. A more
specific designation of the ingestion transfer coefficients isP2345. The values in Table 23 are applicable also to the case ofroutine continuous or long-term averaged releases of radionuclides from nuclear installations. Several additional
radionuclides not normally included in analysis of weapons
fallout but present in releases from nuclear installations suchas 59Fe, 58Co, 60Co and 134Cs have been added to Table 23.
177. The estimates of local and regional collective dosefrom ingestion per unit release of radionuclides from
nuclear installations are presented in Table 25. The results
should be adjusted if it is known that some portion of thediet is derived from non-local foods. Also the representative population density may not apply to specific sites.
The representative values of population densities forvarious steps of the fuel cycle are given in Table 26.
178. Specific values given in Table 25 of the collective
dose per unit release are needed in exposure evaluations forreleases from separate fuel cycle installations. For thegeneral category of particulates released from reactors, a
representative composition maybe assumed (see paragraph96). The weighted average collective doses from ingestion
per unit release of particulates are included in Table 25. Of
course, the specific radionuclide weightings should be
adjusted, if the exact composition of the release is known.
179. An alternative method, the specific-activityapproach,
is used to estimate doses from tritium and 14C. In thisapproach the specific activity of 14C, for example, iningested food and water (activity per gram carbon) is
assumed to be the same as the activity per gram carbon inANNEX A: DOSE ASSESSMENT METHODOLOGIES

air at the point of interest. This is a good approximation forsituations where rapid exchange occurs, such as between
atmosphere and terrestrial biota, and the specific-activity
model provides a good estimate of 14C doses for chronic
releases from nuclear facilities. However, it is necessary toknow the carbon content of plants and animals to apply
this approach rigorously. The specific activity in air isreflected in humans after about one year.
180. The specific-activity model for tritium is expressed interms of the tritium to hydrogen atom ratio. For aqueouscompartments (air moisture, plant water, soil water and soon), the constancy of this ratio is equivalent to assumingthat the HTO concentration in Bq l1 is constant. However,a strict specific-activity approach overestimates doses fortritium, since it assumes a level of equilibrium between
tritium in the environment and in the atmosphere that israrely achieved. Concentrations in precipitation, andtherefore in soil, are lower than those in air, because theairborne plume is not always present when precipitation
occurs. Concentrations in plants will be lower than thosein air by an amount that depends on the transpiration rate.
Concentrations in drinking water tend to be much less thanair concentrations because of the large dilution that occurs
in most drinking water sources. Concentrations in animals
and humans reflect the concentrations in the food products
and drinking water they ingest.
181. The general formula for the specific-activityapproachto evaluate the collective dose is as follows:

Si 
50

N1

5 10 7 s m 3 I g,nd g,i3.15 10 7 s a 1 C a,n2000

x 1.4 2  x dx  N 2

1

(31)

x 1.4 2  x dx

50

where Ig,n is the ingestion intake rate of the stable form of
element n (kg a1); Ca,n is the concentration of the stableform of element n in air (kg m3); dg,i is the effective dose
per unit intake by ingestion of radionuclide i (Sv Bq1); andN1 and N2 are the population densities in the local andregional areas. For tritium, the exponent in the powerfunction of distance should take the value 1.2 rather than
1.4 because of less local retention of deposited tritium.
(i) Tritium
182. Application of equation (31) for tritium requires
estimates of the intake rates of both water-bound andorganically bound hydrogen in foods and drinking water. To
account for the fact that tritium concentrations in the various
foodstuffs ingested are lower than the concentrations inmoisture in air, Ig,n is determined as follows:I g,n 

jfU
i

i

i

(32)

45

where Ui is the intake rate of hydrogen from ingestion offood type i and fi is the ratio of tritium concentration infood type i to the concentration in moisture in air.
183. Representative intake rates of plant foods, animal
foods, and drinking water may be assumed to be 370, 170,
and 500 kg a1, respectively. With typical water content ofplant foods of 85% and of animal foods of 78% [D6] andthe hydrogen content of water being 11.1%, the intake
rates, Ui, of water-bound hydrogen are 35 kg a1 in plantfoods (370 × 0.85 × 0.111), 15 kg a1 in animal foods(170 × 0.78 × 0.111), and 56 kg a1 in drinking water (500× 0.111).
184. In the organic matter of foods, the hydrogen content
is, on average, 5.8% in plant foods and 8.4% in animal
foods [D6]. The intake rates, Ui, of organically boundhydrogen are thus 3.2 kg a1 in plant foods (370 × 0.15 ×
0.058) and 3.1 kg a1 in animal foods (170 × 0.22 × 0.084).
185. The value of fi for plant foods is about 0.8 or less [D1,H3, H4, M4]. For drinking water the value of fi is variable,
depending on local conditions. At sites on large water bodies,
where tritium enters only from the atmosphere, fi tends to beless than 0.1 [L2, N6]. Larger values of fi, even approaching
1.0, might apply to small water bodies, but the low volume orflow rate of such sources would limit the suitability of the site
to supply drinking water. Larger values of fi could also apply
to sites downstream of liquid discharges of tritium [N6] or if
groundwater had been contaminated. Both of these cases,
however, do not pertain to atmospheric releases. For thepresent calculations, fi for drinking water is assumed to be 0.1.
The value for specific sites should be based on local
conditions. For animals, it may be assumed that 40% of water
intake is derived from drinking water [R1]. The value of fi foranimal foods is thus estimated to be 0.5 for combined intakesof drinking water and plants (0.4 × 0.1 + 0.6 × 0.8).
186. It will be assumed that the concentration of organicallybound tritium (Bq l1 water equivalent) is the same as theconcentration of water-bound tritium in both plants andanimals so that the same values of fi apply to the aqueous andorganic phases. The value of Ig,n is then estimated to be
40 kg a1 in water-bound form (0.8 × 35 + 0.5 × 15 + 0.1 × 56)
and 4 kg a1 in organically bound form (0.8 × 3.2 + 0.5 × 3.1).
187. The annual average content of water vapour in air isassumed to be 8.1 g m3 [U6], implying that Ca for hydrogenis 9 104 kg m3. Population densities surrounding the point of
release are given above. The dose per unit intake, dg, waspreviously taken to be 2.2 1011 Sv Bq1 for water-boundtritium [U4, U6], but the value now recommended is 1.8 1011
Sv Bq1 [I5]. The dose coefficient for organically boundtritium is 4.2 1011 Sv Bq1 [I5]. Applying these parameters inequation (31), recalling that for tritium the exponent in the
power function is 1.2, and summing the water- andorganically bound doses, the result is 2.1 man Sv PBq1(local plus regional exposure) (Table 25). The dilution factor
and the dose per unit intake are lower than in previous

46

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

assessments by the Committee, and the allowance has been
made for reduced environmental concentrations relative tomoisture in air. These reductions are partially offset by theuse of a smaller exponent for the decrease in air
concentration with downwind distance and the separate
consideration of organically bound tritium. The net result
is a dose lower by about a factor of about 4 than thepreviously derived value of 9 man Sv PBq1 [U4, U6].
Organically bound tritium contributes about 20% of thedose but would contribute more for diets high in grain or
rice, which have high organic fractions.
(ii) Carbon-14
188. The dose from local and regional exposure to 14C
released to the atmosphere represents only a small
proportion of the total dose commitment. The mainsignificance of 14C stems from its global dispersion andentry into the carbon cycle, leading to long-term exposure(see Section V.B). The local and regional collective dose
commitment was previously assessed by the Committee
using the specific-activity approach. The Committee
assumed in its 1982 Report [U6] that the release of 14C isin the form of CO2 and the concentration of carbon in theatmosphere, Ca, is 0.16 g m3. A more recent, revised value
is 0.18 g m3 [T1]. The intake rate of carbon is 300 g d1 by
men and 210 g d1 by women, averaging 93 kg a1 intake
by ingestion, Ig. The dose per unit intake of 14C byingestion is 5.8 1010 Sv Bq1 [I5]; the value formerly used
was 5.6 1010 Sv Bq1 [U6]. It is assumed that, unlike
tritium, all components of the diet attain the specificactivity level of air at the location of interest downwind
from the source. Substituting these parameters into
equation (31), the result is 270 man Sv PBq1 (local plus
regional exposure) (Table 25).
189. For both tritium and 14C, the approximations of thespecific-activity method are recognized. The assumption
for 14C that all intake attains the specific activity at the
point of calculation is not realistic. For tritium, theconcentrations in the environment, although allowed todiffer from the concentration in air, are probably
overestimates. For both radionuclides, the time distribution
in the delivery of the dose must be ignored. The approachthus probably leads to overestimates of the doses.
Nevertheless, the method has the advantage of being asimple approach that can be easily adjusted for alternative
parameters that might more accurately reflect actual local
conditions.
190. Significant doses from a short-term 14C release will
be received only in the year of the release. Carbon-14 doses
arise only from ingestion and once the food crop of the yearof release is consumed, there are no significant pathways
for further exposure. Small amounts of 14C deposited in thesoil during the release may be re-emitted and taken up by
plants, but concentrations would be very low and doses
imparted by eating the plants would be insignificant
compared with those received in the year of release.

C. RADIONUCLIDES RELEASED TO THE
AQUATIC ENVIRONMENT
191. Radioactive contamination of the aquatic environment
may result in ingestion doses by three pathways: drinking of
freshwater from both surface and ground sources,
consumption of biota living in the water, typically fish, andconsumption of terrestrial foods that have been contaminatedby the use of freshwater for irrigation, by the application of
sediments as soil conditioners, or by the application of aquaticplants as fertilizer. Water consumed byanimals mayalso forma pathway for the transfer of radionuclides to the human diet.
Shoreline deposits of contaminated sediments can contributeto external exposures.
1. Transfer processes
192. Radioactive material released to the aquatic
environment is transported and dispersed by advective andturbulent processes occurring in the water body.
Interactions between radionuclides and suspended matterand sediments may remove radionuclides from thesolution. It is convenient to consider separate categories ofwater bodies for modelling the behaviour of radioactive
material: lakes, rivers, groundwater, coastal seas, andoceans.
(a) Lakes
193. Contaminants in lakes may occur in solution in thewater phase or in the sediments. Most radionuclides occur
in both phases, and the distribution factor Kd describes
their partition between water and sediments. Lakes receive
water from rivers, soil run-off and rainfall and lose waterby outflows and evaporization. The mean residence time of
the water in a lake depends on this in- and outflux of water
to and from the lake. The mean residence time of theradionuclide in the water phase of a lake depends
furthermore on the Kd for the radionuclide and its
radioactive decay. The water chemistry of the lake (pH,mineral and organic matter content, and redox) influence
Kd. These factors also influence the uptake of radionuclides
in biota. Lakes that are low in nutrients usually show
higher concentration factors from water to biota thannutrient-rich lakes.
(b) Rivers
194. Rivers may be considered as lakes with a high in- andoutflux of water. Thus, the mean residence time ofradionuclides in water in a river is usually significantly
shorter than that in a lake for a similar volume of water, so
lower concentrations are usually found in rivers than inlakes for the same input of radionuclides to the twosystems. The amounts of water carried by a river may varyconsiderably throughout the year. In the spring, when thesnow melts, the river may cover an area several times thatcovered in the dry season of the year. The flood land along
a river may retain radionuclides carried by the river water,

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

and this contamination may be released to the river again
in subsequent years. Accordingly, it is more complicated tomodel the behaviour of radionuclides in river systems thanin lakes. Sediments in the river bed may, during flooding
conditions, also be transported to new locations in the river
system and eventually be carried to the sea. Sediments may
also be disturbed by dredging and other activities.
(c) Groundwater
195. Lakes and rivers contain 0.3% and 0.003%,
respectively, of the total freshwater inventory of the world[U14]. Ice sheets and glaciers contain 75% andgroundwater the remaining part, i.e. about one fourth of all
freshwater is present as groundwater. Groundwater is, ingeneral, well protected against atmospheric radioactive
contamination, because adsorption, chemical precipitationand ion exchange prevent or delay the migration of many
radionuclides, such as 90Sr, 137Cs, and 239,240Pu. But someradionuclides, especially those of a noncationic form, e.g.
tritium, 99Tc, and 129I are not completely retained by thesoil. Tritium in the form of HTO is particularly mobile andis readily measurable in young groundwaters (less than 30
years). Groundwater may be contaminated in connection
with underground waste disposal. This has been seen, forinstance, at the Hanford site in the United States, whereliquid waste has been discharged to the ground,contaminating the groundwater, first of all with tritium.
Underground nuclear explosions at, for example, theNevada test site, contaminated groundwater with tritium[M9]. The contamination of groundwater by long-lived
radionuclides may be of interest in connection with thepermanent disposal of high-level waste in undergrounddepositories.
(d) Marine waters
196. The total volume of the water in the ocean is
1.37 1021 l [K1], which is four orders of magnitude morewater than found in rivers and lakes together. However,most of the water in the ocean belongs to the deep ocean
which is not used by man for food production. Fish andother marine foods are mainly produced in the coastal seas,
which have a mean depth of about 50 m and a volume of
1.37 1018 l, or 0.1% of the total water volume of the ocean.

47

desorption of radionuclides from sediments. The decreased
fixation in estuaries is partly counterbalanced by a lower
uptake by biota.
2. Methods for estimating exposures
199. Dose assessments for radionuclides released to theaquatic environment require, in general, information on
the activity of each radionuclide released, the volume of thereceiving water into which the radioactive material isdiluted, the concentration levels reached in fish andshellfish, the factors regarding removal to sediments andexchange rates of water bodies, and the number ofindividuals who use the water for drinking purposes andwho consume fish.
200. The local and regional collective dose commitmentsfrom radionuclides in liquid effluents can be estimatedusing the expression
S ic Ai

V (  i )jN f Idkk k,i k(33)
iwhere Ai (Bq) is the activity of radionuclide i released towater; V (liters) is the volume of the receiving water; τ(a1) is the reciprocal of the mean residence time of aradionuclide in the receiving water assuming no decay(removal to sediments is incorporated implicitly in thisvalue); λi (a1) is the radioactive decay constant forradionuclide i; Nk is the number of individuals for pathwayk; fk,i (Bq kg1 per Bq l1) is the concentration factor for anitem in pathway k for radionuclide i; Ik (kg a1) is theindividual consumption rate of pathway item k; di(Sv Bq1) is the effective dose per unit activity ingested.
201. The quantity A/V(τ + λ) (Bq a l1) is the integral
concentration in water for release of an activity A (Bq) or,
alternatively, the equilibrium concentration in water, Cw,i
(Bq l1), for a constant continuing release rate (Bq a1). The
equilibrium concentration in fish or shellfish is Cw,ifk,i(Bq kg1), where fk,i is the appropriate freshwater or saltwater concentration factor.

197. Some coastal seas are much like closed systems, andthe residence time of the water in such systems is relatively
long. Other coastal waters have a more direct connection
to the open ocean, and the mean residence time there isshorter. In the present context, the North Sea has beentaken as a typical coastal sea, and the mean residence timeof the water of the North Sea has been taken as
representative of all coastal seas.

202. For radionuclide releases to small volumes of water,
the concentrations in water or fish may be high, but the
population that can be served with drinking water or byfish production will be limited. For releases to larger water
volumes, the concentrations will be less, but the populations involved will be correspondingly larger. In fact, theNk/V relationship could be taken, in a crude
approximation, as relatively constant, the inverse of whichindicates the water use with regard to the specific pathway,
k, of each individual in the population.

198. Unlike freshwater systems, where the composition of
the water shows great variation, marine waters generally
have the same mineral composition. The increase insalinity in the transition from river to sea causes a203. For the drinking water pathway, a value for thequotient V/Nk of 2.2 107 l man1 is assumed for estimating
the collective dose commitments from generalized liquid
releases. This value is assumed to be a global average and48

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

is obtained from an estimated global total of 1.3 1017 l of
freshwater in lakes (1 1017 l) and rivers (annual flow
0.3 1017 l) [U7], serving a world population of 6 109.
204. Average fish plus seafood consumption per individual
is about 8 kg a1, ranging from 4 to 6 kg a1 in the Near
East and Africa to 1014 kg a1 in the Far East and Europe[I7]. It may be assumed that the annual consumption is
6 kg a1 ocean fish, 1 kg a1 freshwater fish and 1 kg a1
shellfish. Total freshwater fish consumption by the world
population is thus 6 109 kg a1, which, when a correction ismade for an edible weight of 50%, agrees with theestimated annual global harvest of 1010 kg landed weight[F2]. Dividing by the global freshwater volume given in
the above paragraph, the result is 4.6 108 man kg a1 l1,which will be assumed to be the factor NkIk/V needed for
estimating collective doses from freshwater fishconsumption.
205. The annual global ocean fish and shellfish harvest is1011 kg landed weight [F2], which is consistent with theocean fish and shellfish consumption by the world
population, 42 109 kg a1. The catch mostly takes place
within the continental shelf over an area of 27.5 106 km 2
and with a mean depth of approximately 50 m [K1]. The
volume of these waters is thus 1.4 1018 l. The factor NkIk/Vrequired for the salt-water fish and shellfish pathway is,therefore, 3 108 man kg a1 l1. This is about 35 times
higher than the factor used in the UNSCEAR 1977 Report[U7]. The mean residence time of the water over thecontinental shelf is assumed to be the same as thatobserved for the North Sea, i.e. approximately 3 years for90
Sr and 137Cs [N2] and 3.5 years for 239.240Pu (first pass).
Experience from Chernobyl has shown the turnover timeof 137Cs in freshwater systems to be 0.3 a1, i.e. similar to
the turnover observed in coastal waters. This turnover rateis less by a factor of 3 than the value of 1 a1 used in the
UNSCEAR 1977 Report [U7]. For 90Sr, the turnover rate in
freshwater systems is 0.2 a1, somewhat less than for 137Cs,
owing to a lower sedimentation rate.
206. The specific-activity concepts for tritium and 14C
discussed above apply in aqueous systems as well as interrestrial systems. HTO released to a water body istransported in the same way as other radionuclides but withthe additional process of evaporation, which can have alarge influence on HTO concentrations in some systems[H5]; neglecting this evaporation will result inoverestimates of the tritium concentration. For an atmo-

spheric release, concentrations in water bodies are usuallymuch less than in air because of the large amounts of water
available for dilution. Uptake of HTO by aquatic organisms
is very quick: concentrations in tissue become equal towater concentrations within minutes or hours. Aquaticplants form organically bound tritium through photosynthesis. Fish and invertebrates also produce small
amounts of organically bound tritium from the HTO in
their bodies and can directly incorporate organically boundtritium taken up through ingestion.
207. The calculations made here of local and regional
collective doses from tritium and 14C in liquid effluents are
based on equation (33) rather than on the specific-activity
model. Because tritium concentrations in water and aquaticorganisms are essentially the same, fk,i = 1 for tritium. On
the other hand, fk,i for 14C is very high, since the carbon
content of the organisms is much greater than the carbon
content of water.
208. The parameters used and the estimates of collective
dose per unit release of radionuclides to the aquatic
environment are given in Tables 27 and 28. The estimates
are the local and regional components of collective dose.
Many radionuclides have been included that might have to
be considered in specific circumstances. For releases of allradionuclides other than tritium in liquid effluents fromreactors, it is useful to specifya representative composition,
as was done for particulates in airborne effluents. The
release composition can vary widely depending on thereactor type, the fuel integrity and the waste management
practices. A representative composition is given in Table
29, which is derived from previously reported data [U3,U4]. Although these referred mainly to PWRs and BWRs,
the composition can be taken to be reasonably applicable toall reactor types. For analysis of worldwide releases fromreactors, the Committee has used an average of the results
for releases to freshwater and to salt water. In this case, theestimated collective dose per unit release of therepresentative composition of radionuclides in liquid
effluents is 330 man Sv PBq1. More appropriate selections
and weightings of values can be made in applications toactual circumstances of releases from specific sources.
209. For manyradionuclides, sediment removal considerations and radioactive half-lives limit the contributions to
global collective doses. Only a few radionuclides achieve
widespread, global dispersion, and these are considered in
the following Chapter.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

49

V. GLOBALLY DISPERSED RADIONUCLIDES
A. TRITIUM
210. Estimates of doses from globallydispersed tritium are
required for three sources: natural occurrence, atmosphericnuclear testing, and nuclear power production. The mostdirect estimates of dose are obtained from measurements ofthe environmental concentrations of tritium, which havebeen made at a number of locations worldwide and fromwhich individual doses from natural tritium and tritium
produced in atmospheric testing may be inferred.
Collective doses can be determined from an assumed
variation of dose with latitude and the known population
distribution. Doses from globally dispersed tritium arisingfrom nuclear power production cannot be derived in thisway, since the concentrations are undetectable beyond a
few kilometres from the release point. Instead, the doses
are estimated from model calculations.
211. The Committee based its estimate of the annual
effective dose from natural tritium on measurements of theuniform levels of tritium in surface waters (and in thehuman body) prior to input from man-made sources. The
estimated effective dose to individuals is 10 nSv a1 [U7].
With reference to the total annual production of natural
tritium of 72 PBq a1 (see Annex B, “Exposures fromnatural radiation doses”) and the present world populationof 6 109, the collective dose per unit release is 6 109 × 10
nSv a1 ÷ 72 PBq a1 = 0.8 man Sv PBq1. Considering the
population of each hemisphere (89% north, 11% south),the collective doses per hemispheric input are 1.5 man Sv
PBq1 for the northern hemisphere and 0.2 man Sv PBq1
for the southern hemisphere.
212. The doses from tritium produced in atmospherictesting were estimated initially from measurements of theconcentrations in surface waters [B3]. The estimated dose
commitments were 20 µSv in the northern hemisphere and2 µSv in the southern hemisphere [U7]. Based on estimatedinputs of tritium into the atmosphere from the practice of
1.9 1020 Bq to the northern hemisphere and 0.5 1020 Bq to
the southern hemisphere [U6] and applying the natural
tritium dose/production rate ratio, the estimates of dose
commitment were adjusted to 51 µSv and 14 µSv in thenorthern and southern hemispheres, respectively [U6].
These last values were derived from and correspond to thedose coefficients given at the end of the previous paragraph.
213. The models used to estimate the global doses fromtritium similate the world hydrological cycle. Calculations
are thereby made of the specific activity of tritium in thevarious global water pools. Most tritium is released to theatmosphere as HTO, and tritium gas (HT, T2) istransformed in the soil to HTO. Tritium, therefore, follows
the local and global water cycles. The hydrological models
are invariably formulated in terms of compartments, inwhich the tritium is assumed to be instantaneously anduniformly mixed. Transfers between compartments arequantified using rate constants that are based on the known
rates of water movement due to processes such as
precipitation, evapotranspiration and run-off. Tritium
concentrations in foodstuffs are assumed to equal
concentrations in air moisture, soil water or surface water,
depending on the model. The concentration of tritium inhumans is calculated from an average of the concentrationsin the sources of water ingested, weighted by the relative
amount that each source contributes to intake. Several
models of this kind exist, differing primarily in the number
and size of compartments considered. The compartment
approach is sufficient to calculate mean tritium
concentrations over long times and large spaces. As well as
providing estimates of doses from nuclear power
production, the models can be used to confirm the doses
from natural production and atmospheric testing deduced
from observations.
214. The simplest model for estimating global tritium
doses consists of single compartments representing thecirculating waters of the hemispheres (to an ocean depth of75 m). The model of Kelly et al. [K3], as implemented by
NRPB and the Commissariat à l’Energie Atomique (CEA)[N2], used this basic approach and allowed for slow
exchanges between the hemispheres and the deep oceans.
For a release to the atmosphere or to surface waters, thecollective dose per unit release was determined to be
0.028 man Sv PBq1 relevant to a world population of 4109. The Committee used this result in the UNSCEAR
1982 Report [U6] and adjusted it in the UNSCEAR 1988
Report [U5] to 0.032 man Sv PBq1 for a population of 4.6
109. These results are probably underestimates of doses,
because the tritium is mixed in large compartments that
include the world’s oceans and is diluted more than it
would be in the terrestrial environment normally accessible
by humans.
215. Improved estimates of the global dose from tritium
are obtained using more realistic models developed by the
NCRP [N3], Bergman et al. [B4] and Killough and Kocher[K2]. The seven compartments in the NCRP model
represent atmospheric water, surface soil water, surface
streams and freshwater lakes, groundwater, saline lakes
and inland seas, the ocean surface and the deep ocean(Figure V). Water volumes and mean residence times ofwater in each compartment were estimated, together withfractional transfer rates for movement among thecompartments. The volumes and transfer rates for thehemispheres and the world are listed in Table 30. The
intake of tritium by man was calculated from the predictedenvironmental concentrations and the amount of watertaken in through drinking and food ingestion. Eighty
percent of drinking water was assumed to come fromsurface streams and freshwater lakes and 20% from deep

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

50

groundwater. Tritium concentration in plant water was
assumed equal to 0.7 Ca + 0.3 Cs, where Ca and Cs are theconcentrations in air moisture and soil water, respectively.
Although the NCRP model is not divided into latitude

bands, it can be used to estimate doses from releases todifferent parts of the atmosphere by adjusting the size ofthe compartments to hemispheric or latitudinal water
volumes.

        
       
       

Figure V. Seven-compartment model of the hydrological cycle for global circulation of tritium [N3].

216. The model developed by Bergman et al. [B4]improved on the NCRP model by dividing all compartments into two latitude zones in each of the northernand southern hemispheres. It included a separate reservoir
for organically bound tritium in terrestrial biota and wasable to account for HT releases. The Killough and Kocher
model [K2] separates the atmosphere into stratosphere andtroposphere and further subdivided all atmosphere andocean compartments, allowing the model to account for
latitudinal inhomogeneities. Killough and Kocher noted
that, without the stratospheric compartments, HTO
entering the northern atmosphere is removed too rapidly topermit significant interhemispheric transport, andestimates of doses from atmospheric nuclear testing areunreliable. The use of a diffusive ocean module improved
the ability of the model to estimate concentrations in thesurface waters of the ocean.
217. The estimates of global collective doses fromatmospheric tritium releases obtained with several models
are shown in Table 31, together with the estimates basedon natural tritium production. The estimates of the model
calculations are those available in published reports, sincewith the exception of the NCRP model, the codes are too
poorlydocumented to be run independently. The results are
not easily compared since different source distributions areused. In general, however, there is a relatively good levelof agreement. The dose estimates for releases to the global
atmosphere are within a factor of 2, regardless of whether
latitudinal zonation is considered or not. This level ofagreement is maintained for releases to the northernhemisphere, but differences by a factor of 10 arise forreleases to the southern hemisphere. The estimate of
Bergman et al. [B4] for release to the global stratosphere(0.76 man Sv PBq1) agrees well with the global dose fromnatural production (0.8 man Sv PBq1). The NCRP results
tend to be lower than those of the other models. Killough
and Kocher [K4] found that the NCRP model underestimates observed freshwater concentrations of fallout
tritium by about an order of magnitude and overestimates

ocean concentrations by a factor of 3. Use of the NCRP
model therefore likely leads to underestimates of the global
collective doses from releases of tritium.
218. The global collective dose from near-surface
atmospheric releases from nuclear installations is best
obtained from model estimates of releases to the 30°50°
band of the troposphere in the northern hemisphere. The
Killough and Kocher estimate of 2.3 man Sv PBq1 is themost reliable in this regard. The NCRP model result is 0.7
man Sv PBq1 for this case, but as noted above, this isprobably an underestimate. The northern hemisphericestimate from natural production, 1.5 man Sv PBq1, mayalso reflect doses due to releases from nuclear installations,although it, too, may be an underestimate because therelease is not confined to the latitude band in which thegreatest population density is found.
219. Estimates of the global collective dose arising fromreleases to the ocean from nuclear installations areavailable from both the NCRP and Bergman et al. models.
Both obtain estimates of doses that are about one tenthlower than those resulting from atmospheric releases.
Taking the atmospheric result to be 2.3 man Sv PBq1, thedose from releases to the ocean becomes 0.2 man Sv PBq1.
220. Estimating the global distribution of tritium released
from nuclear installations is a difficult task, and calculated
doses contain an element of uncertainty. Based on acomparison of model estimates with observations [K2] andon the level of agreement among the estimates of the morereliable models, the true value of the global collective doseis believed to lie within a factor of 3 of the values given
above. Much of the uncertainty is due to the large size ofthe compartments used in the models. The average
concentrations assumed throughout these compartments
are incompatible with the rapid changes in concentrationthat occur in the environment surrounding local sources
and the non-uniform population density that actuallyexists.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

221. The seven-compartment NCRP model may be used todemonstrate the spatial and temporal variations in theestimated tritium doses (Tables 32 and 33).These results
should be considered illustrative only, since the NCRP model
does not include latitudinal zonation and it tends tounderestimate doses. However, it is well documented,
transparent and accessible, and its estimates are probably
realistic in terms of trends if not of magnitudes. Such resultsfrom use of other models are unavailable to the Committee.
222. Results of the seven-compartment model [N3] forreleases to different parts of the atmospheric compartment
are presented in Table 32. The slight difference between
the northern and southern hemisphere reflects the fact that
more of the global land surface (67%) is in the northernhemisphere and more of the global ocean surface (57%) isin the southern hemisphere.
223. The time course of the delivery of dose from tritium
released to the atmosphere is indicated in Table 33. In thisexample, the seven-compartment model [N3] is applied to the30°50° latitude band of the northern hemisphere. The
distribution of tritium within the seven compartments isindicated, with the decreasing total reflecting radioactive
decay. The concentrations of tritium within the compartments
may be determined by dividing by the water volumes:
1.7 1012 m3 in the atmosphere, 1.4 1013 m3 soil water,
5.6 1013 m3 in freshwater, 9.9 1013 m3 in saline water,
1.8 1015 m3 in groundwater, 2.7 1015 m3 in the ocean surface,and 1.3 1017 m3 in the deep ocean. The concentration inhumans is determined from the concentrations in theenvironment, weighted for fractional daily intake: 0.99 l fromthe atmosphere, 0.77 l from soil water (foods), 1.22 l fromdrinking water (80% from fresh water and 20% fromgroundwater) and 0.02 l from the ocean surface (seafood) fora total daily water intake of 3 l. The effective dose is largely
received within the first few years of release, since much of
the tritium is by then transferred to the oceans, from whichless than 1% of the water intake by humans is derived.
224. From the above discussion it would appear that someconsolidation of the results of tritium modelling would be
useful in order to be somewhat more certain about the best
estimates of global doses. On the whole, however, dose
estimates can be selected that should be adequate for thegeneral purposes. In summary, the estimates of the global
collective doses per unit release of tritium from various
sources are 0.8 man Sv PBq1 for natural production, 1.5 and
0.2 man Sv PBq1 for northern and southern hemispherereleases from atmospheric testing, and 2 and 0.2 man SvPBq1
for airborne and liquid discharges from nuclear installations.

B. CARBON-14
225. After its release, carbon is distributed among thevarious reservoirs of the global carbon cycle: the atmosphere, the terrestrial biosphere, the hydrosphere, and thelithosphere. The fluxes of radiocarbon and stable carbon

51

between the different reservoirs are governed by the same
exchange processes. Isotopic fractionation is negligibly
small, within the other uncertainties involved. The total
carbon content in the atmosphere is about 7.5 1017 g, of
which the overwhelming bulk is present as CO2. Exchangeof carbon with the terrestrial biosphere and thehydrosphere is estimated to be 2 1017 g a1, with more thanhalf going to the biosphere. The largest reservoir is thelithosphere (7.2 1022 g), but the exchange rates between
this and other compartments are extremely low.
226. Because of the long half-life of 14C, its consequencesmust be evaluated through the collective effective dose
commitment, which is complete about 50,000 years after the
release. About 70% of the collective effective dose
commitment will have been delivered by 10,000 years. Most
models assume that the global population grows until the
middle of the next century and then stabilizes at 1010 people.
227. As with tritium, the most direct estimates of global 14C
dose are obtained from environmental measurements. A
natural production rate of 1 PBq a1 leads to an individual
effective dose rate of 12 µSv a1. This implies a collective
effective dose commitment of 120,000 man Sv PBq1 if it isassumed that the equilibrium population of the world of 1010
is achieved within a short time compared with the mean
environmental lifetime of 14C.
228. Recent interest in climate change has led to thedevelopment of many models to studythe global circulation of
stable carbon. For the most part, these models cannot be used
to calculate global 14C doses without major modifications. The
models discussed below are those developed specifically toassess the doses from man-made sources of radiocarbon. As
was the case for tritium, they are all compartment models ofvarying complexity. The assumption of instantaneous mixing
in compartments is invalid in the short term for 14C but issufficiently accurate for long-term dose assessment. The
models predict activities per gram carbon in eachenvironmental compartment over time. Once mixing isachieved, the specific-activity model may be used to estimate
collective dose commitments from 14C. It is assumed that the
specific activity of 14C in the carbon ingested by humans is thesame as that in the most relevant compartments for foodintake (ground vegetation for terrestrial foods and relevant
surface ocean compartments for marine foods).
229. The long time required to deliver the dose means thatdetails of the source location and distribution are not asimportant for 14C as they were for tritium. For all doses
derived from model calculations, the release was assumed to
be to a single compartment representing the global
atmosphere, and the results apply equally to 14C releases fromnatural production, atmospheric testing and nuclear power
production.
230. The models for global carbon dose consider
radiocarbon only in the form of 14CO2, as this is the only
form in which 14C can enter the food chain. Thus, 14CO2 isANNEX A: DOSE ASSESSMENT METHODOLOGIES

52

the only direct contributor to ingestion dose, which makes
up 99% of the total 14C dose. Assuming that allradiocarbon is released as 14CO2 will overestimate doses ifhydrocarbons are also present in the emissions. However,
the hydrocarbons will be oxidized to 14CO2 within a few
years [E6], and this can be taken into consideration.
231. The ability to make reasonable time-dependent
estimates of regional and global 14C fluxes and doses fromarbitrary release locations over thousands of years requires afairly sophisticated model. It should include the atmosphere,
biosphere with multiple compartments, soil, oceans withmultiple layers (a well-mixed upper layer, unstirred dense
thermocline, and deep water), and, possibly, ocean sediments.
Input fluxes should include both 14C and 12C, so that the
specific activity of the radiocarbon can be calculated. Recent
models incorporate the influx of 12C from the burning of fossil
fuels.
232. The Committee has used a variety of methods toestimate global 14C doses for releases from nuclear
installations. The estimates in the UNSCEAR 1977 Report[U7] were calculated using a model with compartments forterrestrial biosphere, atmosphere and short-term biospherecombined; surface ocean, thermocline layer in the ocean (adiffusive layer), and deep ocean. The parameters were
adjusted to fit measurements of excess 14C in theatmosphere and surface ocean from atmospheric testing.
The incomplete (to 104 years) whole-body collective dose

commitment was found to be 120,000 man Sv PBq1 for a
future world population of 1010 people. In the UNSCEAR
1982 and 1988 Reports [U4, U6], the NRPB/CEA [N2] model
was used to estimate an incomplete collective effective dose
commitment of 67,000 man Sv PBq1 as an average for both
atmospheric and aquatic releases for a population of 1010,
which was assumed constant during the integration period. A
model developed by Emanuel et al. [E2] was used in the
UNSCEAR 1993 Report [U3] that produced estimates of theincomplete collective dose commitment of 85,000 man Sv
PBq1 for a projected world population of 1010 people.
233. Global 14C modelling has been further advanced by
the work of Titley et al. [T1], and this model isrecommended for use in 14C dose assessments. It contains
23 compartments (Figure VI): atmosphere, ocean
sediments, Antarctic Ocean (four layers), Atlantic Ocean(four layers), Pacific Ocean, including the Indian Ocean(three layers), Arctic Ocean (two layers), woody tree parts,
non-woody tree parts, ground vegetation, decomposers,soil, and a compartment representing input from fossil fuelburning. The terrestrial portion of the model was adaptedfrom Emanuel et al. [E1] with minor modifications toallow the transfer of soil via rivers to the ocean surface
compartments. Exchanges between the atmosphere and theterrestrial biosphere are based on estimates of thephotosynthetic uptake of carbon by plants and its release to
the atmosphere by plants, animals, and soil by respiration[C4, E1].

PACIFIC OCEAN
0-500 m

500 2200 m

TREES

TREES

(woody (non-woodyparts)
parts)

2200 4200 m

ANTARCTIC OCEAN

500 - 2200 - 3200 0-500 m 2200 m 3200 m 4000 m

SOIL

ATLANTIC OCEANDECOMPOSERS
SEDIMENTS

ATMOSPHERE

500 - 2200 - 3200 0-500 m 2200 m 3200 m 4200 m

GROUND
VEGETATIONARCTIC OCEAN
0-200 m200-1200 m

FOSSIL FUEL

Figure VI. Compartment model for global circulation of carbon-14 [T1].

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

234. The ocean model in Titley et al. [T1] takes into
account temperature changes, surface areas and varying
amounts of ice cover in winter. Photosynthesis in thesurface ocean layers and subsequent transfer of carbon
down the water column was included and found to beimportant. In contrast, net sedimentation from water to theseabed was found to be a relatively insignificant process.
The parameter values for the oceanic compartments were
derived from several references [B6, M1, S3]. Exchanges
between the oceanic and atmospheric compartments werebased on estimates of the dissolution of CO2 at theocean/atmosphere interface using data from Mobbs et al.
[M1] and Siegenthaler [S3]. Carbon dioxide is very
soluble, and exchange with the atmosphere is rapid in open
aerated water. The model was tested and validated against
stable carbon distributions and 14C specific activities
arising from natural sources and atmospheric nuclear
testing [T1].
235. The model of Titley et al. [T1] provides an estimateof the complete collective effective dose commitment perunit release: 109,000 man Sv PBq1. This is similar to theestimate of Emanuel et al. (108,000114,000 man Sv
PBq1), to the estimate derived from natural 14C production(120,000 man Sv PBq1) and to previous UNSCEAR
estimates. Indeed, the 14C dose estimates of the manymodels in the literature are all very consistent. Killough
and Rohwer [K5] found that the predictions of six models
ranged over a factor of only 1.5. A similar range was foundby Titley et al. [T1] in their comparison of four othermodels. Finally, McCartney et al. [M5] found less than a
15% difference in the results of three models. Killough andRohwer [K5] attribute the consistency to the long half-lifeof 14C relative to its rate of environmental transport, whichmakes the estimated dose commitments insensitive to thedetailed structure of the models or to the values of theparameters used in them.
236. The collective dose coefficient of 109,000 man Sv
PBq1 was calculated with the assumption that the releaseis to the atmosphere, that the future world population
stabilizes at 1010 people, and that the global inventory ofstable carbon does not increase from its present value.
Based on the values provided by the various models, there
is a high probability that a range of 100,000140,000
man Sv PBq1 will encompass the actual collective dose
under these conditions. Assuming fossil fuels continue tobe burned at the present rate of 5 1015 g carbon per year
until supplies are exhausted, the best estimate of thecollective dose (from predictions of the Titley model) is
92,000 man Sv PBq1, with a range of 80,000130,000
man Sv PBq1. Doses following a release to soils or surface
oceans are about the same as those for an atmosphericrelease, but doses from release to deep oceans would be
about 20% lower.
237. The time course of collective dose for a release of 14C
to the atmosphere or to the ocean surface is shown in
Table 34. The equilibrium specific activities assuming

53

fixed, stable carbon inventories match those of natural 14C
production, which is of the order of 1 PBq a1. Estimates ofdose are given for a variable inventory of stable carbon
caused by the burning of fossil fuels. About 9% of thecomplete dose commitment from a single release isdelivered within 100 years, 23% within 1,000 years and
75% within 10,000 years.

C. IODINE-129
238. Because of its very long half-life (1.57 107 a), 129I may
become widely distributed in the global environment much
like stable iodine, 127I, over a long time. Whether released
into the atmosphere or into the aquatic environment, 129I
will eventually reach the oceans in a time period
presumably shorter than its half-life. Iodine is released
from the ocean into the atmosphere as organic iodine(mostly as methyl iodide) [L1] as a consequence of
microbial activity. The emitted organic iodine isdecomposed by sunlight into inorganic iodine compounds.
Both the organic and inorganic forms enter the terrestrial
environment by the processes of wet and dry deposition[W1]. The deposition velocity of inorganic iodine onto
vegetation is about two orders of magnitude higher thanthat of organic forms [N4]. The global iodine cycle and the
dynamic behaviour of iodine in the environment is being
further studied to improve the estimates of doses from 129I
releases.
239. Doses to humans from 129I are delivered principallyby its incorporation into the body by ingestion or
inhalation. Iodine accumulates primarilyin the thyroid, but
the low specific activity of 129I (6.55 MBq g1) limits theactivity of the radionuclide that can be present in the gland[T1]. Artificially produced 129I is released into theenvironment from nuclear installations, and small amounts
were also released in atmospheric nuclear testing.
240. The behaviour of iodine in the terrestrial environmentis influenced by many factors, e.g. soil type, microbialactivity, and chemical form. It is known that stable iodine
accumulates in soil; iodine concentrations in soil are
101,000 times higher than those of the parent rocks. The
levels of 129I in soils collected from the vicinity of nuclear
reprocessing plants are markedly higher than the levels inother places [B5, M7, R6]. Vertical distribution of 129I in
soil showed that most of the radionuclide is retained in thesurface layer (<10 cm). These observations indicate thatthe transfer of 129I from the terrestrial environment to theocean would occur only relatively slowly.
241. A model of the global iodine cycle was developed by
Kocher [K8]. The environmental compartments assumed
in the model are the atmosphere, hydrosphere, lithosphere,
and terrestrial biosphere. It is estimated that the mean
residence time of iodine in surface soil is of the order of
10,000 years and that the mixing of iodine throughout theocean would require 1,000 years or more. Therefore, theANNEX A: DOSE ASSESSMENT METHODOLOGIES

54

most important parameter for determining dose rates andcumulative doses following the release of 129I is the
10,000-year mean residence time of iodine in the surface
soil region. It thus appears that for a realistic long-term
population dose assessment, a progression from local toregional to global-scale models would be required [K8].
When the released 129I reaches equilibrium with stableiodine, the specific-activity method could be used in theassessment.
242. In the specific-activity approach, the activity
concentration of 129I per unit mass of 127I is assumed to be
the same in sea water and in the human thyroid. Assumingthat 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 toan age-weighted annual thyroid dose of 1.5 107 Gy. Since

the sea contains 3.8 1016 g stable iodine (water mass of 61023 g and iodine concentration in water of 0.064 µg g1),
a release of 1 PBq 129I results in a long-term specificactivity of 0.026 Bq g1. The collective thyroid dose
commitment arising from the discharges of 129I would be
about 9 108 man Gy PBq1, assuming a world population of1010 and no sink for iodine in the environment.
243. The compartment model for the global circulation ofiodine is shown in Figure VII. This represents a revision [T1]of the model described by Kocher [K8] and modified bySmith and White [S4]. The inventories of stable iodine in themodel compartments and the fluxes between them were
determined from environmental measurements and from therequirement for mass balance. Iodine-129 released into any
compartment is assumed to be transported with stable iodine,and so the specific activity of 129I can be determined for eachcompartment. Intake of 129I occurs by inhalation and by theingestion of water and terrestrial and marine foods.

OCEAN
ATMOSPHERE

LAND
ATMOSPHERETERRESTRIAL
BIOSPHERE

OCEAN
MIXED LAYER

SOIL
WATER

SOIL SOLIDDEEP OCEANSHALLOWGROUNDWATER

OCEAN
SEDIMENTS

DEEP
GROUNDWATER

SEDIMENTARY
ROCK

Figure VII. Compartment model for global circulation of iodine-129 [T1].

244. Important fluxes added to the global iodine model arefrom the sedimentary rock compartment to the two
subsurface groundwater compartments and to the solid soil
compartment, and from the solid soil compartment to theocean mixed layer. This model includes the transfer ofiodine from soil to the oceans and its subsequent movementback to soil from sedimentary rock. Titley et al. [T1]estimated the residence time of iodine in deep ocean waters
to be 350 years and the flux of iodine from the deep ocean
to the ocean mixed layer to be 2.3 1014 g a1. The amountof iodine transferred annually from the sedimentary rockcompartment back to the solid soil compartment isestimated to be 1.8 1011 g a1 [T1]. The mean residence
times of iodine in the major compartments used in themodel are 0.1 years in the ocean atmosphere, 0.09 years in
the land atmosphere, 5.9 years in the ocean mixed layer, 19

years in the terrestrial biosphere, 3.6 105 years in ocean
sediments, 970 years in shallow subsurface region, and38,000 years in deep subsurface region [T1].
245. The doses to individuals and collective doses
following a release of 129I can be calculated using theestimated time-dependent concentrations in the various
compartments and either a pathway or a specific-activity
analysis [K9]. The pathway analysis procedure involves
identification of a number of exposure pathways; transfer
coefficients are then used to estimate the movement of theradionuclide from the various compartments to humans.
This approach requires considerable judgement because of
the possible variations in the transfer coefficients and in
the assumed intake rates, but the results are then quite
realistic. The specific-activity approach is a means of

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

bypassing all the uncertainties and difficulties associatedwith the pathway analysis.
246. The pathway analysis method was adopted in the 129I
model under consideration. Five exposure pathways wereassumed as follows: inhalation by humans from the land
atmosphere, the daily intake rate of iodine being
0.29 µg d1; deposition from the land atmosphere onto foodcrops ingested directly by humans or by dairy and beef
cattle and subsequently ingested by humans (6.6 µg d1);
ingestion of land surface water directly by humans or by
cattle (5.3 µg d1); ingestion of marine fish and shellfish
from the ocean mixed layer (11 µg d1); root uptake fromthe surface soil region or from the soil water region into
crops consumed by humans or by cattle subsequently
ingested by humans. The intake of iodine through root
uptake considers the concentration of iodine in theterrestrial biosphere and the ingestion of vegetables,
cereals, all other foods, meat and milk. The daily intake ofiodine through root uptake of iodine using average worldconsumption rates is 200 µg d1, and the total daily uptakeof iodine is 220 µg d1 [T1]. The calculation of effective
dose utilized the following values: equivalent dose in the
thyroid per unit intake 1.3 µSv Bq1 (inhalation) and2.1 µSv Bq1 (ingestion) and tissue weighting factor 0.05.
247. A comparison of collective effective dose to the world
population arising from a release of 1 TBq of 129I during
one year to the five different compartments calculated

55

using this model [T1] is given in Table 35. At 108 years thecollective effective dose for release to the land atmosphere(727 man Sv) and to solid soil (828 man Sv) are higher
than the collective effective dose for release to the ocean
compartments, 530, 469, and 469 man Sv for release to theocean atmosphere, the ocean mixed layer and the deep
ocean, respectively. The trend in collective effective doses
from 50 years indicates higher amounts of iodine in theland atmosphere with negligible amounts in the deep
ocean, but by 108 years the amounts in the deep ocean will
have increased, while the amounts in the land atmosphere
will have decreased. Thus the transfer to the deep ocean ismuch faster than the reverse process.
248. The long residence times of iodine in the solid soil
compartment and the deep ocean compartment and the factthat a larger fraction of iodine in the ocean mixed layercompartments is transported downwards rather than to theatmosphere imply that it takes much longer for 129I to reachthe soil water compartment, from which most of the iodine
intake by humans is derived. Collective effective doses
estimated assuming that 129I is discharged into the land
atmosphere compartment are generally higher because of
the direct connection between this compartment and thesoil water compartment. The long residence time in thesedimentaryrock compartment implies that iodine entering
the sedimentary rock compartment is trapped there for atime of the same order as the half-life of 129I before being
cycled back to the soil.

CONCLUSIONS
249. In this Annex, the procedures used by theCommittee for calculating doses from radionuclides inthe environment are reviewed and updated. The
radionuclides considered are those present either because
they occur naturally or they have been released byanthropogenic practices. Although the calculational
procedures are well established from extensive
measurement and modelling experience, the increasing
knowledge of transfer processes and radionuclide
behaviour and better judgement of representative
conditions allow the relevant parameters to be adjusted
and the dose estimates to be improved.
250. For the Committee’s purposes of estimating
average doses under general conditions of release orpresence of radionuclides in the environment, relatively
simple calculational methods are sufficient. Moredetailed, time-dependent or otherwise complex methods
have not been considered. For releases to the atmosphereor to the aquatic environment, such as those that occur
from nuclear installations, average annual doses per unit
release are estimated for populations in the local andregional areas. For longer-lived radionuclides that
become widely dispersed, the average global doses arealso evaluated. The main pathways of external irradiation, inhalation, and ingestion are considered.
251. The Committee has selected representative parameters to reflect the various conditions of release,environmental transport and behaviour, and the personal
habits of intake and metabolism of the various radionuclides. These should provide reasonably accurate
estimates of dose in many applications. Alternative
selections of the parameters may lead to wide variationsin the dose estimates. Therefore, the methods presented
in this Annex should be used with caution. In particular,
it is recommended that site-specific data should be used
as appropriate and when available.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

56

Table 1
Radiation weighting factors[I1]

Type of radiation

Energy range

Radiation weighting factor wR

Photons, electrons, muons

All energies

1

Neutrons
Protons<10 keV, >20 MeV>2 MeV5
5

Neutrons

10-100 keV, >2-20 MeV

10Neutrons
Alpha particles, fission fragments, heavy nuclei>0.1-2 MeV
All energies

20
20

Table 2
Tissue weighting factors[I1, I11]
Weighting factor wT
Tissue or organ
GonadsBreastColon
Red bone marrow
Lungs
StomachUrinary bladder
Liver
Oesophagus
Thyroid
Bone surface
Skin
Remainder
a
b
c

1977

1990

0.25
0.15

0.20
0.05
0.12
0.12
0.12
0.12
0.05
0.05
0.05
0.05
0.01
0.01
0.05 b c

0.12
0.12

0.03
0.03
0.30 aThe value 0.06 is applied to the average dose among each of the five remaining organs or tissues receiving the highest dose, excluding the skin,lens of the eye, and the extremities.
The remainder is composed of the following tissues and organs: adrenals, brain, extrathoracic region of the respiratory tract, small intestine,
kidney, muscle, pancreas, spleen, thymus, and uterus.
The value 0.05 is applied to the average dose to the remainder tissue group. However, when the most exposed remainder tissue or organ receives
the highest committed equivalent dose of all organs, a weighting factor of 0.025 is applied to that organ and a weighting factor of 0.025 is applied
to the average dose in the rest of the remainder.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

57

Table 3
Values of the parameters used to evaluate vertical dispersion in the Gaussian plume model
Stability-dependent parameters
Stability class
a

b

d
-4

5.38 10
6.52 10-4
9.05 10-4
1.35 10-3
1.96 10-3
1.36 10-3

1.060
0.950
0.920
0.889
0.895
0.783

0.112
0.130
0.112
0.098
0.06090.0638

A: extremely unstable
B: moderately unstable
C: slightly unstable
D: neutral
E: slightly stable
F: moderately stable

c

0.815
0.750
0.718
0.688
0.684
0.672

Roughness-dependent parameters

Roughness length
(m)

p

q

s
-4

6.25 107.76 10-4
0
18.6
4.29 103
4.59 104

0.048
0.02690-0.098-0.0957-0.128

1.56
2.02
2.72
5.16
7.37
11.7

0.01: Lawns, water bodies
0.04: Plowed land
0.1: Open grassland
0.4: Rural areas, small villages
1.0: Forest, cities
4.0: Cities with tall buildings

r

0.45
0.37
0
-0.225-0.60-0.78

Table 4
Representative values of meteorological and release parameters
Parameter
Effective release height (H)Direction frequency (fi)Sector width (∆θ)
Roughness length (z0)

Units

Value

m
Dimensionless
Radians
m

30
0.083
0.524
0.4
Stability classFrequency of occurrence (f1)Wind speed (ui)Inversion height (hi)Dimensionless
m s-1
m

A

B

C

D

E

F

0.05
2
2 000

0.10
3
1 500

0.20
4
1 200

0.30
5
800

0.20
3
400

0.15
2
200

Table 5
Dilution factors for the representative source and long-term average conditions
Downwind distance(km)

Dilution factor(Bq m-3 per Bq s-1)

0.5
1
2
5
10
20
50
100
200
500
1 000
2 000

9.7 10-7
5.3 10-7
2.5 10-7
7.1 10-8
2.5 10-8
8.7 10-9
2.2 10-9
7.6 10-10
2.7 10-10
6.7 10-11
2.4 10-11
8.2 10-12

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

58

Table 6
Analysis of variability in results of the Gaussian plume model
Varied parameter

Value of varied parameter

Dilution factor
at 1 km

Exponentof power function

Wind speed (ui)Twice the representative value
Half the representative value

2.6 10-7
1.1 10-6

1.39
1.55

Mixed layer height (hi)Twice the representative value
Half the representative value

5.3 10-7
5.3 10-7

1.55
1.37

Frequency ofstability class (fi)0.02, 0.05, 0.15, 0.55, 0.15,0.08 for classes A, B, C, D, E, F
(high proportion of neutral classes)
0.02, 0.05, 0.15, 0.25, 0.3, 0.23 for classes A, B, C, D, E, F
(high proportion of stable classes)
0.15, 0.25, 0.3, 0.2, 0.05, 0.05 for classes A, B, C, D, E, F
(High proportion of unstable classes)

5.0 10-7

1.44

5.8 10-7

1.49

4.4 10-7

1.41

0.1 m
1.0 m

5.1 10-7
5.3 10-7

1.46
1.44

Effective releaseheight (H)

0 m (ground-level release)
60 m

1.2 10-6
2.2 10-7

1.70
1.36

Dry deposition
velocity (vd)0.01 m s-1
4 10-4 m s-1
0 m s-1

5.2 10-7
5.3 10-7
5.3 10-7

1.74
1.31
1.21

Wash-out coefficient (Λ)

3 10-4 s-1
3 10-5 s-1

5.3 10-7
5.3 10-7

1.46
1.46

0

5.3 10-7

1.21

Surface roughness (z0)vd and Λ

Table 7
Outdoor effective dose rate to the adult per unit concentration in soil for the significant naturally occurring
radionuclides
Effective dose rate per unit concentration (nSv h-1 per Bq kg-1)
Radionuclide
40
K
Th series
238
U series232ab[B8] a[S10, S11] b[E7] b

0.029
0.46
0.30

0.030
0.42
0.31

0.033
0.51
0.35

Calculated as E = X × 0.0087 Gy R -1 × 0.7 Sv Gy-1.
HE + 0.01 H skin.

Table 8
Conversion coefficients from air kerma to effective dose for terrestrial gamma rays
[S11]
Effective dose per unit air kerma (Sv Gy-1)
Radionuclide
40

K
Th
238
U
Average232

Infants

Children

Adults

0.926
0.907
0.899
0.91

0.803
0.798
0.766
0.79

0.709
0.695
0.672
0.69

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

59

Table 9Effective dose factors for cloud immersion
Effective dose per unit time-integrated concentration in air(nSv per Bq d m-3)

Radionuclide

89

Sr
Sr b
95
Zr
95
Nb
99
Mo b
103
Ru b
106
Ru b
110m
Ag b
115
Cd b
125
Sb b
127
Sb b
129m
Te b
131m
Te b
131
I
132
Te b
133
I
134
Cs
136
Cs
137
Cs b
140
Ba
140
La
141
Ce
143
Ce
144
Ce b
239
Np
90

a
b

[K7] a[E7] a

0.033
0.062
2.9
3.0
1.1
1.8
0.87
11
1.4
1.6
2.6
0.29
6.0
1.4
9.8
2.3
6.0
8.5
2.2
0.72
9.3
0.29
1.0
0.27
0.64

0.039
0.079
3.1
3.3
1.1
2.0
0.99
12
1.6
1.8
3.0
0.32
6.5
1.6
11
2.6
6.6
9.3
2.4
0.76
10
0.31
1.1
0.31
0.68

Calculated as H E + 0.01 H skin.
Decay products included.

Table 10
Collective effective doses from immersion exposure to noble gases released from reactors

Radionuclide

41

Ar
Kr
85
Kr
87
Kr
88
Kr
131m
Xe
133m
Xe
133
Xe
135m
Xe
135
Xe
138
Xe
85m

a
b
c

Half-life

1.827 h4.48 h10.72 a76.3 m
2.84 h11.9 d2.188 d5.245 d15.29 m9.09 h
14.17 mEffective dose rate
per unit
concentration a [E7](nSv per Bq a m-3)
2 080
243
7.92
1 340
3 260
13.8
46.5
50.8
653
385
1 850

Collective dose per unit release b ( man Sv PBq1)
Local

Regional

Total

0.90
0.15
0.007
0.47
1.73
0.012
0.039
0.043
0.062
0.28
0.16

0.005
0.004
0.007
0.001
0.021
0.009
0.013
0.025
-c
0.016
-c

0.90
0.15
0.014
0.47
1.75
0.021
0.052
0.068
0.062
0.30
0.16

Evaluated as H E + 0.01 H skin.
Release from model reactor site; population density 400 km -2 in local area (1-50 km) and 20 km -2 in the regional area (50-2,000 km).
Negligible result.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

60

Table 11
Collective effective dose from immersion exposure for representative composition of noble gases released
from reactors

PWRs

BWRs

PWRs

BWRs

Local

Regional

Total

Local

Regional

Total

Ar
Kr
85
Kr
87
Kr
88
Kr
131m
Xe
133m
Xe
133
Xe
135m
Xe
135
Xe
138
Xe

0.005
0.004
0.016
0.009
0.004
0.006
0.006
0.81
0.002
0.14
0.003

0.03
0.06
0.01
0.08
0.15
0.03
0
0.20
0.06
0.17
0.20

0.005
0.001
0.00010.004
0.007
0.00008
0.00020.035
0.00010.039
0.00050.000020.000020.00010.00001
0.00008
0.00006
0.00008
0.020
0.002
-

0.005
0.00060.00020.004
0.007
0.00010.00030.055
0.00010.041
0.00050.026
0.009
0.00009
0.039
0.25
0.0004
0
0.009
0.004
0.049
0.032

0.00010.00020.00009
0.00010.003
0.0003
0
0.005
0.003
-

0.027
0.010
0.00020.039
0.26
0.0007
0
0.014
0.004
0.052
0.032

Total

1.0

1.0

0.09

0.02

0.11

0.42

0.01

0.43

41

85m

a

Weighted collective dose per unit release a (man Sv PBq1)

Fractional release[U4]

Radionuclide

Collective dose per unit release (values from Table 10) multiplied by the release fraction. The results apply for the model reactor site.

Table 12
Transfer coefficients P25 from deposition to external exposure from radionuclides produced in atmosphericnuclear testingRadionuclide
54

Mn
Zr c
Ru
106
Ru
125
Sb
131
I
137
Cs
140
Ba c
141
Ce
144
Ce c
95

103

a
b
c

Half-life

Absorbed dose rate in air
per unit deposition density a(nGy a-1 per Bq m-2)

Effective dose commitment
per unit deposition density b(nSv per Bq m-2)

312.3 d64.02 d39.26 d373.6 d2.76 a8.02 d30.07 a12.75 d32.5 d284.9 d

12.9
45.0
10.8
3.21
6.52
13.0
8.89
73.5
1.49
0.693

4.02
2.87
0.42
1.19
6.54
0.10
97.2
0.93
0.048
0.20

Ref. [B9]; converted with 0.869 rad per R and 0.01 Gy per rad. Assumes relaxation lengths of 0.1, 1, and 3 cm for radionuclides of half-lives<30 d, 30-100 d, and >100 d, respectively.
Derived from absorbed dose rate in air times 0.7 Sv Gy -1 times 0.36 (occupancy/shielding factor) times mean-life (in years) of radionuclide (1.44
× half-life).
Includes decay product.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

61

Table 13Annual components of dose from external exposure to radionuclides following a single deposition event
Annual effective dose per unit deposition density (nSv per Bq m-2)

Year
following
deposition

131

140

141

103

0.10

0.93

0.048

0. 42
0.001

2.82
0.054
0.001

0.12
0.05
0.020
0.008
0.003
0.001
0.00060.00020.0001

2.23
0.99
0.44
0.20
0.09
0.039
0.017
0.008
0.003
0.002

0.59
0.30
0.15
0.08
0.039
0.020
0.010
0.005
0.003
0.001

1.45
1.13
0.88
0.68
0.53
0.41
0.32
0.25
0.19
0.15

2.21
2.16
2.11
2.07
2.02
1.97
1.93
1.88
1.84
1.80

0.10

0.93

0.048

0.42

2.87

0.20

4.02
0.001

1.19
0.001

6.01
0.49
0.04

20.0
15.9
30.6
21.0
9.69

0.10

0.93

0.048

0.42

2.87

0.20

4.02

1.19

6.54

97.2

I

1
2
3
4
5
6
7
8
9
10
Total
1-10
11-20
21-50
51-100101-"
Commitment
1-"

Ba

Ce

Ru

95

Zr

144

Ce

54

Mn

106

125

Ru

Sb

Table 14
Effective dose equivalent factors for external irradiation outdoors from deposited radionuclides[B9, U4]

Radionuclide

103

Ru
Ru
131
I
134
Cs
137
Cs

106

a
b

Assumes relaxation length in soil of 1 cm.
Assumes relaxation length in soil of 3 cm.

Effective dose equivalent per unit deposition density(nSv per Bq m-2)
30 days to 1 year a

After 1 year b

0.691
2.09
0.015
18.6
8.04

0.00128
1.65
0.0
36.2
264

137

Cs

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

62

Table 15
Estimates of collective dose from external exposure per unit release of radionuclides from nuclear
installations

Radionuclide
51

Cr
Mn
59
Fe
58
Co
60
Co
65
Zn
95
Zr b
103
Ru
106
Ru
124
Sb
131
I
134
Cs
136
Cs
137
Cs
140
Ba b
141
Ce
144
Ce b
241
Am
54

Local

Regional

Total

0.021
4.0
1.1
1.1
71
2.1
2.9
0.42
1.2
2.3
0.10
18
0.92
97
0.93
0.048
0.20
44

0.6
120
33
32
2 100
63
85
13
35
69
3.1
540
27
2 890
28
1.4
5.8
1 310

0.3
54
15
14
940
28
38
5.7
16
31
1.4
240
12
1 300
12
0.65
2.6
590

0.9
170
48
46
3 040
92
120
18
51
100
4.5
780
40
4 190
40
2.1
8.5
1 890

740

340

1 080

Particulates cabcCollective effective dose per unit release a (man Sv PBq-1)

Transfer coefficient
P25(nSv per Bq m-2)

Estimated from dispersion relationship: 5 10 -7 x-1.4, where x is the distance from the release point; deposition velocity = 0.002 m s -1; andpopulation density = 400 km -2 in local area (1-50 km) and 20 km -2 in the regional area (50-2,000 km). Reduction due to urban runoff (factor of0.75) also assumed.
Includes decay product.
Weighted average for assumed representative composition: 13% each of 54Mn, 58Co, 60Co, 89Sr, 134Cs, 137Cs, and 140Ba; 0.9% each of 51Cr, 59Fe,
65
Zn, 90Sr, 90Y, 95Zr, 124Sb, 136Cs, 141Ce, and 144Ce.

Table 16
Age-weighted breathing rate for the world population
Age group

Breathing rate a (m3 d-1)

Fraction of population b

Weighted rate (m3 a-1)

0-12 months
1-2 years
3-7 years
8-12 years
13-17 years
Adults (>17 years)

2.86
5.16
8.72
15.3
20.1
22.2

0.02
0.04
0.10
0.10
0.09
0.65

21
75
320
560
660
5 300

1.0

6 900

Sum
a
b

Ref. [I4].
Estimated from [U15].

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

63

Table 17
Committed effective doses per unit intake by inhalation of radionuclides[I4, I5]

Radionuclide
54

Mn
Fe
89
Sr
90
Sr
91
Y
95
Zr
95
Nb
99
Mo
103
Ru
106
Ru
110m
Ag
115
Cd
125
Sb
127
Sb
129m
Te
131m
Te
132
Te
131
I
133
I
134
Cs
136
Cs
137
Cs
140
Ba
140
La
141
Ce
143
Ce
144
Ce
55

a
b
c
d

Absorption
type a
M
M
M
M
M
M
M
M
M
M
M
M
M
M
M
M
M
F
F
F
F
F
M
M
M
M
M

Effective dose per unit intake (nSv Bq-1)
Infants bChildren c

Adults d

6.2
1.4
24
110
30
16
5.2
4.4
8.4
110
28
4.8
16
7.3
26
5.8
13
72
18
7.3
5.2
5.4
20
6.3
11
3.9
160

2.4
0.62
9.1
51
11
6.8
2.2
1.5
3.5
41
12
1.7
6.8
2.7
9.8
1.9
4.0
19
3.8
5.3
2.0
3.7
7.6
2.0
4.6
1.3
55

1.5
0.38
6.1
36
7.1
4.8
1.5
0.89
2.4
28
7.6
0.98
4.8
1.7
6.6
0.94
2.0
7.4
1.5
6.6
1.2
4.6
5.1
1.1
3.2
0.75
36

Absorption rates in body fluids are fast (F), moderate (M), and slow (S).
From 1 year to 2 years.
More than 7 years to 12 years.
More than 17 years.

Radionuclide
238

232

Absorption
type aEffective dose per unit intake (nSv Bq-1)
Infants bChildren c

Adults d

U series:
238
U
234
U
230
Th
226
Ra
210
Pb
210
Po

M
M
S
M
M
M

9 400
11 00035 000
11 000
3 700
11 0004 000
4 800
16 000
4 900
1 500
4 600

2 900
3 500
14 000
3 500
1 100
3 300

Th series:
232
Th
228
Ra
228
Th

S
M
S

50 000
10 000130 000

26 0004 600
55 00025 0002 600
40 000M
S
M

10 00069 000550 000

4 300
39 000260 000

3 100
34 000220 000

M
M
M
M
M
M

4.2
74 00077 00077 000970
69 0001.4
44 00048 00048 000830
40 0000.93
46 00050 00050 000900
42 000235

U series:
235
U
231
Pa
227
Ac
239

Np
Pu
239
Pu
240
Pu
241
Pu
241
Am
238

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

64

Table 18
Transfer coefficients for the inhalation pathway applicable to the deposition of radionuclides produced in
atmospheric nuclear testingRadionuclide
54

Mn
Fe
89
Sr
90
Sr
91
Y
95
Zr
95
Nb
103
Ru
106
Ru
125
Sb
131
I
137
Cs
140
Ba
141
Ce
144
Ce
238
Pu
239
Pu
240
Pu
241
Pu
241
Am
55

a
b

Effective dose per unit intake a
P45(nSv Bq-1)

Effective dose per unit deposition density b
P245(nSv per Bq m-2)

1.5
0.38
6.1
36
7.1
4.8
1.5
2.4
28
4.8
7.4
4.6
5.1
3.2
36
46 00050 00050 000900
42 0000.020
0.0050
0.080
0.47
0.093
0.063
0.020
0.032
0.37
0.063
0.097
0.061
0.067
0.042
0.47
610
660
660
12
550

Absorption assumed to be Type F (fast) for 131I and 137Cs and Type M (moderate) for all other radionuclides.
Equal to P14P45/P12, where P14 = 20 m3 d-1 (adult breathing rate) and P 12 = 0.0176 m s -1 (the deposition velocity applicable to fallout fromatmospheric testing).

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

65

Table 19
Estimates of collective dose from inhalation exposure per unit release of radionuclides from nuclear
installations

Radionuclide
51

Cr
Mn
55
Fe
59
Fe
58
Co
60
Co
65
Zn
89
Sr
90
Sr
90
Y
91
Y
95
Zr
103
Ru
106
Ru
124
Sb
131
I
134
Cs
136
Cs
137
Cs
140
Ba
141
Ce
144
Ce
238
Pu
239
Pu
240
Pu
241
Pu
241
Am
54

Particulates babCollective effective dose per unit release a (man Sv PBq1)

Transfer coefficient
P245(nSv per Bq m-2)

Local

Regional

Total

0.0037
0.17
0.044
0.43
0.19
1.2
0.19
0.71
4.2
0.16
0.82
0.56
0.28
3.2
0.74
0.86
0.76
0.14
0.53
0.13
0.37
4.2
5 320
5 790
5 790
100
4 860

0.15
6.9
1.7
17
7.3
46
7.3
28
165
6.4
33
22
11
130
29
34
30
5.5
21
5.0
15
165
211 000
229 000
229 000
4 130
193 000

0.07
3.1
0.8
7.6
3.3
21
3.3
13
74
2.9
15
9.9
5.0
58
13
15
14
2.5
9.5
2.3
6.6
74
95 000103 000
103 000
1 860
86 700

0.21
10
2.5
25
11
66
11
41
240
9.3
47
32
16
190
43
49
44
8.0
31
7.3
21
240
306 000
332 000
332 000
5 990
279 000

23

10

33

Estimated from dispersion relationship: 5 10 -7 x-1.4, where x is the distance from the release point; deposition velocity = 0.002 m s -1; andpopulation density = 400 km -2 in local area (1-50 km) and 20 km -2 in the regional area (50-2,000 km).
Weighted average for assumed representative composition: 13% each of 54Mn, 58Co, 60Co, 89Sr, 134Cs, 137Cs, and 140Ba; 0.9% each of 51Cr, 59Fe,
65
Zn, 90Sr, 90Y, 95Zr, 124Sb, 136Cs, 141Ce, and 144Ce.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

66

Table 20
Committed effective doses per unit intake by ingestion of natural radionuclides[I5]
Effective dose per unit intake (nSv Bq-1)
Radionuclide3
H (water)
H (organic)

3

7

Be

Fractional absorption
Infants a

Children b

Adults c

1.0
1.0

0.048
0.12

0.023
0.057

0.018
0.042

0.005

0.13

0.053

0.028

14

C

1.0

1.6

0.80

0.58

22

Na

1.0

15

5.5

3.2

40

1.0

42

13

6.2

0.02
0.02
0.0005
0.2
0.2
0.5

120
130
410
960
23
3 600
8 800

68
74
240
800
5.9
1 900
2 600

45
49
210
280
3.5
690
1 200

0.00050.2
0.0005

450
5 700
370

290
3 900
150

230
690
72

0.02
0.00050.0005

130
1 300
3 100

71
920
1 500

47
710
1 100

K

238

U series:
238
U
234
U
230
Th
226
Ra
222
Rn d
210
Pb
210
Po

232

Th series:
232
Th
228
Ra
228
Th

235

U series:
235
U
231
Pa
227
Ac

a
b
c
d

From 1 year to 2 years.
More than 7 years to 12 years.
More than 17 years.
Ref. [N5].

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

67

Table 21
Food consumption rates by individuals a
Consumption rate (kg a-1)
Country /regionPopulation
Milk

Grain

Leafy
vegetables

Fruit/
vegetables

Meat

Total

North Europe
Denmark
Finland
NorwaySweden

5.11
4.87
4.16
8.35

173
263
202
222

80
73
65
77

18
6b
37
36

150
169
120
121

66
71
76
56

487
582
500
512

Central Europe
Austria
Czechoslovakia
Germany
Hungary
PolandRomania
Switzerland

7.56
15.48
77.66
10.62
37.46
22.73
6.49

145
134
109
185
160
150
180

66
132
84
110
180
190
99

71
25
28
25
20
40
29

136
107
145
160
132
240
230

99
86
63
80
67
86
110

517
484
429
560
559
706
648

West Europe
Belgium
FranceIreland
Luxembourg
Netherlands
United Kingdom

9.86
53.6
3.54
0.37
14.49
55.87

180
130
163
110
145
163

65
84
68
95
65
68

55
84
40
33
65
40

150
132
69
150
135
100

40
73
50
88
70
71

490
503
390
476
480
442

South Europe
Bulgaria
Greece
Italy
PortugalSpain
Yugoslavia

8.89
9.83
56.91
9.94
37.3
22.49

123
80
90
45
104
146

179
100
110
125
88
146

20
30
50
113
124
55

76
250
150
105
132
128

64
60
60
42
62
55

462
520
460
430
510
530

USSR

279

332 b

133

37

118

63

683

West Asia
CyprusIsraelSyrian Arab Rep.
Turkey

0.64
3.87
8.98
52

83
120
70
125

94
130
190
200

87
140
30
100

315 b
190
340 b
150

83
60
22
40

662
640
652
615

East Asia
China
India
Japan

1046.4
750.9
121.0

5b
39
50

229
183
193

29
28
30

173
89
180

30
5b
120

466
344
573

North America
CanadaUnited States

25.4
238.7

181
174

93
91

21
25

301 b
260

130
146

726
696

Average values cCountries ofEast and West Asia

25

210

30

140

25

430

Countries ofEurope, USSR, andNorth America

200

110

40

165

85

600

85

170

35

150

50

490

World

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

68

Table 21, continued
Consumption rate (kg a-1)
Country /regionPopulation
Milk

Grain

Leafy
vegetables

Fruit/
vegetables

Meat

Total

Representative values dCountries ofEast and West Asia

90

210

30

140

60

530

Countries ofEurope, USSR andNorth America

150

110

40

170

85

555

120

170

35

150

70

545

World
a
b
c
d

Population and consumption rates valid for 1986 [U4].
Unusually high or low values.
Average values are population-weighted results.
Rounded, generic values (unusually high and low values excluded).

Table 22
Parameters of empirical models for transfer of 90Sr and 137Cs from deposition to diet to dose a
Pathway

Sr

137

Cs

b1 (Bq a kg-1 per Bq m-2)
b2 (Bq a kg-1 per Bq m-2)
b3 (Bq a kg-1 per Bq m-2)
λ' (a-1)
P23 (Bq a kg-1 per Bq m-2)

0.001
0.001
0.000110.06
0.00380.00380.00290.000052
0.03
0.0084

Diet to bodyc (Bq a kg-1 per Bq a kg-1)
g (Bq a kg-1 per Bq a kg-1)
λb (a-1)
P34 (Bq a kg-1 per Bq a kg-1)

17.5
3.7
0.13
48

2.6

Body to dose-1

P45 (nSv per Bq a kg )

290

2 500

Diet to intake b

P34 (Bq per Bq a kg-1)

500

500

P45 (nSv per Bq)

28

13

P2345 (nSv per Bq m-2 )53

55

Deposition to doseb

90

Deposition to diet

Diet to doseaTransfer parameter

Annual dose in a specific year is the deposition density of 90Sr or 137Cs in that year times the annual component of P 23 times the annualcomponent of P 34 times P45 plus the contribution from intake in earlier years, which equals the residual body burden (for 90Sr) reduced byexponential decay and removal (e -&b) times P45.
Assumes consumption intake of food of 500 kg a -1.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

69

Table 23
Transfer coefficients for radionuclides in the ingestion pathway

Radionuclide

Deposition to diet
P23
(mBq a kg-1 per Bq m-3)

Deposition to intake a
P234(Bq per Bq m-2)

Intake to dose b
P45(nSv Bq-1)

Deposition to dose
P2345
(nSv per Bq m-2)

7c
25 c

0.56
2
6
0.76
2.1
2.9
3.6
0.03
1.9
0.1
0.07
1
0.07
2
0.3
4.2
0.005
0.07
0.1
0.05
0.7
0.7
0.04
0.2
0.04

0.038
0.71
0.33
1.8
0.74
3.4
3.9
2.6
28
0.95
0.58
2.5
61 e
19
3
13
2.6
0.71
5.2
230
250
250
4.8
200
120

0.02
1.4
2.0
1.4
1.6
9.9
14
0.08
53
0.10
0.04
2.5
4.3
38
0.90
55
0.013
0.05
0.52
12
180
180
0.19
40
5

51

Cr
Mn
55
Fe
59
Fe
58
Co
60
Co
65
Zn
89
Sr
90
Sr
95
Zr
95
Nb
124
Sb
131
I
134
Cs
136
Cs
137
Cs
140
Ba
141
Ce
144
Ce
238
Pu
239
Pu
240
Pu
241
Pu
241
Am
244
Cm
54

a
b
c
d
e

26 c
36 c
45 c
3.8
1.3 c
0.9 c
13 c
0.6 d
4
0.6
8.4
0.9 c
1.3 c

May be derived from P 23 by multiplying by total dietary consumption of 500 kg a -1.
To adults unless otherwise stated.
To grain. To derive P 24, grain consumption of 80 kg a -1 has been assumed.
For milk. To derive P24, milk consumption of 0.3 l d -1 has been assumed.
Population-weighted value.

Table 24
Annual components of dose from ingestion exposure to radionuclides following a single deposition eventYearfollowing
deposition
1
2
3
4
5
6
7
8
9
10
Total
1-10
11-20
21-50
51-100101-"
Commitment
1-"aAnnual effective dose per unit deposition density (nSv per Bq m-2)
131

140

4.2

0.013

4.2

4.2

I

55

90

0.08
0.0005

1.00
0.60
0.089
0.069
0.054
0.042
0.033
0.025
0.020
0.015

6.15
7.73
2.47
2.30
2.14
1.99
1.86
1.73
1.62
1.51

24.7
19.2
0.32
0.31
0.30
0.29
0.28
0.27
0.27
0.26

0.013

0.08

1.95
0.049
0.004

29.5
10.7
10.3
1.9
0.10

46.2
2.2
3.7
2.0
0.49

0.013

0.08

2.0

53

55

Ba

89

Sr

Fe a

Sr

137

Cs

A transfer model does not exist. Using 137Cs as a guide, it is assumed that 50% of commitment arises in first year after deposition, 30% in second
year, and remainder at uniform rate over the mean life of 55Fe.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

70

Table 25
Estimates of collective dose from ingestion exposure per unit release of radionuclides from nuclear fuelcycle installations

Radionuclide
Hb
Cb
51
Cr
54
Mn
55
Fe
59
Fe
58
Co
60
Co
65
Zn
89
Sr
90
Sr
95
Zr
124
Sb
131
I
134
Cs
136
Cs
137
Cs
140
Ba
141
Ce
144
Ce
238
Pu
239
Pu
240
Pu
241
Pu
241
Am

Local

Regional

Total

0.021
1.4
2.0
1.4
1.6
9.9
14
0.078
53
0.10
2.5
4.3
38
0.90
55
0.013
0.050
0.52
12
180
180
0.19
40

1.1
190
0.8
56
79
54
62
390
560
3.1
2 110
3.8
99
170
1 510
36
2 160
0.5
2.0
21
460
6 930
6 930
7.6
1 580

1.0
80
0.4
25
36
24
28
180
250
1.4
950
1.7
45
76
680
16
970
0.2
0.9
9.3
210
3 120
3 120
3.4
710

2.1
270
1.2
82
110
79
89
570
810
4.5
3 060
5.5
144
250
2 180
52
3 140
0.7
2.9
30
660
10 10010 100
11
2 300

570

260

830

3

14

Particulates cabcCollective effective dose per unit release (man Sv PBq1) a

Transfer coefficient
P2345(nSv per Bq m-2)

Population density: local (1-50 km): 400 persons km 2; regional (50-200 km) 20 persons km -2.
Doses estimated using specific-activity model.
Weighted average for assumed representative composition: 13% each of 54Mn, 58Co, 60Co, 89Sr, 134Cs, 137Cs, and 140Ba; 0.9% each of 51Cr, 59Fe,
65
Zn, 90Sr, 90Y, 95Zr, 124Sb, 136Cs, 141Ce, and 144Ce.

Table 26
Population densities surrounding nuclear fuel cycle installations
Population density surrounding nuclear fuel cycle sites (inhabitants km-2)
Country / region
World average a

a
b
c

Area
Local bRegional c

Uranium mining

Fuel fabrication

Reactors

3
25

25

400
20

Representative values used in UNSCEAR assessments.
0-100 km for mining; 0-50 km for reactors.
100-2,000 km for mining and fuel fabrication; 50-2,000 km for reactors.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

71

Table 27
Collective dose per unit release of radionuclides in liquid effluents to fresh water

Radionuclide

3

H
C
24
Na
35
S
45
Ca
51
Cr
54
Mn
55
Fe
57
Co
58
Co
59
Fe
60
Co
65
Zn
89
Sr
90
Sr
95
Zr
95
Nb
97
Zr
99
Mo
103
Ru
106
Ru
110m
Ag
113
Sn
122
Sb
124
Sb
125
Sb
129
I
131
I
132
Te
133
I
135
I
134
Cs
136
Cs
137
Cs
140
Ba
141
Ce
143
Ce
144
Ce
147
Pm
239
Pu
14

a

12.26 a5 730 a14.36 h87.5 d162.2 d27.7 d312.1 d2.73 a271.8 d70.8 d44.5 d5.271 a244.3 d50.5 d28.78 a64.02 d34.98 d16.9 h
2.75 d39.26 d373.6 d249.8 d115.1 d2.73 d60.2 d2.76 a1.6 107 a
8.02 d3.2 d20.8 h
6.57 h2.06 a13.16 d30.07 a12.75 d32.5 d1.38 d284.9 d2.623 a24 110 a

Time integralof unit activity
in water a(Bq a)

Drinking water
treatment
removal factor

3.90
5.00
0.0024
0.32
0.57
0.11
0.99
2.20
0.88
0.27
0.17
3.02
0.81
0.19
4.46
0.24
0.13
0.0028
0.011
0.15
1.14
0.82
0.42
0.011
0.23
2.22
5.00
0.032
0.013
0.00340.0011
1.49
0.051
2.81
0.050
0.12
0.0054
0.82
2.15
3.00

1
1
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.8
0.8
0.5
0.8
0.8
0.2
0.2
0.2
0.5
0.1
0.1
0.1
0.1
0.1

Concentration factor
for fish b(Bq kg-1 fish
per Bq l-1 water)

Dose per unit
activity
ingested(nSv Bq-1)

1
50 000
20
800
20
200
400
200
300
300
200
300
1 000
60
60
300
300
300
10
10
10
5
3 000
100
100
100
40
40
400
40
40
2 000
2 000
2 000
4
30
30
30
30
30

0.018
0.58
0.43
0.13
0.71
0.038
0.71
0.33
0.21
0.75
1.8
3.4
3.9
2.6
28
0.96
0.59
2.1
0.6
0.73
7.0
2.8
0.73
1.7
2.5
1.1
110
22
3.8
4.3
0.93
19
3.0
13
2.6
0.71
1.1
5.2
0.26
250

Collective dose per unit activity
released (man Sv PBq-1)
Drinking
water

Fish

Total

1.3
54
0.009
0.39
3.7
0.038
6.5
6.7
1.7
1.8
2.8
95
29
4.6
1 150
2.1
0.73
0.054
0.060
1.0
74
21
2.8
0.17
5.2
23
8 120
10
0.44
0.22
0.015
100
0.57
130
1.2
0.16
0.011
7.9
1.0
1 380

0.003
6 690
0.0009
1.6
0.37
0.038
13
6.7
2.6
2.8
2.8
140
150
1.4
350
3.2
1.1
0.081
0.003
0.051
3.7
0.53
42
0.084
2.6
11
1 020
1.3
0.89
0.027
0.002
2 620
14
3 370
0.02
0.12
0.008
5.9
0.78
1 040

1.3
6 740
0.010
1.9
4.1
0.075
19
13
4.3
4.6
5.6
240
180
6.0
1 500
5.3
1.8
0.13
0.063
1.1
77
22
45
0.25
7.9
34
9 140
12
1.3
0.24
0.017
2 720
15
3 500
1.2
0.28
0.019
14
1.8
2 420

Time integrals of unit activity in freshwater were calculated from the empirically derived values of the mean residence times in water of 90Sr andCs (5 and 3 years, respectively), assuming that those radionuclides with high K d, i.e. 144Ce and 239Pu, behave as 137Cs and the otherradionuclides behave as 90Sr, in both cases correcting for physical delay. The formula is: A 0 (J + 8)-1, where A0 is unit activity (1 Bq), J is thereciprocal of the mean residence time, and 8 is ln 2 / half-life.
Ref. [I9].

137bHalflife

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

72

Table 28
Collective dose per unit release of radionuclides in liquid effluents to salt water

Radionuclide

3

H
C
24
Na
35
S
45
Ca
51
Cr
54
Mn
55
Fe
57
Co
58
Co
59
Fe
60
Co
65
Zn
89
Sr
90
Sr
95
Zr
95
Nb
97
Zr
99
Mo
103
Ru
106
Ru
110m
Ag
113
Sn
122
Sb
124
Sb
125
Sb
129
I
131
I
132
Te
133
I
135
I
134
Cs
136
Cs
137
Cs
140
Ba
141
Ce
143
Ce
144
Ce
147
Pm
239
Pu
14

a
b

Halflife

12.26 a5 730 a14.36 h87.5 d162.2 d27.7 d312.1 d2.73 a271.8 d70.8 d44.5 d5.271 a244.3 d50.5 d28.78 a64.02 d34.98 d16.9 h
2.75 d39.26 d373.6 d249.8 d115.1 d2.73 d60.2 d2.76 a1.6 107 a
8.02 d3.2 d20.8 h
6.57 h2.06 a13.16 d30.07 a12.75 d32.5 d1.38 d284.9 d2.623 a24 110 a

Time integralof unit activity
in water a(Bq a)
2.56
3.00
0.0024
0.31
0.53
0.11
0.87
1.70
0.79
0.26
0.17
2.15
0.73
0.19
2.80
0.23
0.13
0.0028
0.011
0.15
0.99
0.74
0.40
0.011
0.22
1.71
3
0.031
0.013
0.00340.0011
1.49
0.051
2.81
0.050
0.12
0.0054
0.82
1.67
3.50

Concentration factor b(Bq kg-1 fish per Bq l-1 water)
Fish

Shellfish(crustacea)1
20 000
0.1
2
2
200
400
3 000
1 000
1 000
3 000
1 000
1 000
2
2
20
30
20
2
2
2
500
50 000
400
400
400
10
10
1 000
10
10
100
100
100
10
50
50
50
500
40

1
20 000
0.1
1
5
500
500
5 000
5 000
5 000
5 000
5 000
50 000
2
2
200
200
200
10
100
100
5 000
50 000
400
400
400
10
10
1 000
10
10
30
30
30
1
1 000
1 000
1 000
1 000
300

Dose per unit
activity
ingested(nSv Bq-1)
0.018
0.58
0.43
0.13
0.71
0.038
0.71
0.33
0.21
0.75
1.8
3.4
3.9
2.6
28
0.96
0.59
2.1
0.6
0.73
7.0
2.8
0.73
1.7
2.5
1.1
110
22
3.8
4.3
0.93
19
3.0
13
2.6
0.71
1.1
5.2
0.26
250

Collective dose per unit activity released(man Sv PBq-1)
Fish

Shellfish(crustacea)

Total

0.0012890
0.000003
0.0021
0.019
0.021
6.4
43
4.3
4.9
23
190
73
0.025
4.0
0.12
0.060
0.00300.00030.0055
0.36
27
370
0.19
5.7
19
85
0.18
1.2
0.00380.0003
73
0.39
94
0.033
0.11
0.00775.5
5.6
900

0.0002150
0.00020.00800.0086
1.3
12
3.6
4.1
6.4
160
610
0.0042
0.67
0.19
0.067
0.00500.0003
0.046
3.0
45
62
0.031
0.94
3.2
14
0.030
0.21
0.00060.00004
3.6
0.020
4.7
0.0006
0.37
0.026
18
1.9
1120

0.00141 040
0.000003
0.0022
0.027
0.029
7.7
55
7.8
9.1
29
350
680
0.029
4.7
0.31
0.13
0.00800.0006
0.052
3.3
71
430
0.22
6.6
23
99
0.21
1.4
0.00440.0003
77
0.41
98
0.034
0.49
0.033
24
7.5
2 020

Obtained from estimated mean residence times in water of 3 years for 90Sr, 137Cs and other radionuclides and 3.5 years for 239Pu. The time integralof unit activity is 1/( J+8), where J is the reciprocal of the mean residence time and 8 is ln2 / half-life.
Ref. [I10].

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

73

Table 29
Collective effective dose for representative composition of particulates released from reactors in liquid
effluents

Radionuclide

Fractional
release

58

Co
Co
51
Cr
131
I
137
Cs
24
Na
54
Mn
65
Zn
134
Cs
133
I
55
Fe
59
Fe
89
Sr
95
Nb
110m
Ag
125
Sb
135
I
140
Ba

0.20
0.20
0.10
0.10
0.10
0.05
0.05
0.05
0.05
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01

Total
Average

1.0

60Collective dose per unit release(man Sv PBq1)

Contribution to collective dose per unit total release(man Sv PBq1)

Freshwater

Saltwater

Freshwater

Saltwater

4.6
240
0.075
12
3 500
0.010
19
170
2 720
0.24
13
5.6
6.0
1.8
22
34
0.017
1.2

9.1
350
0.029
0.21
98
0.000003
7.7
680
77
0.0044
55
29
0.029
0.13
71
23
0.0003
0.034

0.92
47
0.00751.2
3500.00051.0
8.7
1400.0049
0.13
0.056
0.060
0.018
0.22
0.34
0.0002
0.012

1.8
69
0.0029
0.021
9.8
0.39
34
3.8
0.00009
0.55
0.29
0.00030.00130.71
0.23
0.0003

550

120
330

Table 30
Parameters of the seven-compartment model of the world hydrological cycle [N3]Compartment

Northern hemisphere

Southern hemisphere

World

Volume (1012 m3)
Atmospheric water
Soil water aFreshwater
Saline water
Groundwater a
Ocean surface b
Deep ocean b

6.33
45.13
95
100
5 624
11 568553 980

6.67
21.87
31
4
2 726
15 432739 020

13
67
126
104
8 350
27 0001 293 000

Transfer rate (1012 m3 a-1)
Atmosphere

Soil

Freshwater
Saline water
Groundwater

Ocean
Deep ocean
a
b
c

–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–

Soil
Freshwater
Saline water
Ocean
AtmosphereFreshwater
Groundwater
Atmosphere
Ocean
Atmosphere
Soil
Freshwater
Saline water
Atmosphere
Deep ocean
Ocean

66.85
0.452
0.096
137.1
45.91
19.80
17.11
0.75
20.25
0.48
15.97
0.754
0.385
157.4 c
685.5
685.5

32.45
0.148
0.004
190.3 c
22.59
9.60
8.29
0.25
9.75
0.02
8.03
0.246
0.015
200.0
914.5
914.5

99.3
0.6
0.1
320
68.5
29.4
25.4
1.0
30.0
0.5
24.0
1.0
0.4
350
1 600
1 600

Land surface area: 67.35% in northern hemisphere, 32.65% in southern hemisphere.
Ocean surface area: 42.84% in northern hemisphere, 57.16% in southern hemisphere.
Transfer of 7.4 10 12 m3 a-1 from ocean surface to atmosphere (northern hemisphere) and atmosphere to ocean surface (southern hemisphere)
added to achieve balance.

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

74

Table 31
Comparison of model estimates of global collective doses from tritium released to the atmosphere

Model
NCRP [N3] Seven-compartment model
30--50- northern hemisphereNorthern hemisphere
Southern hemisphere
World

Normalized collective dose a(man Sv PBq-1)

0.38
0.13
0.11
0.06

0.7
0.67
0.07
0.35

Bergmann et al. [B4]0--90- northern troposphere
0--90- southern troposphere
Whole stratosphere

0.95
0.65
0.76

Killough and Kocher [K2]World troposphereNorthern troposphere30--50- northern troposphere

0.94
1.4
2.3

Natural tritium production
Northern hemisphere
Southern hemisphereWorldaNormalized effective dose(nSv PBq-1)

0.27
0.27
0.14

1.5
0.2
0.8

World population: 6 10 9.

Table 32
Results of model calculation of release of 1 PBq of tritium to the atmosphere a
Integrated concentrations in 70-year period (Bq a m-3)
Regiona

Release to
30o-50o N latitude

Release tonorthern hemisphere

Release to
southern hemisphere

Release toworld

Atmosphere
Surface soil water
Freshwater
Saline water
Groundwater
Ocean surface
Deep ocean
Man

24.3
17.0
6.9
0.12
0.83
2.9
0.059
14.7

6.5
5.3
4.1
0.12
0.26
0.66
0.014
4.9

5.4
4.4
3.6
0.10
0.22
0.54
0.011
4.3

3.0
2.4
1.9
0.055
0.12
0.30
0.0061
2.3

Effective dose commitment (nSv)

0.38

0.13

0.11

0.06

Obtained with use of seven-compartment model [N3].

1 000
945
893
844
798
754
713
674
637
602
569
430
324
245
184
139
105
79
60
45
34
26
19
0
39
76
108
136
158
177
191
202
211
217
219
197
166
135
108
84
65
50
39
29
22
17
0
704
664
607
550
497
448
403
362
325
291
167
95
55
32
18
11
6.5
3.9
2.4
1.6
1.0
0.7

0
60
71
71
68
65
62
59
56
53
51
39
30
22
17
13
10
7.2
5.4
4.0
3.0
2.3
1.7

0
0.54
0.53
0.52
0.50
0.48
0.46
0.44
0.42
0.40
0.38
0.30
0.24
0.18
0.14
0.11
0.09
0.07
0.05
0.04
0.03
0.02
0.02

0
63.6
62.9
51.5
40.4
31.5
24.5
19.2
15.1
11.9
9.4
3.28
1.37
0.67
0.37
0.22
0.13
0.08
0.06
0.04
0.03
0.02
0.01

0
75.0
17.8
5.4
2.6
1.9
1.6
1.4
1.3
1.1
1.0
0.60
0.36
0.21
0.13
0.082
0.052
0.034
0.023
0.015
0.011
0.00740.0053

1 000
2.9
0.90
0.46
0.33
0.28
0.25
0.22
0.20
0.18
0.16
0.09
0.053
0.031
0.018
0.011
0.00640.00390.00240.00160.00100.000680.00047

0
1
2
3
4
5
6
7
8
9
10
15
20
25
30
35
40
45
50
55
60
65
70

Total

Total

Deep ocean

Ocean surface

Groundwater

Saline water

Freshwater

Surface soil water

Atmosphere

Activity (TBq)

Time after
release(years)

10.23
1.43
0.65
0.41
0.31
0.25
0.20
0.17
0.14
0.12
0.054
0.028
0.016
0.00930.00570.00360.00230.00150.00100.000700.000490.00035
14.7

Integrated
concentration in man(Bq a m-3)

0.38

0.26
0.037
0.017
0.011
0.00800.00640.00520.00430.00360.00300.00140.00072
0.000410.000240.000150.000092
0.000059
0.000039
0.000026
0.000018
0.000013
0.000009

Annual

0.26
0.30
0.32
0.33
0.33
0.34
0.35
0.35
0.35
0.36
0.37
0.37
0.37
0.38
0.38
0.38
0.38
0.38
0.38
0.38
0.38
0.38

Cumulative

Effective dose (nSv)

Table 33
Distribution of tritium and dose from release of 1 PBq to the atmosphere of the 30o-50o N band of the northern hemisphere; results of seven-compartment model

ANNEX A: DOSE ASSESSMENT METHODOLOGIES
75

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

76

Table 34
Results of model calculation of the release of 1 PBq of 14C to the environment[T1]
Inventory of 14C
(TBq)
Year
Atmosphere

Groundvegetation

Inventory ofstable carbon(1012 g)groundvegetation a

Integrated specific activity(Bq g-1)
Atmosphere b

Groundvegetation b

Effective dose(µSv)

Groundvegetation a

Annualdose aCumulative
dose a0.000130.00056
0.00260.00520.0080
0.012
0.016
0.021
0.030
0.040
0.057
0.095
0.13
0.17
0.18

0.0076
0.024
0.038
0.024
0.011
0.00560.00350.00230.00140.00110.000860.00058
0.000310.000088
0.000002

0.0076
0.032
0.13
0.27
0.41
0.61
0.85
1.1
1.6
2.1
3.0
4.9
6.9
8.5
9.2

Release to atmosphere
1
2
5
10
20
50
100
200
500
1 000
2 000
5 000
10 000
20 00050 000

885
712
437
253
137
76.3
50.0
33.0
21.1
16.3
13.7
9.41
5.02
1.43
0.033

18.6
40.4
46.4
28.7
13.9
7.18
4.65
3.05
1.94
1.50
1.26
0.866
0.462
0.131
0.0030Inventory of 14C
(TBq)
Year
Ocean
surface

Groundvegetation

69 10069 30070 00070 90071 80073 20074 50076 00078 50080 40082 30084 40084 60083 90082 000Inventory ofstable carbon(1012 g)groundvegetation a0.000590.00170.00390.00610.0085
0.012
0.017
0.022
0.032
0.045
0.064
0.11
0.16
0.20
0.22

0.000130.00056
0.00260.00530.0081
0.012
0.016
0.022
0.032
0.045
0.064
0.11
0.16
0.20
0.22

Integrated specific activity(Bq g-1)
Ocean
surface b

Groundvegetation b

Effective dose(µSv)

Groundvegetation a

Annualdose aCumulative
dose a0.000001
0.00001
0.00009
0.000400.00120.00350.0062
0.010
0.019
0.029
0.045
0.083
0.12
0.16
0.18

0.00005
0.000330.00160.00350.00470.00420.00310.00230.00160.00120.00090.00070.00050.00020.00005

0.00005
0.00050
0.0050
0.022
0.067
0.18
0.32
0.53
1.0
1.5
2.4
4.3
6.3
7.9
8.6

Release to ocean surface
1
2
5
10
20
50
100
200
500
1 000
2 000
5 000
10 000
20 00050 000
a
b

977
914
777
599
368
137
81
53
30
23
19
14
7.0
2.0
0.046

0.124
0.69
3.2
5.6
6.3
4.6
3.5
2.6
1.8
1.4
1.2
0.90
0.46
0.13
0.003069 10069 30070 00070 90071 80073 20074 50076 00078 50080 40082 30084 40084 60083 90082 0000.000470.00140.00380.0071
0.012
0.019
0.024
0.030
0.042
0.055
0.075
0.12
0.17
0.21
0.24

0.00000
0.00001
0.00009
0.000410.00130.00360.0066
0.011
0.021
0.032
0.052
0.098
0.15
0.19
0.22

Assuming variable inventory of carbon in environment as a result of input from burning of fossil fuels.
Assuming fixed inventories of carbon in the environment: 750 10 15 g (atmosphere); 1,050 10 15 g (Atlantic Ocean surface) and 69 10 15 g (groundvegetation).

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

77

Table 35
Estimates of collective dose to the world population per unit release of 129I to different environmental
compartments calculated using a global circulation model[T1]
Collective effective dose per unit release (man Sv TBq-1)
Time(years)

1
2
5
10
20
50
100
200
500
1 000
2 000
5 000
10 000
20 00050 000100 000
1 000 000
10 000 000
100 000 000

Release toland atmosphere

Release to
ocean atmosphere

Release to
ocean mixed layer

Release to
deep ocean

Release tosolid soil

67.1
81.7
81.8
82.0
82.4
83.4
85.1
88.4
98.0
113
138
192
236
263
282
303
450
643
727

13.9
19.3
19.4
19.4
19.5
19.8
20.2
21.0
23.3
26.9
33.2
46.8
59.0
69.5
85.6
106
254
446
530

0.00432
0.01580.04200.06490.08040.0908
0.104
0.130
0.212
0.361
0.698
1.94
4.43
9.82
25.0
45.8
193
385
469

0.0000036
0.000032
0.000288
0.00106
0.003130.009960.02160.0453
0.121
0.259
0.577
1.78
4.24
9.62
24.8
45.6
193
385
469

0.0292
0.0977
0.305
0.649
1.34
3.40
6.80
13.5
32.9
62.9
115
223
309
359
382
403
551
744
828

ANNEX A: DOSE ASSESSMENT METHODOLOGIES

79

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