United Nations Scientific Committee on the Effects of Atomic Radiation

S O U R C E S , E F F E C T S A N D R I S K S O F I O N I Z I N G R A D I AT I O N

UNSCEAR 2016 Report

Report to the General Assembly
SCIENTIFIC ANNEXES A, B, C and D

SOURCES, EFFECTS AND RISKSOF IONIZING RADIATION
United Nations Scientific Committee on the
Effects of Atomic Radiation
UNSCEAR 2016
Report to the General Assembly,
with Scientific Annexes

UNITED NATIONSNew York, 2017

NOTE
The report of the Committee without its annexes appears as Official Records of the General
Assembly, Seventy-first Session, Supplement No. 46 and corrigendum (A/71/46 and Corr.1). The
report ­reproduced here includes the corrections of the corrigendum.
The designations 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, or­concerning 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 thedata 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
Sales No. E.17.IX.1
ISBN: 978-92-1-142316-7
eISBN: 978-92-1-060002-6© United Nations, January 2017. All rights reserved, worldwide.
This publication has not been formally edited.
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A/71/46/Corr.1

United Nations

General Assembly
Official Records
Seventy-first session
Supplement No. 46 (A/71/46)

13 October 2016
Original: English

Report of the United Nations Scientific Committee on the
Effects of Atomic Radiation
Corrigendum
Paragraph 48
The fourth sentence should readWith regard to the construction phase of the electricity-generating technologies, by
far the largest collective dose to workers per unit of electricity generated was foundin the solar power cycle, followed by the wind power cycle.

V.16-08901 (E)*1608901*

CONTENTS
Page

Report of the United Nations Scientific Committee on the Effects of Atomic Radiation
to the General Assembly .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .

1

Scientific Annexes
Annex A.	 Methodology for estimating public exposures due to radioactive discharges.  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 19
Annex B.	 Radiation exposures from electricity generation .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 135
Annex C.	 Biological effects of selected internal emitters—Tritium.  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 241
Annex D.	 Biological effects of selected internal emitters—Uranium.  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 361

iii

Report of the United Nations Scientific Committee on the
Effects of Atomic Radiation to the General AssemblyContents
ChapterPage

I.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1

II.

Deliberations of the United Nations Scientific Committee on the Effects of Atomic
Radiation at its sixty-third session . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

2A.

Completed evaluations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

2

B.

Present programme of work . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

3

1.

Developments since the Committee’s 2013 report on the levels and effects ofradiation exposure due to the nuclear accident following the great east-Japan
earthquake and tsunami . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

3Cancer epidemiology of exposures at low dose-rates due to environmental
radiation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

4

3.

Selected evaluations of health effects and risk inference from radiation exposure . .

44.

Collection of data on radiation exposures, in particular on medical andoccupational exposures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

4Outreach activities . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

5

C.

Long-term strategic directions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

6

D.

Future programme of work . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

7E.

Administrative issues . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

8

Scientific report . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

9A.

Methodology for estimating public exposures due to radioactive discharges . . . . . . . . . .

9B.

Radiation exposures from electricity generation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

10C.

Biological effects of selected internal emitters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

13

Members of national delegations attending the fifty-seventh to sixty-third sessions of theUnited Nations Scientific Committee on the Effects of Atomic Radiation . . . . . . . . . . . . . . . . .

16

Scientific staff and consultants cooperating with the United Nations Scientific Committee on
the Effects of Atomic Radiation in the preparation of its scientific report for 2016 . . . . . . . . .

18

2.

5.

III.

Appendices
I.
II.

v

REPORT TO THE GENERAL ASSEMBLY

Chapter I
Introduction
1.
Since the establishment of the United Nations Scientific Committee on the
Effects of Atomic Radiation by the General Assembly in its resolution 913 (X) of
3 December 1955, the mandate of the Committee has been to undertake broad
assessments of the sources of ionizing radiation and its effects on human health andthe environment. 1 In pursuit of its mandate, the Committee thoroughly reviews andevaluates global and regional exposures to radiation. The Committee also evaluates
evidence of radiation-induced health effects in exposed groups and advances in theunderstanding of the biological mechanisms by which radiation-induced effects onhuman health or on non-human biota can occur. Those assessments provide thescientific foundation used, inter alia, by the relevant agencies of the United Nations
system in formulating international standards for the protection of the generalpublic, workers and patients against ionizing radiation; 2 those standards, in turn, arelinked to important legal and regulatory instruments.
2.
Exposure to ionizing radiation arises from naturally occurring sources (such asradiation from outer space and radon gas emanating from rocks in the Earth) andfrom sources with an artificial origin (such as medical diagnostic and therapeutic
procedures; radioactive material resulting from nuclear weapons testing; energy
generation, including by means of nuclear power; unplanned events such as thenuclear power plant accidents at Chernobyl in 1986 and that following the greateast-Japan earthquake and tsunami of March 2011; and workplaces where there may
be increased exposure to radiation from artificial or naturally occurring sources).

__________________
1 The United Nations Scientific Committee on the Effects of Atomic Radiation was established by

the General Assembly at its tenth session, in 1955. Its terms of reference are set out in
resolution 913 (X). The Committee was originally composed of the following Member States:
Argentina, Australia, Belgium, Brazil, Canada, Czechoslovakia (later succeeded by Slovakia),
Egypt, France, India, Japan, Mexico, Sweden, Union of Soviet Socialist Republics (later
succeeded by the Russian Federation), United Kingdom of Great Britain and Northern Ireland,and United States of America. The membership of the Committee was subsequently enlarged by
the Assembly in its resolution 3154 C (XXVIII) of 14 December 1973 to include the Federal
Republic of Germany (later succeeded by Germany), Indonesia, Peru, Poland and the Sudan. By
its resolution 41/62 B of 3 December 1986, the Assembly increased the membership of theCommittee to a maximum of 21 members and invited China to become a member. In its
resolution 66/70 of 9 December 2011, the Assembly further enlarged the membership of theCommittee to 27 and invited Belarus, Finland, Pakistan, the Republic of Korea, Spain andUkraine to become members.
2 For example, the international basic safety standards for radiation protection and safety ofradiation sources, currently co-sponsored by the European Commission, the Food andAgriculture Organization of the United Nations, the International Atomic Energy Agency, theInternational Labour Organization, the Nuclear Energy Agency of the Organization for
Economic Cooperation and Development, the Pan American Health Organization, the United
Nations Environment Programme and the World Health Organization.

1

2

UNSCEAR 2016 REPORT

Chapter II
Deliberations of the United Nations Scientific Committee on
the Effects of Atomic Radiation at its sixty-third session
3.
The Committee held its sixty-third session in Vienna from 27 June to 1 July
2016. 3 Yoshiharu Yonekura (Japan), Chair; John Hunt (Brazil), Peter Jacob(Germany) and Hans Vanmarcke (Belgium), Vice-Chairs; and Michael Waligórski
(Poland), Rapporteur, served as officers of the Committee.
The Committee took note of General Assembly resolution 70/81 on the effects
4.
of atomic radiation. It recalled that it had expected to report its long-term strategic
directions beyond the period covered by its present strategic plan (2014-2019), so as
to help to inform future deliberations of the Assembly on the Committee’s
membership.

A.

Completed evaluations
5.
The Committee discussed four substantive evaluations in detail, adopted thescientific report based on the findings of those evaluations (see chapter III) andrequested that the scientific annexes be published in the usual manner, subject to theagreed modifications.
The Committee had decided, at its fifty-sixth session, to initiate work on a new
6.
estimation of human exposures to ionizing radiation from electricity generation.
Accordingly it had decided to review and update its previous methodology for
estimating public exposures from discharges published in its 2000 report. The
Committee discussed and approved for publication the scientific annex updating themethodology and associated electronic workbooks.
The Committee recalled that progress on the scientific annex on radiation
7.
exposures from electricity generation had been hampered by, among other things,
gaps in data on occupational exposures and on releases associated with electricity
generated from non-nuclear energy sources. In comparison, there were abundantdata for the nuclear energy industry, although those data remained somewhat
deficient as regards decommissioning and other aspects of the so-called back end of
the nuclear fuel cycle. The evaluation has been completed on the basis of reasonable
and transparent assumptions where precise data were not available. The electronic
workbooks for implementing the methodology had been used in 2015 to complete,in an internally consistent manner, the assessment of radiation exposures of
populations from various types of electricity generation.
At its fifty-sixth session, held from 10 to 18 July 2008, during deliberations on
8.
its future programme of work, the Committee had decided that work should beundertaken to address radiation doses and the risks and effects from internallydeposited radionuclides. At its fifty-seventh session, held from 16 to 20 August 2010,__________________
3 The sixty-third session was attended by observers for the International Atomic Energy Agency,

the International Labour Organization, the World Health Organization, the European Union, theInternational Agency for Research on Cancer, the International Commission on RadiologicalProtection and the International Commission on Radiation Units and Measurements.

REPORT TO THE GENERAL ASSEMBLY 3

the Committee had further decided to focus on tritium and radioisotopes of uranium.
At the current session, the Committee agreed that the review of the literature wasnow complete, that the material had been streamlined and its structure harmonized,
and that final conclusions had been drawn from the material evaluated. The
Committee accordingly approved the evaluations for publication.

B.

Present programme of work

1.

Developments since the Committee’s 2013 report on the levels and effects ofradiation exposure due to the nuclear accident following the great east-Japan
earthquake and tsunami
9.
The Committee recalled its assessment of the exposures and effects due to the
nuclear accident after the 2011 great east-Japan earthquake and tsunami, as
presented in its report to the sixty-eighth General Assembly in 2013 (A/68/46) andthe supporting detailed scientific annex. 4 It had concluded in that report that, ingeneral, doses were low and that therefore associated risks were also likely to below. Cancer rates were expected to remain stable. Nevertheless, in the report theCommittee had noted a possibility that the risk of thyroid cancer among thosechildren most exposed to radiation could increase. However, it also noted that thelikelihood of a large number of radiation-induced thyroid cancers in Fukushima
Prefecture — such as after the Chernobyl accident — could be discounted because
absorbed doses to the thyroid after the Fukushima accident were substantially lower.
It had concluded that no discernible changes in the incidence of birth defects andhereditary disease were expected, and that the effects on terrestrial and marine
ecosystems would be transient and localized. Cancer rates for workers were
expected to remain stable.
10. Following its assessment, the Committee put in place arrangements forfollow-up activities to enable it to remain abreast of additional relevant information
as it was published. The Committee’s report of its sixty-second session to theseventieth General Assembly included the findings from the follow-up activities it
had conducted up to that time.
11. The Committee continued to identify further information that had become
available up to the end of 2015, and systematically appraised relevant new
publications to assess their implications for its 2013 report. A notable publication
was the report by the International Atomic Energy Agency (IAEA) on the accident
at the Fukushima Daiichi nuclear power station. 5 It describes the accident and its
causes, evolution and consequences based on an evaluation of data and informationfrom a large number of sources available at the time it was written. That report anda large proportion of the new publications again confirmed the main assumptions
and findings in the Committee’s 2013 report. None of the publications materiallyaffected the main findings in the 2013 report or challenged its major assumptions.
Several publications were identified for which further analysis or more conclusive
evidence from additional research was needed. On the basis of the material
reviewed, the Committee saw no need, at the current time, to make any change to its__________________
4 United Nations publication, Sales No. E.14.IX.1.
5 International Atomic Energy Agency, The Fukushima Daiichi Accident: Report by the Director

General (GC(59)/14), accompanied by technical volumes 1-5.

4 UNSCEAR 2016 REPORT

overarching conclusions. However, several of the research needs identified by theCommittee had yet to be addressed fully by the scientific community.
12. The Committee plans to continue to identify and systematically appraise new
information on the accident, and evaluate the outcomes periodically at its annual
sessions. It also plans to actively engage with those responsible for formulating,
implementing and advising on major research programmes in Japan, in order torapidly assimilate emerging issues, and highlight questions needing further research.
At an appropriate time, dependent on the outcomes, the Committee expects toconsider the need to update its 2013 report.
13. The Committee requested the secretariat, subject to available resources, topublish the findings of its systematic review of new scientific literature as anon-sales publication in English and also to foster its publication in Japanese.
2.

Cancer epidemiology of exposures at low dose-rates due to environmental
radiation
14. The Committee discussed progress on an evaluation of epidemiological studies
of cancer incidence from low-dose-rate exposures due to environmental sources ofradiation. The Committee acknowledged that the scientific review had improved
considerably. It welcomed the development of an appendix on quality criteria for theCommittee’s reviews of epidemiological studies. The Committee requested that thescientific review and quality criteria now be brought into accordance with eachother. It requested that the appendix be finalized for publication as an independent
annex because of its wider application, and expected that both the review andquality criteria could be approved for publication at the sixty-fourth session.

3.

Selected evaluations of health effects and risk inference from radiation exposure
15. The Committee considered progress on evaluations of selected health effects
and the inference of risk from exposure to ionizing radiation. Four scenarios were
proposed for evaluation based on agreed criteria and preliminary literature reviews:
leukaemia after exposure at low dose; solid cancer risk after acute and protracted
exposure; thyroid cancer risk after exposure during childhood or adolescence; andrisk of circulatory diseases after acute and protracted exposure. The Committee
expected that the evaluations would be conducted in line with quality criteria for theCommittee’s reviews of epidemiological studies, and expected to discuss draft
evaluations at the sixty-fourth session.

4.

Collection of data on radiation exposures, in particular on medical andoccupational exposures
16. The Committee took note of a progress report by the secretariat on thecollection, analysis and dissemination of data on radiation exposures, in particular
on medical and occupational exposures. The Committee welcomed the fact that theGeneral Assembly, in its resolution 70/81, had encouraged Member States tonominate a national contact person to facilitate coordination of the collection
and submission of data on the exposure of the public, of workers and of patients.
Fifty-one Member States had nominated national contact persons by the
sixty-third session of the Committee.

REPORT TO THE GENERAL ASSEMBLY 5

17. In 2014, the secretariat had launched an online platform for the collection ofdata on medical exposures and had invited all Member States to take part in theCommittee’s Global Survey of Medical Radiation Usage and Exposures. In preparation
for the Global Survey it had fostered close cooperation with IAEA, theWorld Health Organization and the International Radiation Protection Association.
Twenty countries had submitted their first data on medical exposure; however, not
all their submissions were complete. Because of the relatively low response rate todate, and because of delays brought about by changes in the United Nations
administrative and financial platform (Umoja), the cut-off date for data submission
would be extended until May 2017. The Committee requested the secretariat toprepare a first evaluation of the results for the Committee’s review at its sixty-fourth
session, including a detailed literature review. It also requested the secretariat toaccelerate the survey on occupational exposures, fostering close cooperation with
the International Labour Organization and other relevant bodies, and to begin
detailed work on defining and collecting data on public exposures to radiation fromnatural and artificial sources.
5.

Outreach activities
18. The Committee took note of a progress report by the secretariat on outreachactivities. It acknowledged in particular the work done in Japan to disseminate theCommittee’s 2013 report on the levels and effects of radiation exposure due to the
Fukushima Daiichi accident and the white paper on developments since that report.
It noted that the General Assembly had encouraged the secretariat to continue todisseminate the findings to the public. The Committee also welcomed the outreachactivities surrounding the sixtieth anniversary of the Committee’s inception, thethirtieth anniversary of the Chernobyl accident, and the fifth anniversary of thenuclear accident in Japan. The updated publication of the United Nations
Environment Programme (UNEP) entitled Radiation: Effects and Sources, which isintended as a basic scientific guide for the public, was published in English;
publication in other languages is envisaged. The secretariat had also prepared amemory stick preloaded with all the Committee’s publications and all the
resolutions relevant to its activities, in all official languages of the United Nations
where available, as a handy reference tool.
19. With regard to the sixtieth anniversary of the Committee, the Mayor and
Governor of the City of Vienna hosted a reception for invited dignitaries, scientists
and diplomats at the Vienna Town Hall to commemorate the anniversary. The
Secretary-General of the United Nations, Ban Ki-moon, sent a video message forthe occasion, in which he said: “From assessing the significance of fallout in the
1950s to evaluating the effects of radiation on the human genome today, theCommittee has always taken an independent and impartial approach. This is crucial
on issues that are often highly emotional and political.” Other speakers delivered
messages from the heads of their organizations, including the World HealthOrganization, IAEA, the Preparatory Commission for the Comprehensive
Nuclear-Test-Ban Treaty Organization and UNEP. The messages commended theCommittee for the expertise and independence shown in its scientific reviews,
lauded its efforts to share its scientific findings with a broader audience andencouraged it to further enhance those efforts.

6 UNSCEAR 2016 REPORTC.

Long-term strategic directions
20. The Committee considered its long-term strategic directions beyond the period
covered by its present strategic plan (2014-2019). The Committee took note of thereport of the Secretary-General to the General Assembly at its sixty-ninth session on
the impact of the increase in the membership to 27 States, and possible approaches
to further increase procedures (see A/69/350). The Committee also took note ofGeneral Assembly resolution 70/81 on the effects of atomic radiation, in which theAssembly requested the Secretary-General to provide it at its seventy-second
session with a list of the Member States that had expressed their particular interestin membership in the Committee between the sixty-sixth and seventy-second
sessions.
21. The Committee envisages to direct its future work mainly at the following
scientific areas:(a) Improving the evaluation of exposure levels for people in everyday life,
in occupational environments, during medical procedures, and as a result ofaccidents;(b) Improving the understanding of mechanisms of radiation action andbiological reaction at all levels of biological organization, i.e. from the molecularlevel to the population level;(c) Obtaining more definitive evidence relating to health effects, inparticular health effects from low-dose-range and chronic exposure, and soundestimates of the health implications of exposure of populations to radiation.
22. The Committee also expects that rapidly emerging issues or significant events
may lead to short-term or longer-term reprioritization, and the programme of work
is changed accordingly at each session. As an example, in recent times theCommittee redirected its efforts towards a timely scientific evaluation of the levels
and effects of radiation exposure due to the 2011 nuclear accident in Japan. 4
23. The Committee is of the view that it will be able to continue delivering
authoritative scientific evaluations in the scientific areas outlined above. The
Committee fully supports the Secretary-General’s view that the primary purpose of
any increase in the number of States members should be to enhance the capability of
the Committee to conduct its scientific work. The Committee believes that there is alimit of about 30 States members with which the Committee’s secretariat at its
present size could reasonably cope while at the same time supporting the scientificwork of the Committee. Any increase above that number would require further
strengthening of the secretariat’s human resources (see paragraphs 35 and 40 of
A/69/350).
24. The Committee thus considers that any discussion of membership should focus
on the Committee’s ability to continue its delivery of authoritative scientificevaluations as well as on the secretariat’s ability to support the Committee in doingso. However, in view of the ever-increasing scientific database it may be necessary
to implement a range of strategies that will support the Committee’s efforts to serve
the scientific community as well as wider audiences. Such strategies may also allow
for the inclusion of scientists outside the Committee’s current membership. Therealready are examples of such arrangements, and they have proved beneficial to the

REPORT TO THE GENERAL ASSEMBLY 7

Committee’s work while causing only minor or negligible additions to the workloadof the secretariat.
25. While recognizing the importance of including all States members in, inter alia,
the implementation of the Committee’s strategies, future deliberations and the
production of scientific documents, and while giving due regard to available
resources, the Committee could consider including the following in the strategies
referred to in the previous paragraph:(a) Establishing standing working groups focused on areas such as sourcesand exposure, or health and environmental effects;(b) Inviting, on an ad hoc basis, scientists from other States Members of theUnited Nations to participate in evaluations regarding the above areas;(c) Increasing the Committee’s efforts to present its evaluations, andsummaries thereof, in a manner that attracts readers without compromisingscientific rigour and integrity;(d) While maintaining its lead in providing authoritative scientificevaluations to the General Assembly, liaising closely with other relevantinternational bodies to avoid duplication of efforts to the extent possible.
26. Over the coming sessions, the Committee will work towards implementing theabove strategies.

D.

Future programme of work
27. The Committee discussed preliminary plans for five projects and two smaller
activities. The five topics for which projects were proposed were: (a) second cancers
after radiotherapy; (b) assessment of the impact on biota of radiation exposure dueto the nuclear industry; (c) biological mechanisms that may influence health effects
from low-dose radiation exposure; (d) effects of exposure to radon in homes andworkplaces; and (e) epidemiological studies of radiation and cancer. Having
considered the current work programme and the capacity of both the Committee andits secretariat, the Committee decided:(a) To start projects in 2016 based on topics (c) and (d), and to focus the
project for topic (c) on cancer and hereditary effects;(b) To start a project in 2017 based on the proposal for topic (e) in a version
further elaborated by the delegation of the United States of America;(c) To ask the French delegation to elaborate working material for a morein-depth discussion on the proposal for topic (a) with a view to accepting the
proposal in 2017.
28. The Committee also requested the secretariat to prepare a short paper on thescientific view of the Committee on the dose and dose rate effectiveness factor, andanother on the evaluation of thyroid cancer data in regions affected by the nuclear
power plant accident at Chernobyl in 1986, with a view to discussion andacceptance at the sixty-fourth session of the Committee.

8 UNSCEAR 2016 REPORT

E.

Administrative issues
29. The Committee recognized that, because of the need to maintain the intensityof its work — particularly its work to develop exposure databases and to improvethe dissemination of its findings to the public, including in official languages of theUnited Nations other than English — regular pledges to make voluntary
contributions to the general trust fund established by the Executive Director of
UNEP would be pivotal. The Committee suggested that the General Assembly might
encourage Member States to consider making regular pledges of voluntary
contributions to the general trust fund for that purpose, or to make contributions inkind.
30. The Committee agreed to hold its sixty-fourth session in Vienna from 29 May
to 2 June 2017. It elected new officers to guide the Committee at its sixty-fourth andsixty-fifth sessions: Hans Vanmarcke (Belgium), Chair; Peter Jacob (Germany),
Patsy Thompson (Canada), Michael Waligórski (Poland), Vice-Chairs; and Gillian
Hirth (Australia), Rapporteur.

REPORT TO THE GENERAL ASSEMBLY 9

Chapter III
Scientific report
31. Four scientific annexes (published separately) provide the rationale for thefindings set out below.

A.

Methodology for estimating public exposures due to radioactive
discharges
32. From time to time, the Committee has undertaken estimations of public
exposures from radioactive discharges to the environment under normal operations,primarily from facilities in the nuclear fuel cycle. On each occasion, the Committee
reviewed its methodology for estimating exposures in the light of scientificdevelopments and, where appropriate, revised it. The Committee decided to update
and extend its past evaluations of human exposures to ionizing radiation fromelectricity generation. Consequently, the Committee has reviewed and updated its
previous methodology for estimating public exposures from discharges that had
been published in its 2000 report. Because of the need to be applied more flexibly
for different types of electricity generation and in the interest of transparency, themethodology was updated to provide results in terms of estimated radiation doses
specific to the discharge of each significant radionuclide.
33. The updated methodology can be used to estimate characteristic individual
doses and collective doses resulting from discharges to the atmosphere, rivers andlakes, and the sea. Characteristic individual doses are doses indicative of thosereceived by a typical person living in the area around the discharge point.
A collective dose is the product of the mean dose to a specified population from aparticular source, and the number of people in that population, integrated over adefined period of time. In other words, a collective dose is the dose received by all
members of a particular population combined, over a defined period of time.
However, the calculated doses are metrics to be used only for the comparison ofdifferent sources of exposure, not to estimate implications for health. Moreover, themethodology applies only to routine discharges that can be assumed to becontinuous; more sophisticated methodologies are needed to assess exposures from
accidental releases.
34. Radioactive discharges can lead to exposures of the public in a number ofways, and the updated methodology takes the most important of these into account,namely exposures from radionuclides external to the body, i.e., in the atmosphereand on the ground, and exposures from radionuclides inside the body following
intakes by inhalation and ingestion. To enable the estimation of exposures for both
nuclear and non-nuclear forms of electricity generation, the methodology was
extended to cover a wide range of radionuclides. The methodology uses models
based on experimental data and other field observations in order to estimate thetransfer of radionuclides through the environment and thus the resulting exposuresof the public. The updated methodology now takes into account an additional routeof exposure not previously considered, namely the ingestion of crops irrigated withwater that contains radionuclides as a consequence of discharges to fresh water.

10 UNSCEAR 2016 REPORT

35. In the past, world average values for population densities and foodconsumption were used because those were considered sufficient to estimate global
exposures from nuclear facilities. However, non-nuclear power stations are foundthroughout the world, and population densities and food consumption vary much
more around them. Therefore the Committee has decided to include regional factors.
Even so, the regions now considered are still very large, and other approaches
would be necessary to make assessments for individual sites. Exposures are
estimated using a series of mathematical models, for which the Committee haschosen parameter values that result in realistic exposure estimates. This is incontrast to a more cautious approach often used for regulatory purposes, whereby
values are selected so as to deliberately overestimate exposures.
36. As before, estimates can still be made of collective doses to populations on a
local, regional and global scale, as appropriate. In addition, the methodology
provides information on collective doses resulting from a year’s continuous
discharge into the atmosphere to different population groups as a function of theirdistance from the discharge point. The estimates of collective doses to the world
population are now available integrated over periods of 100, 500 and 10,000 years.
37. The methodology has been implemented in a series of electronic workbooks toprovide transparency and facilitate use and revision by the Committee in any futurestudies. The workbooks contain information on the most important exposure
pathways and radionuclides and are made available for download on the website of
the Committee (www.unscear.org).
38. The Committee is satisfied that the updated methodology, as implemented in
the workbooks, is robust, builds on the strong position of previous versions and issuitable for estimating exposures of regional and global populations from routinedischarges of radionuclides to various environments.

B.

Radiation exposures from electricity generation
39. The world’s mix of electricity-generating technologies changes over time inresponse to the landscape of climatic, environmental, resource, political andeconomic challenges. Governments and researchers may conduct various
comparative studies that take into account, among other things, the impact ofdifferent technologies on the public, on workers and on the environment. Exposure
to ionizing radiation is only one of the many factors that may be taken into accountas part of such assessments. However, the Committee considers that an update andextension of its past evaluations of radiation exposures of the public and of workers
from electricity generation could be a useful source of information for such studies.
40. While interest in the exposure of the public and of workers to radiation due toelectricity generation from nuclear power dates back to the earliest use of thetechnology, radiation exposures from the use of other electricity-generating
technologies have not been so comprehensively studied. The Committee has
periodically reviewed exposures of both the public and of workers related toelectricity generation from nuclear power, and has also conducted evaluations forother forms of electricity generation, albeit to a lesser extent. 6 These evaluations__________________
6 Sources and Effects of Ionizing Radiation — 1977 Report to the General Assembly, with Annexes

REPORT TO THE GENERAL ASSEMBLY 11

have used a variety of methodologies and relied on data from industrial activities
outside the nuclear sector that are not generally monitored or reported in asystematic manner, which has made meaningful comparisons between the radiation
exposures from the different electricity-generating technologies challenging.
41. Assessing the collective dose from accidents was out of the scope of theevaluations of radiation exposures of the public and of workers from electricity
generation; however, the Committee has conducted assessments of past accidents in
its 2008 report; of the Chernobyl accident in its reports of 1988, 2000 and 2008; andof the Fukushima Daiichi nuclear accident in its 2013 report. It is difficult to makedirect comparisons between exposures from accidents and those resulting fromroutine discharges. One of the reasons is that the distribution of doses to the public
immediately after an accident is much more localized geographically, whereas thecollective doses from normal operations for electricity generation are more evenly
distributed over regional or global populations. Nevertheless, the collective dose to
the global population from serious accidents, such as those that occurred at the
Chernobyl and Fukushima Daiichi nuclear power stations, were orders of magnitudelarger than the collective doses to the world population from one year’s normal
operation of significant technologies of electricity generation, as assessed in the
study.
42. As stated above, the Committee has updated its methodology for estimating
public exposures due to radioactive discharges. The methodology is now more
flexible to be able to address a wider range of electricity-generating technologies. In
addition to including an extensive analysis of the available data, the updated
methodology provides the Committee with a sounder basis for comparative studies
than was possible before. In parallel, the Committee also re-evaluated occupational
exposures arising from different electricity-generating technologies, using data
mainly from dosimetry records of worker exposures. These evaluations comprise thebasis for the current comparative study on radiation exposures of both the public
and of workers from electricity generation.
43. The Committee conducted the comparative study by investigating sources of
exposure from electricity-generating technologies based on nuclear power; thecombustion of coal, natural gas, oil and biofuels; and geothermal, wind and solar
power. Two electricity-generating technologies (nuclear power and coal
combustion) were investigated in detail, because a more robust database existed forthem. The Committee evaluated the main sources of radioactive discharges fromtheir life cycle. For the life cycle of nuclear power the sources of radioactive
discharges included uranium mining, milling and mill tailings, power plant
operation and reprocessing activities. For the life cycle associated with the
combustion of coal those sources were coal mining, the operation of coal-fired
power plants (both modern and older-style), and deposits of coal ash. For the sake__________________(United Nations publication, Sales No. E.77.IX.1); Ionizing Radiation: Sources and Biological
Effects — 1982 Report to the General Assembly, with Annexes (United Nations publication,
Sales No. E.82.IX.8); Sources, Effects and Risks of Ionizing Radiation — 1988 Report to theGeneral Assembly, with Annexes (United Nations publication, Sales No. E.88.IX.7); Sources andEffects of Ionizing Radiation — 1993 Report to the General Assembly, with Scientific Annexes(United Nations publication, Sales No. E.94.IX.2); and Sources and Effects of Ionizing
Radiation — 2000 Report to the General Assembly, with Scientific Annexes, Volume I: Sources,(United Nations publication, Sales No. E.00.IX.3).

12 UNSCEAR 2016 REPORTof simplicity we will refer to these cycles as the nuclear fuel cycle and the coal
cycle, respectively.
44. To compare exposures, the Committee focused on two metrics. The firstconsisted of the collective doses to defined population groups resulting from one
year’s global and regional electricity generation by each technology, integrated over
specific time periods. The second metric consisted of the relevant collective doses
divided by the amount of electricity generated by each technology. The reference
year used for the comparisons was 2010.
45. The Committee estimated that the contribution from the coal cycle was morethan half of the total collective dose to the local and regional public from thedischarges due to a single year’s global electricity generation. That estimate wasbased on the assumption that the discharges originated from modern coal plants. The
nuclear fuel cycle, on the other hand, contributed less than a fifth. The contribution
from the coal cycle comes from discharges of natural radionuclides (primarily radonand its radioactive progeny) during coal mining, combustion of coal at power plants
and coal ash deposits. Similarly, almost half of the exposures of the global publicfrom the nuclear fuel cycle result from discharges of natural radionuclides during
uranium mining and milling activities. These values depend on the share of eachtechnology in total electricity production; in 2010 the coal cycle contributed about
40 per cent, the largest amount. Although radon and its progeny are relatively
important contributors to the collective doses to the public for both the nuclear fuelcycle and the coal cycle, the associated individual doses are small compared withdoses due to inhalation of radon and its progeny at levels that occur naturally inhomes.
46. The Committee found, however, that the contribution of a given technology tothe exposures of the global public was not simply a function of how muchelectricity that technology generated. There were also differences in the collective
doses per unit of electricity generated by each technology to take into account. In
normal operations, the coal cycle gave a higher collective dose per unit of electricity
generated than the nuclear cycle, and a significantly higher dose per unit ofelectricity produced than the other technologies evaluated, with the exception ofgeothermal power. Based on the limited information available about radon
discharges from geothermal power plants, the collective dose per unit of electricity
generated by geothermal power could be significant. However, because the use ofgeothermal technology is not widespread, its contribution to radiation exposures of
the global public is smaller than that from the coal cycle.
47. Previous investigations into electricity generation from nuclear power haveexamined the contribution to public exposures made by long-lived radionuclides,
such as carbon-14, which after being discharged circulate globally and continue tocontribute to radiation exposures of the public centuries into the future, albeit atextremely small individual doses. The contribution of the globally circulating
radionuclides to the collective dose to the global public depends on the length oftime for which the collective dose is integrated. Public exposures due to one year’s
discharge of these globally circulating radionuclides continue to increase slowly
over time. Over long integration times, such as hundreds of years, theseradionuclides result in a larger collective dose to the global public from the nuclear
fuel cycle than from the coal cycle.

REPORT TO THE GENERAL ASSEMBLY 13

48. The Committee also assessed occupational exposures. The largest collective
dose to workers per unit of electricity generated resulted from coal mining, becauseof exposures to naturally occurring radionuclides. Of all the collective doses
evaluated, both to the public and to workers, the exposure of workers from coalmining made the largest contribution, although it has fallen over time because of
improving mining conditions. With regard to the construction phase of theelectricity-generating technologies, by far the largest collective dose to workers perunit of electricity generated was found in the solar power cycle, followed by thewind power cycle. The reason for this is that these technologies require large
amounts of rare earth metals, and the mining of low-grade ore exposes workers
to natural radionuclides during mining.
49. The total collective dose per unit of electricity generated in the coal cycle(i.e., the dose to the global public and all exposed workers combined) was larger
than that found in the nuclear fuel cycle. This held true even if long-lived
globally-circulating radionuclides were integrated over 500 years. Whenconsidering the amount of electricity generated in the year 2010 by each technology,
the coal cycle resulted in the largest collective dose to the global public and workers
combined, followed by the nuclear fuel cycle. Of the remaining technologies,
geothermal energy and the combustion of natural gas were the next largest
contributors.
50. Great care should be taken when interpreting and using these results. Their
only function is to provide an insight into the different magnitudes of radiation
exposure resulting from each technology. They are unfit to be used as the only
metric to determine whether one energy generation technology is preferable toanother. As stated earlier, a number of factors determine why countries may select acertain mix of energy generation technologies. Radiation exposure is only one of
them.

C.

Biological effects of selected internal emitters
51. “Internal emitters” is the commonly used term for radionuclides deposited inbody organs and tissues following their intake, principally by inhalation oringestion, but also potentially through wounds or intact skin. Depending on theradionuclide concerned and the physicochemical form of the intake, internal
emitters vary enormously by type, pattern and duration of their radioactive
emissions and energy deposition within and between organs and tissues.
52. It is important to study exposures to internal emitters directly because theradiation from some radionuclides is short-ranged and, to varying degrees, densely
ionizing. Moreover, such radionuclides may be distributed unevenly among bodytissues. Consequently, the nature of the dose delivered by some internal emittersdiffers markedly from that delivered by radiation penetrating from external sources
such as the atomic bombs detonated in Hiroshima and Nagasaki, Japan. Most of theevidence of risk from radiation comes from studies of human exposure topenetrating radiation, while very few direct data are available on the health effects
from internal exposure. Therefore doses to organs from internal emitters have to be
estimated using models, and risk factors are derived principally from studies on
external penetrating radiation. Under those circumstances it is highly desirable to14 UNSCEAR 2016 REPORT

validate the underlying assumptions by obtaining real observations of populations
exposed internally to radiation from specific radionuclides.
53. In response to initiatives in a number of countries to estimate the appropriate
doses from tritium and radioisotopes of uranium and understand the corresponding
health effects, the Scientific Committee has reviewed the relevant information onthese radionuclides. Two scientific annexes provide the rationale for theCommittee’s conclusions set out here.
54. Tritium ( 3 H) is a radioactive isotope of hydrogen that decays solely by
low-energy beta-particle emission. It occurs both naturally, mainly as a result ofinteraction between cosmic-ray particles and the upper atmosphere, and artificially,in the operation of nuclear reactors and other industrial installations, as a substanceused in biomedical research and, in the past, as an ingredient used in a variety of
consumer products. In the future tritium is expected to be used on a large scale infusion reactors. In the environment and at the workplace tritium is encounteredmainly as tritiated water in liquid or vapour form. An aspect of environmental andfood-chain transfer that warrants further investigation is the accumulation of tritiumin the organic component of foodstuffs, referred to as organically-bound tritium.
55. Uranium is a naturally occurring element and is ubiquitously distributed in theenvironment. There are three naturally occurring radioisotopes of uranium: 234 U,
235
U and 238 U. These are present in rocks and soils and hence in the human diet.
They decay mainly by alpha-particle emission and have very long half-lives.
Internal exposures of workers to uranium are mainly the result of mining activities
and its use as a nuclear fuel. In daily life, people are exposed to uranium originating
mainly from drinking water and foodstuffs. Concern has been expressed over
exposures of military personnel and members of the public to depleted uranium(isotope mixtures containing a low percentage of 235 U) used in munitions, forexample by the General Assembly in its resolution 69/57 on the effects of the use ofarmaments and ammunitions containing depleted uranium.
56. Whereas absorbed doses to body organs as a result of exposures to externalsources of radiation are calculated using anatomical models of the human body,commonly referred to as phantoms, estimating doses from internal emitters
additionally requires biokinetic models that describe the behaviour of radionuclides
following their intake into the body, principally by inhalation or ingestion. Suchmodels consider the deposition of inhaled particles and vapour in the respiratory
tract and the passage of ingested radionuclides through the alimentary tract. Models
also represent the subsequent distribution of radionuclides to body organs andtissues from blood, their retention in those sites of deposition, and their excretion.
The reliability of models used to estimate doses from individual elements and theirradioisotopes depends on the quality of available experimental and human data.
57. For tritium, models are available in the form of tritiated water, representing its
distribution throughout body organs and tissues according to their water content.
Less information is available with which to construct adequate models for the
behaviour of various forms of organically bound tritium and other tritiatedcompounds, including amino acids, some of which are involved in the synthesis of
DNA and associated proteins. The absorption of uranium depends partly on whether
it is inhaled or ingested and varies substantially according to the physical andchemical form of the uranium. Uranium absorbed to blood accumulates mainly inREPORT TO THE GENERAL ASSEMBLY 15

the skeleton but with some retention also in the kidneys during the rapid urinary
excretion of a large fraction.
58. Different types of radiation vary in their effectiveness in causing cancer andother health effects. Two broad categories of radiation are photons and charged
particles such as electrons and alpha particles. Some types of charged particles are
generally more effective at causing cancer per unit of absorbed dose thanpenetrating photons. The assessment of such differences relies largely on
experimental data about their relative biological effectiveness (RBE), defined as theratio of the absorbed dose of a reference radiation to the absorbed dose of a test
radiation required to produce the same biological effect.
59. There is extensive literature on studies of RBE for tritium beta particle
emissions. Values of RBE for a range of biological end points range from about
unity to several-fold higher compared to gamma rays and X-rays. However, theability to draw conclusions for carcinogenesis is limited by the very small numberof relevant studies in mammals. Limited information is available that could be used
to estimate RBE values for alpha-particle emissions from isotopes of uranium.
However, RBE values for alpha particles depend on the particles’ energy, range andthe dense deposition of energy along short tracks, and the values of RBE will belargely independent of the radionuclide concerned other than when the radionuclide
determines the origin within body tissues of the emission. Typical values of RBEreported for alpha particles relative to gamma rays or X-rays are around ten for theend points of liver and lung cancer, with lower values for leukaemia.
60. While the tumorigenic effects of uranium in animals are likely related toradiological toxicity due to alpha-particle emissions, some effects are clearly related
to the chemical toxicity of uranium species, particularly in the kidneys. Chemical
toxicity is the limiting factor determining currently acceptable levels of uranium indrinking water.
61. A number of epidemiological studies have been conducted of workers andmembers of the public who may have been exposed to tritium. However, none ofthese studies have so far been informative in showing an increased frequency ofcancer in the exposed populations that could be attributed to radiation exposure
from tritium. Epidemiological studies of nuclear workers have shown a weakassociation between exposures from uranium and rates of lung cancer, but the data
are not sufficiently conclusive to demonstrate a causal relationship.
62. The Committee considered studies on the health effects of depleted uranium
used in munitions for military applications. No clinically significant pathologies
related to exposure from depleted uranium were found in military personnel ormembers of the public. This is consistent with expectations, given the low levels of
measured or assessed exposures.
63. The Committee recognizes that continued research and review is needed toassess the effects of internal exposures. Further work is required to understand theeffects of uneven delivery of doses from internal emitters within tissues and cells
relative to the uniform delivery of doses from external exposure to penetratingradiation. The complexity of changing exposures and tissue sensitivities during inutero and early postnatal development should also be a focus for further research.

16 UNSCEAR 2016 REPORTAppendix I
Members of national delegations attending thefifty-seventh to sixty-third sessions of the United Nations
Scientific Committee on the Effects of Atomic Radiation
Argentina

A. J. González (Representative), A. Canoba, P. Carretto,
M. di Giorgio, M. G. Ermacora

Australia

C.-M. Larsson (Representative), C. Baggoley, M. Grzechnik,
G. Hirth, P. Johnston, S. B. Solomon, R. Tinker

BelarusA. Stazharau (Representative), J. Kenigsberg (Representative),
A. Nikalayenka, A. Rozhko, V. Ternov, N. Vlasova

Belgium

H. Vanmarcke (Representative), S. Baatout, H. Bijwaard,
H. Bosmans, G. Eggermont, H. Engels, F. Jamar, L. Mullenders,
H. Slaper, P. Smeesters, A. Wambersie, P. Willems

BrazilJ. G. Hunt (Representative), D. R. Melo (Representative),
M. Nogueira Martins (Representative), D. de Souza Santos,
L. Holanda Sadler Veiga, M. C. Lourenço, E. Rochedo

CanadaP. Thompson (Representative), N. E. Gentner (Representative),
B. Pieterson (Representative), C. Purvis (Representative),
D. Boreham, K. Bundy, D. B. Chambers, J. Chen, P. Demers,
S. Hamlat, R. Lane, C. Lavoie, E. Waller, D. Whillans

China

Pan Z. (Representative), Chen Y., Dong L., Du Y., Gao H.,
Li F., Lin X., Liu J., Liu S., Liu Y., Pan S., Qin Q., Song G.,
Su X., Sun Q., Wang Y., Xuan Y., Yang H., Yang X., Zhang W.,
Zhou P., Zhu M.

Egypt

W. M. Badawy (Representative), T. S. El-Din Ahmed Ghazey(Representative), M.A.M. Gomaa (Representative), T. Morsi

Finland

S. Salomaa (Representative), A. Auvinen, R. Bly, E. Salminen

FranceL. Lebaron-Jacobs (Representative), A. Rannou (Representative),
E. Ansoborlo, J.-M. Bordy, M. Bourguignon, I. Clairand,
I. Dublineau Naud, A. Flüry-Hérard, J.-R. Jourdain,
R. Maximilien, F. Ménétrier, E. Quémeneur, M. Tirmarche

Germany

P. Jacob (Representative), W. Weiss (Representative),
S. Baechler, A. Böttger, A. A. Friedl, K. Gehrcke, T. Jung,
G. Kirchner, J. Kopp, R. Michel, W.-U. Müller, W. Rühm, H. Zeeb

India

R. A. Badwe (Representative), S. K. Apte (Representative),
K. S. Pradeepkumar (Representative), K. B. Sainis(Representative), B. Das, P. C. Kesavan, Y. S. Mayya

Indonesia

E. Hiswara (Representative), Z. Alatas (Representative),
S. Widodo (Representative), G. B. Prajogi, G. Witono,
B. Zulkarnaen

REPORT TO THE GENERAL ASSEMBLY

Japan

Y. Yonekura (Representative), K. Akahane, M. Akashi, S. Akiba,
T. Aono, N. Ban, M. Chino, H. Fujita, K. Kodama, M. Kowatari,
M. Nakano, O. Niwa, K. Ozasa, S. Saigusa, K. Sakai, G. Suzuki,
M. Takahashi, T. Takahashi, Y. Yamada, H. Yamagishi, H. Yasuda

MexicoJ. Aguirre Gómez (Representative)

PakistanZ. A. Baig (Representative), M. Ali (Representative), R. Ali

Peru

A. Lachos Dávila (Representative), L. V. Pinillos Ashton(Representative), B. M. García Gutiérrez

PolandM. Waligórski (Representative), L. Dobrzyński, M. Janiak,
M. Kruszewski

Republic of
Korea

B. S. Lee (Representative), M. Baek, K.-W. Cho(Representative), K.-H. Do, J.-I. Kim, K. P. Kim, S. H. Kim,
D.-K. Keum, J. K. Lee, J. E. Lee, S. H. Na (Representative),
S. Y. Nam, S. W. Seo

RussianFederation

A. Akleyev (Representative), M. Kiselev (Representative),
R. Alexakhin, T. Azizova, S. Geraskin, V. Ivanov,
N. Koshurnikova, A. Koterov, A. Kryshev, I. Kryshev, B. Lobach,S. Mikheenko, O. Pavlovsky, A. Rachkov, S. Romanov,
A. Samoylov, A. Sazhin, S. Shinkarev

Slovakia

L. Auxtová (Representative), E. Bédi (Representative),
M. Zemanová (Representative), M. Chorváth, A. Ďurecová,
V. Jurina, Ž. Kantová, K. Petrová, L. Tomášek, I. ZachariášováSpain

M. J. Muñoz González (Representative), D. Cancio, M. T. Macías
Domínguez, J. C. Mora Cañadas, B. Robles Atienza, E. Vañó
Carruana

Sudan

N. A. Ahmed (Representative), I. Salih Mohamed Musa
(Representative), E.A.E. Ali (Representative), A. E. Elgaylani(Representative), M.A.H. Eltayeb (Representative),
I. I. Suliman

SwedenI. Lund (Representative), L. Hubbard (Representative),
L. Moberg (Representative), A. Almén, E. Forssell-Aronsson,
L. Gedda, J. Johansson Barck-Holst, J. Lillhök, A. Wojcik

UkraineD. Bazyka (Representative)

United Kingdomof Great Britain
and Northern
Ireland

S. Bouffler (Representative), J. Cooper (Representative),
J. Harrison (Representative), A. Bexon, J. Simmonds,
R. Wakeford, W. Zhang

United States ofAmerica

R. J. Preston (Representative), F. A. Mettler Jr. (Representative),
A. Ansari, L. R. Anspaugh, J. D. Boice Jr., W. Bolch, H. Grogan,
N. H. Harley, E. V. Holahan Jr., B. A. Napier, D. Pawel,
G. E. Woloschak

17

18

UNSCEAR 2016 REPORTAppendix II
Scientific staff and consultants cooperating with the United
Nations Scientific Committee on the Effects of Atomic
Radiation in the preparation of its scientific report for 2016
L. Anspaugh
B. Lauritzen
M. Balonov
I. Dublineau
H. Grogan
L. Hubbard
B. Lambert
C. Robinson
E. Rochedo
R. Shore
J. Simmonds
R. Wakeford
Secretariat of the United Nations Scientific Committee on the Effects of Atomic
Radiation
M. J. Crick
F. Shannoun

ANNEX A
METHODOLOGY FOR ESTIMATING PUBLICEXPOSURES DUE TO RADIOACTIVEDISCHARGES

19

Corrigendum to Sales No. E.17.IX.1

15 December 2020

Sources, Effects and Risks of Ionizing Radiation: United
Nations Scientific Committee on the Effects of Atomic
Radiation 2016 Report to the General Assembly, with
Scientific Annexes
Corrigendum
1.

Annex A (Methodology for estimating public exposures due to radioactive
discharges), page 81, table 25
Column headed “Asia and Pacific”
In the entry for radionuclide 54Mn, for 5.3 × 10−9 read 2.7 × 10−9
In the entry for radionuclide 58Co, for 8.6 × 10−10 read 4.6 × 10−10
In the entry for radionuclide 60Co, for 2.6 × 10−9 read 1.5 × 10−9
In the entry for radionuclide 65Zn, for 3.7 × 10−9 read 3.6 × 10−9
In the entry for radionuclide 106Rua, for 6.0 × 10−11 read 3.7 × 10−11
In the entry for radionuclide 131I, for 8.8 × 10−14 read 7.1 × 10−14
In the entry for radionuclide 134Cs, for 3.1 × 10−11 read 2.1 × 10−11
In the entry for radionuclide 137Csa, for 1.5 × 10−11 read 1.2 × 10−11
In the entry for radionuclide 232Tha, for 1.6 × 10−8 read 1.1 × 10−8
In the entry for radionuclide 238Ua, for 3.3 × 10−11 read 3.2 × 10−11
In the entry for radionuclide 241Am, for 1.2 × 10−9 read 1.1 × 10−9
Column headed “Europe”
In the entry for radionuclide 54Mn, for 5.2 × 10−9 read 2.6 × 10−9
In the entry for radionuclide 58Co, for 8.3 × 10−10 read 4.3 × 10−10
In the entry for radionuclide 60Co, for 2.5 × 10−9 read 1.3 × 10−9
In the entry for radionuclide 65Zn, for 2.1 × 10−9 read 2.0 × 10−9
In the entry for radionuclide 106Rua, for 5.2 × 10−11 read 2.9 × 10−11
In the entry for radionuclide 131I, for 8.5 × 10−14 read 6.8 × 10−14
In the entry for radionuclide 134Cs, for 3.2 × 10−11 read 2.3 × 10−11
In the entry for radionuclide 137Csa, for 1.6 × 10−11 read 1.3 × 10−11

V.20-07528 (E)*2007528*

Corrigendum to Sales No. E.17.IX.1

In the entry for radionuclide 232Tha, for 1.5 × 10−8 read 9.1 × 10−9
In the entry for radionuclide 241Am, for 6.8 × 10−10 read 6.0 × 10−10
Column headed “Latin America”
In the entry for radionuclide 54Mn, for 5.2 × 10−9 read 2.6 × 10−9
In the entry for radionuclide 58Co, for 8.1 × 10−10 read 4.1 × 10−10
In the entry for radionuclide 60Co, for 2.4 × 10−9 read 1.2 × 10−9
In the entry for radionuclide 65Zn, for 1.4 × 10−9 read 1.3 × 10−9
In the entry for radionuclide 90Sra, for 5.1 × 10−13 read 5.0 × 10−13
In the entry for radionuclide 106Rua, for 4.8 × 10−11 read 2.5 × 10−11
In the entry for radionuclide 131I, for 5.7 × 10−14 read 4.0 × 10−14
In the entry for radionuclide 134Cs, for 2.6 × 10−11 read 1.6 × 10−11
In the entry for radionuclide 137Csa, for 1.2 × 10−11 read 7.9 × 10−12
In the entry for radionuclide 232Tha, for 1.3 × 10−8 read 7.4 × 10−9
In the entry for radionuclide 241Am, for 4.0 × 10−10 read 3.3 × 10−10
Column headed “North America”
In the entry for radionuclide 54Mn, for 5.3 × 10−9 read 2.7 × 10−9
In the entry for radionuclide 58Co, for 8.5 × 10−10 read 4.5 × 10−10
In the entry for radionuclide 60Co, for 2.6 × 10−9 read 1.4 × 10−9
In the entry for radionuclide 65Zn, for 5.6 × 10−9 read 5.5 × 10−9
In the entry for radionuclide 106Rua, for 5.6 × 10−11 read 3.3 × 10−11
In the entry for radionuclide 131I, for 8.3 × 10−14 read 6.6 × 10−14
In the entry for radionuclide 134Cs, for 3.2 × 10−11 read 2.2 × 10−11
In the entry for radionuclide 137Csa, for 1.6 × 10−11 read 1.2 × 10−11
In the entry for radionuclide 232Tha, for 1.7 × 10−8 read 1.2 × 10−8
In the entry for radionuclide 241Am, for 1.0 × 10−9 read 9.2 × 10−10
Column headed “West Asia”
In the entry for radionuclide 54Mn, for 5.2 × 10−9 read 2.6 × 10−9
In the entry for radionuclide 58Co, for 8.0 × 10−10 read 4.0 × 10−10
In the entry for radionuclide 60Co, for 2.3 × 10−9 read 1.2 × 10−9
In the entry for radionuclide 65Zn, for 5.9 × 10−10 read 5.3 × 10−10
In the entry for radionuclide 90Sra, for 2.2 × 10−13 read 2.1 × 10−13
In the entry for radionuclide 106Rua, for 4.6 × 10−11 read 2.3 × 10−11
In the entry for radionuclide 131I, for 4.7 × 10−14 read 3.0 × 10−14
In the entry for radionuclide 134Cs, for 2.4 × 10−11 read 1.4 × 10−11
In the entry for radionuclide 137Csa, for 1.0 × 10−11 read 6.3 × 10−12
In the entry for radionuclide 232Tha, for 1.2 × 10−8 read 6.3 × 10−9
In the entry for radionuclide 238Ua, for 5.8 × 10−12 read 5.6 × 10−12
In the entry for radionuclide 241Am, for 2.3 × 10−10 read 1.5 × 10−10

2/3

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Corrigendum to Sales No. E.17.IX.1

2.

Annex A (Methodology for estimating public exposures due to radioactive
discharges), page 99, table A2
For the existing equation T2 substitute
𝐶𝐶𝑎𝑎𝑎𝑎𝑎𝑎,𝑝𝑝𝑝𝑝𝑝𝑝𝑝𝑝𝑝𝑝𝑝𝑝𝑝𝑝 (𝑥𝑥) = 𝐶𝐶𝑎𝑎𝑎𝑎𝑎𝑎,𝑝𝑝𝑝𝑝𝑝𝑝𝑝𝑝𝑝𝑝𝑝𝑝 (𝑥𝑥)V.20-07528

𝜆𝜆progeny

𝜆𝜆progeny −𝜆𝜆parent

�1 − 𝑒𝑒−�𝜆𝜆progeny −𝜆𝜆parent ��

𝑥𝑥
�
𝑢𝑢𝑎𝑎

�

(T2)

3/3

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 21

Contents
I.

INTRODUCTION ................................................................................................................................... 23

II. GENERAL CONSIDERATIONS .......................................................................................................... 24
A. Dosimetric quantities ................................................................................................................ 24
B. Integration times ........................................................................................................................ 29
C. End points considered .............................................................................................................. 30
D. Discharges considered.............................................................................................................. 31
III. DESCRIPTION OF THE METHODOLOGY ...................................................................................... 32
A. Assessment of doses from discharges to atmosphere.................................................. 34
B. Assessment of doses from discharges to a freshwater environment ................................. 48
C. Assessment of doses from discharges to a marine environment ............................. 64
IV. GLOBALLY DISPERSED RADIONUCLIDES ................................................................................... 70
A. Tritium ............................................................................................................................................ 70
B. Carbon-14 ...................................................................................................................................... 71
C. Iodine-129 ..................................................................................................................................... 72
D. Krypton-85 .................................................................................................................................... 72
V. LIMITATIONS OF THE MODELS AND DATA USED ................................................................... 73
A. Discharges to atmosphere ...................................................................................................... 74
B. Discharges to freshwater and marine environments .................................................... 75
C. Global modelling ........................................................................................................................ 76
D. Comparison of the current and previous methodologies ........................................... 77
VI. APPLICATION OF THE METHODOLOGY ...................................................................................... 78
VII. CONCLUSIONS ..................................................................................................................................... 90
VIII. ACKNOWLEDGEMENTS .................................................................................................................... 91
TABLES ........................................................................................................................................................... 97
REFERENCES ............................................................................................................................................... 127
Electronic attachments (http://www.unscear.org/unscear/en/publications/2016.html)
1. Comparison of the Committee’s earlier and current methodologies
2. Summary of the Committee’s pre-2010 methodologies and their development
3. Population densities around uranium mines and nuclear power sites
4. Sensitivity studies in support of the methodology development
5. Electronic workbooks that implement the current methodology

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 23

I. INTRODUCTION
1.
For many years the Committee has had a methodology for assessing the radiation exposures of thegeneral public from discharges of radionuclides to the environment, where the term discharge refers toauthorized releases from normal operations. The methodology does not apply for accidental releases.
Every few years, the Committee has updated its estimates of the global radiological impact of thevarious types of nuclear installation, using this methodology and the latest available information on thelevel and nature of discharges. The most recent versions of the methodology, which in turn were built
on earlier versions, were published in annex A of the UNSCEAR 2000 Report [U6] and annex B of the
UNSCEAR 2008 Report [U10]. Although used successfully for many years, the Committee decided at
its fifty-sixth session in 2008 to review and, where appropriate, update its methodology [U10] as part ofits strategic plan for 2009–2013, with a view to subsequently updating its assessments of the levels ofradiation exposure from energy production.
2.
The global impact of energy production was one of the Committee’s thematic priorities identified
in its strategic plan for 2014−2019 (see [U12]). The Committee decided to specifically assess the global
radiological impact of discharges from both nuclear and non-nuclear sources of electrical energy
production, and thus to review its methodology so that it could be used to:(a) Evaluate the radiation levels worldwide to which people are usually exposed as a consequenceof electricity generation;(b) Assess the typical variations in exposure worldwide to different sources of radioactive
discharges;(c) Identify sources of possible concern for public exposure;(d) Allow users to derive benchmarks for comparison purposes and to derive relationships fortheir investigative work;(e) Analyse temporal trends in the contributions of different sources to overall public exposure.
3.
The Committee recognizes that it is impractical to carry out a full site-specific assessment foreach site of electricity generation in the world that discharges radionuclides to the environment.
However, it decided that, where possible, its approach should be able to take into account differencesamong regions of the world. It also decided that the methodology should be robust, transparent andapplicable to the different electrical energy sources, and that it should build on the experience gained inapplying the previous methodology. One change agreed was that for ease of application to different
electrical energy sources and to all parts of the nuclear fuel cycle, the methodology should provide
estimates of the doses 1 per unit discharge of key radionuclides to different environments. With
information on the discharges from the various facilities, these factors could then be used to estimate
doses per unit electrical energy produced (see annex B).
1

Where the word “dose” is used in this annex without qualification, it should be taken to mean effective dose to an individual orcollective effective dose to a population, according to context, unless otherwise stated.

24 UNSCEAR 2016 REPORT

4.
The methodology is thus designed to assess individual and collective doses from unit discharge ofeach of the key radionuclides to atmosphere, lakes, rivers and seas. The individuals considered arethose living in the area local to the point of discharge with behaviour indicative of most people living inthat area; the doses to these characteristic individuals are referred to here as “characteristic individual
doses” and are not to be confused with those to the so-called “representative person” (previously
termed by the International Commission on Radiological Protection, ICRP, as the “critical group”) used
for radiation protection purposes [I13]. The Committee agreed that, where possible, estimates should beavailable for collective doses as a function of distance from the point of discharge and of the levels of
individual dose that make up these collective doses. The Committee also agreed to take account ofsome differences among geographical regions, notably on population densities. For ease of applicationand transparency, the methodology has been implemented through a series of Excel® workbooks, whichshould also facilitate any future updates of the Committee’s assessments.
5.
The Committee intends to use the methodology to conduct its worldwide assessments of theradiological impact of discharges of radionuclides. The methodology aims to provide best estimates ofradiation doses, in contrast to other established methodologies (e.g. [I2]) which aim to ensure that doses
are not underestimated and thus adopt cautious parameter values and assumptions. The Committee’s
methodology only applies to continuous routine releases to the environment and is not intended to becomprehensive, covering all situations and other uses (such as for estimating the risks of radiationinduced health effects).
6.
This scientific annex describes the methodology to assess doses from discharges to atmosphere,
freshwater bodies and sea, together with the approach adopted for globally dispersed radionuclides. The
limitations of the models and data used are also discussed. An appendix provides detailed informationon the implementation of the methodology, including the end points considered, the relevant
mathematical equations and model parameters, a specification of the series of Excel® workbooks that
implement the methodology, information on how the workbooks can be used to assess exposures fromdifferent types of discharge, and a description of how progeny of radionuclides were treated and quality
assurance conducted. In addition more detailed supporting information and the workbooks themselves
are provided in electronic attachments, which can be downloaded only from the UNSCEAR website(http://www.unscear.org/unscear/en/publications/2016.html).

II. GENERAL CONSIDERATIONS

A. Dosimetric quantities
7.
Although for strict scientific purposes other quantities are appropriate, the Committee has formany years used two protection quantities, namely effective dose and collective effective dose for its
evaluations of levels and trends of exposures. These have the advantage of simplifying comparison ofdoses from different types of radiation and different distributions of dose within the body, and of
averaging over age and sex; moreover, many regulatory authorities keep records in terms of thesequantities. The quantity effective dose is based on another underlying radiation protection quantity, theequivalent dose, which takes into account the inferred differences in biological effectiveness ofdifferent types of radiation by applying defined radiation weighting factors to the absorbed dose to an

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 25

organ or tissue. Equivalent doses to different organs and tissues are then combined using defined tissue
weighting factors, which take into account the inferred differences in detriment from irradiation of theparticular organ or tissue, averaged over the sex and age distributions of typical populations. This
enables external and internal exposures from a source of radiation to be added to produce an overall
effective dose 2 from that source, which can then be compared with effective doses from other sources.
In this annex, the definition of effective dose used to express both individual and collective doses isgenerally that of Publication 60 of the ICRP [I10]. Although ICRP made some modifications to theweighting factors in the definition of effective doses in its 2007 recommendations [I13], it has not yet
published a full set of dose coefficients for external and internal exposure and so the Committee has
used the previous values, unless otherwise stated.
8.
It is important to note that there are differences between the quantities of individual and collective
dose, although they are calculated in similar ways. In this methodology and any related studies, thecollective (effective) dose is always estimated for a defined population over a specified period of time.
The collective dose used here is the product of the mean effective dose to a specified population from aparticular source, and the number of people in that population, integrated over a defined period of time.
In other words, a collective dose is the dose received by all members of a particular population
combined, over a defined period of time. When evaluating collective doses, the population and time
period should always be specified and any underlying assumptions—such as the population remaining
the same over the time period—should be acknowledged. The Committee has used the quantity,collective dose, for many years to compare the radiation exposures of populations from different
sources of ionizing radiation, or following different protection measures. In particular, the quantity
provides a convenient basis for comparing the impacts of radioactive discharges from nuclear and othersources of electrical energy, and for incorporation by analysts into measures of economic utility fordecision-aiding. However, collective dose should only be used for comparative purposes; it isinappropriate to use it for quantifying exposure in epidemiological studies or for making risk
projections. Specifically, the Committee has stated that it does not recommend the aggregation of very
low doses over extended time periods to large numbers of individuals to estimate absolute numbers ofradiation-induced health effects within a population exposed to incremental doses at levels equivalent
to or lower than normal natural background levels [U13].
9.
The Committee decided to retain its previous approach where collective dose was estimated for
local, regional and global components. The spatial variation in collective dose can be estimated fordischarges to atmosphere because (a) dispersion can be expressed as a function of distance from thedischarge point and (b) data on the distribution of population density are also available as a function of
distance. However, for aquatic discharges, when using a simple generic approach, it is not possible todetermine collective doses as a function of distance. Because people over a wide area may all obtain
their drinking water and fish from the same water body, there is not a simple relationship between
distance from the discharge point and dose. A simplified approach is therefore used based on adischarge to a single water body so that, for aquatic discharges, less spatial resolution in collective
doses is possible than for discharges to atmosphere.
10. The Committee considered that the assessment of characteristic individual doses is also useful inaddressing the thematic priority in its strategic plan mentioned above. As noted previously, as far as
possible, the dose estimates are based on best estimates of parameter values. For example, fordischarges to atmosphere, the characteristic individual dose estimated is the dose to someone living
2
For both individual and collective doses, the effective doses estimated here are the sum of the effective doses from external
exposure received during the period of interest and the committed effective doses from intakes by inhalation and ingestion during
the same period.

26 UNSCEAR 2016 REPORT

5 km from the discharge point who is assumed to obtain a proportion of their food from this location,rather than a more cautious assumption of living 1 km or less from the discharge point. For aquatic
discharges, the characteristic individual dose estimated is the dose to someone assumed to be living in
the area near to the discharge location and to obtain part of their food—and in the case of discharges tofreshwater bodies, their drinking water—from the receiving water body.
11. In this regard, for estimating characteristic individual doses, the Committee considered it to be
more realistic to assume that only a proportion of their food consumed was locally produced (locally
produced food is that grown by the individual or bought fresh from local markets) rather than an
approach often used for regulatory purposes of assuming all their food consumed was 100% locally
produced. There is little information on how much locally produced food people consume. In France, acomprehensive investigation was carried out to determine the amount of locally produced food thatpeople ate [B3]. This showed that rural populations in France consumed more locally produced foodthan urban populations, as would be expected, and that there was a significant variation for different
foods. For example, for milk consumption in France as a whole, only 3% was found to be locally
produced but in rural areas, the figure was 30%, on average. For leafy vegetable consumption in France
as a whole, just over 25% was found to be locally produced but in rural areas the figure was 70%. For
cereal consumption in France, the percentage locally produced was very low, at 0.1% both for thecountry as a whole and for rural areas. While recognizing that there are likely to be wide variations inpractice across the world, the Committee agreed that a simple assumption of 25% of the food consumed
being locally produced was reasonable for its generic methodology and that this value be used for bothterrestrial and aquatic foods. This value is consistent with the value used in the Committee’s recent
assessment of doses in Japan following the nuclear accident at the Fukushima-Daiichi nuclear power
station [U12].
12. The methodology employs published dose coefficients to estimate doses from external andinternal exposure. The dose coefficients used for internal exposure are those for adult members of the
public and are the committed effective doses to 70 years of age per unit intake of radionuclide given by
ICRP [I14]. The dose coefficients for external exposure from radionuclides in the air or deposited on
the ground following discharges to air or water bodies were selected from literature sources relevant tothe situation being considered. The main dose coefficients used for internal and external exposure aresummarized in table 1 both for radionuclides that are discharged and for their key progeny, whereappropriate (see also the appendix). It should be noted that for dose coefficients for internal exposure
the contribution of short-lived progeny is included in the dose coefficient assigned to the parent.
However, for the dose coefficients for external exposure the dose coefficients are for the parent alone
and any contribution from short-lived progeny has to be added separately using the values given intable 1(b). Except for isotopes of argon (e.g. 41Ar), krypton (e.g. 85Kr), xenon (e.g. 133Xe, 135Xe and138Xe) and radon (e.g. 222Rn), which are all gases, the discharges are assumed to be particulate with adefault size of 1 µm AMAD. The dose coefficients for tritium in the form of tritiated water (HTO) andof organically bound tritium (OBT) are provided for use with the “specific activity” model for tritium inthe environment described in chapter III. The chemical form assumed for each element is generally thedefault type for public exposure given in ICRP Publication 72 [I12].

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 27

Table 1. Radionuclide-specific dose coefficients for internal exposure from inhalation andingestion and for external exposure from the plume and deposition
Values are given to one decimal place only as this is considered sufficient for this purpose(a) Radionuclides discharged
Radionuclide

ProgenyaLungabsorption
typeb
[I12]

3

Committed effective dosecoefficients for adults to age
70 years for internal exposurec

Effective dose coefficients forexternal exposured

Inhalation
Dinh(Sv/Bq)[I14]

Ingestion
Ding(Sv/Bq)
[I14]

Plume
Dex,cloud(Sv/(Bq s/m3))[E2, P2](Sv/(Bq /m2))e
[P2]

0

0

Deposition
Dex,deposit

H

—

M

4.5 × 10−11

0

C

—

M

2.0 × 10

5.8 × 10

2.6 × 100M

1.4 × 10

7.7 × 10−18

3.1 × 10

0

14

S

35 f

−9
−9

−10
−10

−18

41

Arf

—

n/a

0

0

6.2 × 10−14

0

54

Mn

—

M

1.5 × 10−9

7.1 × 10−10

3.9 × 10−14

5.2 × 10−14

58

Co

—

M

1.6 × 10−9

7.4 × 10−10

4.5 × 10−14

5.0 × 10−9

60

Co

—

M

1.0 × 10−8

3.4 × 10−9

1.2 × 10−13

2.0 × 10−7

65

Zn

—

M

1.6 × 10−9

3.9 × 10−9

2.8 × 10−14

9.0 × 10−9

85

Krf

—

n/a

0

0

9.9 × 10−17

0

90

Sr

90

Y

M

3.6 × 10

2.8 × 10

9.8 × 100RhM

2.8 × 10

7.0 × 10

0

0

c

−8

−8

−17

106

Ru

129

I

—

F

3.6 × 10−8

1.1 × 10−7

2.8 × 10−16

3.1 × 10−9

I

—

F

7.4 × 10−9

2.2 × 10−8

1.7 × 10−14

2.4 × 10−10

133

Xef

—

n/a

0

0

1.2 × 10−15

0

135

Xe

f

135

Cs

n/a

0

0

−14

1.0 × 10

0

138

Xef

138

Cs

n/a

0

0

5.4 × 10−14

0

134

Cs

—

F

6.6 × 10−9

1.9 × 10−8

7.1 × 10−14

6.2 × 10−8

137

Csc

137m

F

4.6 × 10−9

1.3 × 10−8

9.3 × 10−17

4.6 × 10−13

210

Pb

M

1.1 × 10−6

6.9 × 10−7

4.5 × 10−17

2.5 × 10−10

210

Po

—

M

3.3 × 10−6

1.2 × 10−6

3.9 × 10−19

1.2 × 10−13

222

Rnf,g

Po 214Pb, 214Bi, 214Po

n/a

4.8 × 10−9

n/a

n/a

n/a

226

Ra

—

M

3.5 × 10

2.8 × 10

2.8 × 10

3.8 × 10−9

230

Th

—

S

1.4 × 10−5

2.1 × 10−7

1.5 × 10−17

0

232

Th

Ra, 228Ac, 228Th, 212Pb

S

2.5 × 10−5

2.3 × 10−7

7.2 × 10−18

8.7 × 10−11

234

U

—

M

3.5 × 10−6

4.9 × 10−8

6.1 × 10−18

5.5 × 10−11

238

UTh, 234mPa

M

2.9 × 10−6

4.5 × 10−8

2.5 × 10−18

1.9 × 10−11

239

Pu

—

M

5.0 × 10−5

2.5 × 10−7

3.5 × 10−18

3.5 × 10−11

240

Pu

—

M

5.0 × 10−5

2.5 × 10−7

3.4 × 10−18

1.5 × 10−11

241

Am

—

M

4.2 × 10−5

2.0 × 10−7

6.7 × 10−16

7.9 × 10−9

c

106

131 d

Ba

218

228

234

−8

−6

−9

−7

−16

28 UNSCEAR 2016 REPORT(b) Radionuclides considered to have different chemical forms in the environment or as progeny ofothersRadionuclide

Lung absorption typeb[I12]

Committed effective dosecoefficients for adults to age 70
years for internal exposurec

Effective dose coefficients forexternal exposured

Inhalation
Dinh(Sv/Bq)[I14]

Ingestion
Ding(Sv/Bq)
[I14]

Plume
Dex,cloud(Sv/(Bq s/m3))[E2, P2]

Deposition
Dex,deposit(Sv/(Bq /m2))e
[P2]
0

HTO

n/a

1.8 × 10−11

1.8 × 10−11

0

OBT

n/a

4.1 × 10

4.2 × 10

0

Y

—

—

—

7.9 × 10

0

106

Rh

—

—

—

9.4 × 10−15

5.5 × 10−15

135

Csf

F

6.9 × 10−10

2.0 × 10−9

9.5 × 10−18

2.4 × 10−11

138

Csf

F

2.4 × 10−11

9.2 × 10−11

1.2 × 10−13

3.9 × 10−12

Ba

—

—

—

2.7 × 10−14

1.7 × 10−7

212

Pb

—

—

—

5.7 × 10−15

4.7 × 10−12

214

Pb

M

1.4 × 10−8

1.4 × 10−10

1.1 × 10−14

5.6 × 10−13

210

Bi

—

—

—

2.6 × 10−16

2.2 × 10−11

214

Bi

—

1.4 × 10−8

1.1 × 10−10

7.1 × 10−14

2.5 × 10−12

214

Po

n/a

0

0

3.8 × 10−18

1.9 × 10−23

218

Po

n/a

0

0

4.2 × 10−19

2.3 × 10−18

228

Ra

M

2.6 × 10−6

6.9 × 10−7

0

0

228

Ac

F

2.5 × 10

228

Th

S

234

Th
Pa

90

137m

234m

−11

−11

0
−16

4.3 × 10−14

4.0 × 10

1.7 × 10−11

4.0 × 10−5

7.2 × 10−8

8.1 × 10−17

6.9 × 10−11S

7.7 × 10−9

3.4 × 10−9

2.9 × 10−16

1.3 × 10−11

n/a

0

0

1.2 × 10−15

1.1 × 10−14

−8

−10

a

The progeny considered in the methodology are listed. For further details see the appendix; The symbol “—” means either thatthe progeny are not considered or that the progeny are stable isotopes.

b

This refers to the rate of absorption from the lung. The symbols relate to fast (F), medium (M) or slow (S) absorption.

c

For these radionuclides with short-lived progeny, the dose coefficients for internal exposure for the parent include thecontribution from the progeny and coefficients are not given separately for the progeny.
dIn all cases the dose coefficients for external exposure are for the parent only and the contribution from progeny has to beconsidered separately.

eIntegrated effective dose to 100 years following a continuous deposition rate of 1 Bq/(m2 s) for one year. The Petoussi-Henss etal. values are not used directly but are modified to allow for changes in external dose with time [G2].
fConsidered for discharges to atmosphere only.

g

See section III.A.3 for more information on the treatment of radon inhalation.

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 29

B. Integration times
13. The methodology is applicable to discharges that can be assumed to be continuous and takes into
account (a) the build-up of long-lived radionuclides in the environment and (b) the continued exposure
to long-lived radionuclides after discharges have ceased. This is done by considering a year’s dischargeof a radionuclide, its dispersion in the environment and the subsequent exposures of people over many
years; the resulting dose rates are then integrated to various times. Using these integrals, it is also
possible to consider that the discharges may continue for many years from the same site. This isbecause the integrated dose to say 100 years from one year’s discharge is numerically equal to the dosein the 100th year from a continuous discharge at a constant rate. The Committee agreed that 100 years
was generally a reasonable assumption for the length of continuous discharge for estimating
characteristic individual doses as it covers the lifetime of individual power plants and also thepossibility that additional power plants may be constructed on the same site. Separate considerations
apply for the disposal of waste residues (mill tailings) from uranium and coal mining operations as
discussed below.
14. Collective dose rates can be integrated to infinity (referred to as the collective dose commitment) to
take account of all possible future doses from a discharge. However collective dose rates are generally
integrated to shorter times; values ranging from 80 to 10,000 years have been used in the Committee’s
past reports, see electronic attachments 1 and 2 (collective doses integrated to various times are sometimes
referred to as truncated collective doses). For most radionuclides, integrating collective dose rates to
100 years gives a significant proportion of the collective dose commitment [S9]. However, some relatively
long-lived, environmentally mobile radionuclides (3H, 14C, 85Kr and 129I) become globally dispersed andcan continue to contribute to the integrated collective doses for many years. For both 3H and 85Kr (halflives 12.6 years and 10.7 years, respectively), the integrated collective dose does not increase beyond
100 years. However, for 14C (half-life 5,730 years), the integrated collective dose increases by a factor of
more than five between 1,000 and 10,000 years; however the additional collective dose is then negligible
if the integration is continued to 1,000,000 years. For the very long-lived 129I (half-life 15.7 million years),
the integrated collective dose increases by about a factor of six between 1,000 and 10,000 years and by afactor of over 150 between 10,000 and 1,000,000 years (based on table 70 of [U5]). However, fordischarges from nuclear sites, the integrated collective dose from 14C rather than that from 129I dominates
the total integrated collective dose and so most of the collective dose will be received in the first
10,000 years [S9, U5]. Also, when comparing options explicit consideration of collective doses at longtimes is often not necessary because the collective doses at these times are all similar and moreovercomprise extremely low levels of individual dose [S9, U5].
15. In general, the uncertainties associated with estimating collective doses increase with integration
time, for example, because of possible major changes in environmental conditions and population
dynamics. The integrated collective dose to 500 years from a continuous discharge over one year hasoften been used to provide an upper estimate of the highest future annual collective dose rate if thepractice, and hence discharge, were to continue at a constant rate for that number of years [L1].
Five hundred years is used because it is assumed to be the maximum duration of the generation ofelectrical energy by nuclear power. An integration time of 10,000 years was used in the Committee’s
assessments conducted before the year 2000. Ten thousand years is consistent with the duration of thewarm period between glacial periods [L1] and therefore the maximum number of years over which any
reasonable assumptions about the nature of the global environment can be made.
16. When comparing the radiological impact of different sources, it is important not to introduce any
bias when choosing integration times for collective dose. For example, choosing a short time (e.g. of afew hundred years) for nuclear power generation would mean that a significant portion of the integrated

30 UNSCEAR 2016 REPORT

collective dose is not included. For non-nuclear energy production, notably that using fossil fuels, themajor radionuclides discharged are from the 238U decay chain, particularly 226Ra and its progeny
including 222Rn. These radionuclides do not become dispersed globally and so once discharged to theenvironment need to be considered only for relatively short timescales; a few hundred years is sufficient
to allow for the ingrowth of the radiologically significant radionuclide, 210Po.
17. The conventional mining and milling of uranium ore gives rise to various waste residues, which
are referred to as mill tailings. These contain enhanced levels of naturally occurring radionuclides andthe radon is emitted from the mill tailings for many years following disposal because of the longradioactive half-life of the parent 238U. In assessing the radiological impact of mill tailings, it isnecessary to allow for the number of years over which the emissions will continue. Previously, theCommittee estimated collective doses from mill tailings assuming that the emissions of radon continued
for 10,000 years [U6]. Since the year 2000, improvements have been made to the treatment of milltailings and the rehabilitation of areas where mill tailings had been disposed of. As discussed in thecompanion report on radiation exposures from electricity generation (see annex B), enhanced levels ofradon emissions may continue for periods of less than 100 years. The Committee therefore agreed thatthe best estimate of the period of radon emissions from mill tailings should be reduced to 100 years, butthat the effect of considering periods of 500 years and 10,000 years should also be evaluated. The sameconsiderations also apply to the disposal of mine spoil or ash from coal used for electricity generation.

C. End points considered
18. Taking account of the considerations discussed in sections A and B above, the Committee hasdeveloped its methodology to estimate the following characteristic individual doses (expressed as
effective doses, in sieverts, Sv) from each radionuclide discharged continuously at a rate of 1 Bq/s:(a) For discharges to atmosphere, the characteristic individual dose in the 100th year (as noted
above, the dose in the 100th year of continuous discharge is equal to the integrated dose to
100 years from 1 year’s discharge) at 5 km from the discharge point (annual individual doses arealso calculated at 50, 300, 750 and 1,250 km from the discharge point for use in estimatingcollective doses);(b) For discharges to freshwater bodies, the characteristic individual dose in the 100th year;(c) For discharges to a marine environment, the characteristic individual dose in the 100th year.
19. In addition, integrated collective doses (man Sv) from each radionuclide discharged continuously
at a rate of 1 Bq/s for a year are calculated as follows:(a) For discharges to atmosphere, collective doses are integrated to 100 years (this is equivalent tothe collective dose in the 100th year of continuous discharge) in the following distance bands:
0−100 km (local), 100–500 km, 500–1,000 km and 1,000–1,500 km. The results for 100–1,500 km
are summed to constitute regional components of collective dose;(b) For discharges to freshwater bodies, the collective dose is integrated to 100 years. It should be
noted that for freshwater bodies, instantaneous mixing in a single water volume is assumed and,with this assumption, it is not possible to distinguish between local and regional components ofcollective dose;ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 31(c) For discharges to a marine environment, local and regional components of collective dose areintegrated to 100 years;(d) For globally circulating radionuclides only (3H, 14C, 85Kr and 129I), the global collective dose is
integrated to 100, 500 and 10,000 years.
20. The results can be scaled to obtain the characteristic individual and collective doses for specificdischarges and sources. In chapter III, results are calculated for discharges from different geographicalregions of the world, considering different population distributions (generic, nuclear coastal and nuclear
inland). Results are also calculated for a generic discharge location situated in a region of very low
population density, which may apply to sites of uranium mines and mill tailings.

D. Discharges considered
1.

Electrical energy production

21. One of the aims of the methodology is to allow exposures from discharges of radionuclides toatmosphere and water bodies to be assessed for a range of facilities associated with electricity
generation, covering both nuclear and non-nuclear power production. The methodology could also be
broadly used to assess routine discharges from other sources, but—as noted earlier—it is not intended
for site-specific or regulatory purposes. The assessment of exposures due to electricity generation
requires adequate and representative discharge data from the various facilities for input into the models
that implement the methodology. For example, discharges from the nuclear facilities involved in power
production and the fuel cycle listed in table 2 could be considered.
Table 2. Types of nuclear facility for which the methodology may be relevantLife-cycle stage
Front end

Facility types
Uranium mine with mill
Enrichment and fabrication

Nuclear reactor operation

Nuclear power reactors (e.g. pressurized water reactors,
boiling water reactors, fast breeder reactors, gas cooled
reactors, heavy water cooled reactors, and light-watercooled, graphite-moderated reactors)
Research reactors

Back end

Reprocessing
Decommissioning
Waste management

22. There are also a number of non-nuclear means of electrical energy production that may haveassociated discharges of naturally occurring radionuclides because of the materials used in construction
or at some point in their production cycle. The Committee decided that these should also be covered by
the methodology and they are considered in annex B. These could include the use of: (a) fossil fuels(coal, gas and oil); (b) geothermal sources; and (c) solar power, wind and biomass.

32 UNSCEAR 2016 REPORT2.

Radionuclides discharged

23. The radionuclides covered by the methodology are listed in table 1 and are considered by theCommittee to give rise to most of the doses from radioactive discharges from different energy sources.
This list of radionuclides includes those that have been demonstrated to give rise to the highest
integrated collective doses in previous assessments [C3, U6]. Naturally occurring radionuclides are also
covered by the methodology in order to assess the radiological impact of discharges from mining andmilling of primary materials, fuel production, and non-nuclear forms of electrical energy production [N4].
24. Although the methodology is designed to use data on discharges of individual radionuclides, inmany cases discharge data are provided only for groups of radionuclides as a whole. For example,
discharges may be given for “total alpha” and then this value will need to be apportioned among therelevant radionuclides included in the methodology (e.g. 239,240Pu, 241Am). For discharges specified as
total beta/gamma or particulate there may be many radionuclides that could be discharged, with themost important for nuclear power plants generally being 60Co, 90Sr, 134Cs and 137Cs but with significant
contributions from others such as 54Mn or for United Kingdom gas cooled reactors only 35S. In using themethodology, it will be necessary to specify a fraction of the discharge for each relevant radionuclide
and also which representative radionuclides to consider as defaults for the sum of discharges from any
radionuclides not included in table 1.
25. The methodology is designed to allow different assumptions to be made about the state ofequilibrium between components of the decay chains of the naturally occurring radionuclides. Accountis taken of the time necessary for ingrowth to occur and differences between the importance ofingrowth of progeny for different types of discharge and for different exposure pathways. Where
appropriate, the contribution of progeny is taken into account according to the following assumptions:(a) Where the progeny are short-lived compared to the parent, their contribution is not considered
explicitly but is generally included with the parent, for example, contributions from 90Y and137m
Ba are included in the dose coefficients for intakes by inhalation and ingestion for 90Sr and137
Cs, respectively;(b) Where the progeny are long-lived compared to the parent, their contribution is not consideredfurther if, over a period of 100 years, there is unlikely to be sufficient ingrowth of the progeny.
Examples where this is the case are 239Pu/235U, 240Pu/236U, and 241Am/237Np;(c) The key radionuclides in the decay chains of the naturally occurring radionuclides are
considered separately and other progeny are only considered if they make a significant
contribution to exposures or if it is reasonable to assume that they would be discharged insecular equilibrium with the parent.
Further information for specific radionuclides and exposure pathways and details of how this is appliedin practice are given in the appendix.

III. DESCRIPTION OF THE METHODOLOGY
26. When radionuclides are discharged to the environment there are various ways in which people canbe exposed to radiation. Figure I illustrates the more important exposure pathways.

Figure I. Exposure pathways following discharge of radioactive material to the environment

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 33

34 UNSCEAR 2016 REPORT

A. Assessment of doses from discharges to atmosphere
27. This section describes the methodology for estimating characteristic individual and local andregional components of collective dose to the public from a discharge of radionuclides to atmosphere.
The appendix gives full details of all of the equations used. The exposure pathways included in themethodology for discharges to atmosphere are:(a) Internal exposure from inhalation of radionuclides in the plume;(b) External exposure (beta and gamma emitters) from radionuclides in the plume;(c) External exposure from deposited radionuclides;(d) Internal exposure from ingestion of radionuclides incorporated in food.
28. Other exposure pathways such as internal exposure from inhalation of resuspended radionuclides
deposited on the ground and from inadvertent ingestion of soil are also possible. However a review ofthe literature indicated that the doses from these exposure pathways are negligible compared with thosefrom the exposure pathways listed above for continuous discharges of radionuclides to atmosphere [J5].
29. Characteristic individual doses are calculated for a member of the public assumed to be living at a
distance of 5 km from the point of discharge. Local and regional components of collective dose fromdischarges to atmosphere are derived directly by multiplying individual doses calculated at the
midpoints of distance bands or annuli (of 0–100, 100–500, 500–1,000 and 1,000–1,500 km) around thepoint of discharge by data specific to each geographical region on the number of people within eachannulus. Region-specific or generic values are used for parameters in the dose calculation as discussed
in the following sections.

1.

Population distribution

30. The Committee’s previous approach was based on simplified assumptions regarding population
density around nuclear sites representative of the situation in the 1980s. Given the importance of
population information in determining the overall magnitude of collective doses, and that there havebeen significant changes in population patterns throughout the world, the Committee decided to update
the information on population density. For comparison purposes and to examine the effect of different
population densities, four different sets of population data are considered in the methodology. As adefault case for comparison purposes, the calculation of collective dose from discharges to atmosphereis based on information on population densities for 2010 weighted by the population in eachgeographical region [F1]. The population distribution used for the default case is given in table 3 where
the same population density is used for each annulus (the population is the assumed density multipliedby the area of the annulus). In addition to values by region, a world-average population distribution has
been obtained as an average of the values for each region weighted by the populations in each region.
The regions considered in the methodology are those used by the United Nations Environment
Programme (UNEP), illustrated in figure II below.

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 35

Table 3. Default population distributions for different geographical regions
UNEP region

Population density,
weighteda (km−2)

Population within each annulus
0–100 km100–500 km

500–1,000 km

1,000–1,500 km

Africa1

7.9 × 106

2.5 × 107

5.9 × 108

1.9 × 10

3.1 × 108

Asia and Pacific

2.8 × 102

8.8 × 106

2.1 × 108

6.6 × 108

1.1 × 109

Europe1.3 × 102

4.0 × 106

9.7 × 107

3.0 × 108

5.0 × 108

Latin America and Caribbean

1.4 × 102

4.3 × 106

1.0 × 108

3.2 × 108

5.3 × 108

North America

3.2 × 101

1.0 × 106

2.4 × 107

7.6 × 107

1.3 × 108

West Asia

1.0 × 102

3.2 × 106

7.7 × 107

2.4 × 108

4.0 × 108

World averageb

1.6 × 102

5.0 × 106

1.2 × 108

3.8 × 108

6.3 × 108a

The population densities are taken from FAO [F1] and are values for each region weighted as described in this reference.

b

The world-average value is the average of the values for each region weighted by the population in each region.

Figure II. The geographical regions used in the methodology (taken from UNEP [U3])31. For nuclear power stations, a more specific analysis of population data is possible. In
collaboration with the Metadata and Socio-Economics Unit of UNEP/DEWA/GRID-Geneva, 3 ageographic information system has been used to express the number of people living within specified
distance bands around each station. Details of the approach and data used are given in electronic
attachment 3; in essence, the numbers of people living within different distances (100, 500, 1,000 and
1,500 km) around the site of each operating nuclear power station have been analysed. These data havebeen used to derive simplified averages by UNEP region for the population in each annulus. The
averages have been estimated separately for inland sites and for sites located on the coast, where thepopulation distribution is likely to be affected by the presence of a large body of water (see table 4).
Again a world-average value has also been calculated using all the available data. Because there were,at the time of analysis, no significant operating nuclear power stations in West Asia and only a coastal
3

United Nations Environment Programme (UNEP) Division of Early Warning and Assessment (DEWA) Global Resource
Information Database network (GRID) http://www.grid.unep.ch/.

36 UNSCEAR 2016 REPORTsite for Africa, no population densities are given for these in table 4. The results for the default
population distribution in table 3 could be used to consider nuclear sites in these regions if needed.
Table 4. Population distribution around operating nuclear power stations as a function of UNEP
region and location
Based on information provided by UNEP for the year 2008

UNEP regiona

Population within each annulusb
0–100 km100–500 km

500–1,000 km

1,000–1,500 km

COASTAL
Africa6

3.9 × 10

1.1 × 106

8.7 × 106

3.4 × 107

Asia and Pacific8.3 × 106

8.8 × 107

2.0 × 108

3.5 × 108

Europe3.3 × 106

6.4 × 107

1.3 × 108

1.5 × 108

Latin America and Caribbean

3.6 × 106

6.2 × 107

4.8 × 107

4.3 × 107

North America

4.7 × 106

4.0 × 107

7.0× 107

6.7 × 107

n/a

n/a

n/a

n/a

5.6 × 106

6.6 × 107

1.4 × 108

2.0 × 108

n/a

n/a

West Asia
World averagec

INLAND
African/an/aAsia and Pacific7

1.9 × 108

2.0 × 108

4.6 × 10

5.0 × 108

Europe

4.6 × 106

8.9 × 107

1.8 × 108

1.8 × 108

Latin America and Caribbean

4.6 × 106

1.5 × 107

3.7 × 107

3.4 × 107

North America

3.0 × 106

3.9 × 107

7.7 × 107

8.5 × 107

n/a

n/a

n/a

n/a

West Asia
World averagec

4.6 × 106

7.2 × 107

1.5 × 108

1.5 × 108aNote that at the time of analysis there were no significant operating nuclear power stations in West Asia or inland in Africa (seefigure II for definition of UNEP regions) and hence population data are not presented for these regions.

b

Data are presented to one decimal place, which is considered sufficient for this purpose.

c

The world-average is the average of the population in each annulus for all nuclear sites for which data are available.

32. The other case considered in the methodology is for uranium mines and mill tailings sites. Mostof these facilities are situated in areas of very low population densities, as illustrated in electronic
attachment 3. Therefore, the methodology also considers a population distribution with a density of
5 km−2 to gain insight into the effect on collective doses. Using this low population density out to1,500 km results in estimated collective doses from atmospheric discharges that are about two orders ofmagnitude lower than those estimated using the population distributions in tables 3 and 4. The low
value for population density should therefore only be used where appropriate and where it can beassumed to apply over the integration period for the collective dose.
33. There are two main approaches to assessing collective doses from ingestion of terrestrial foods.
The first uses information on the distribution of agricultural production to obtain an overall collective
dose from the production but with no indication of the individual doses that make up the collective
dose. The second approach, which the Committee adopted here, assumes that people obtain their food

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 37

from the area where they live. This enables both individual and collective doses to be estimated.
However, as discussed earlier, in many parts of the world it is not realistic to assume that people derive
all their food from local produce. This is particularly so within a small area that has a high population
density, and therefore low agricultural production. The Committee therefore decided that collective
doses from ingestion (where it is assumed that 100% of the food is obtained from the area considered)should only be calculated for distances of at least 100 km. For the characteristic individual dose, basedon an individual living at a distance of 5 km from the site, 25% of their food is assumed to be locally
produced as discussed earlier (paragraph 11).

2.

Dispersion in the atmosphere

34. In previous assessments, the Committee adopted a simple generic atmospheric dispersion model.
It considered the uncertainties associated with this approach in some detail in its 2000 Report [U6].
Given the generic nature of its assessment objectives, the Committee concluded that the long-term
sector-averaged Gaussian model was likely to provide an appropriate level of accuracy. Crawford et al.
[C5] acknowledge that simple Gaussian models continue to be used because they produce results that
often agree fairly well with measured data and because their results are relatively easy to obtain. The
Committee has decided to retain this approach.
35. The variation in activity concentration in air, Ca(x), with downwind distance beyond 1 km isapproximately given by the following equation:

Ca ( x) = D1 ⋅ Q ⋅ x − n(1)where D1 is the annual average dilution factor at 1 km (s/m3), Q is the discharge rate (Bq/s), x is thedownwind distance (km) and n is an empirically-determined index. The Committee agreed to retain theparameter values used in the UNSCEAR 2000 Report, in which a value of 5.3 × 10−7 s/m3 was found tobe the best approximation for D1 with a value for n of 1.42 for all radionuclides except noble gases,
tritium and carbon-14 [U6]. The dilution factor, D1, was found to be relatively insensitive to changes inthe values of parameters, except for wind speed and release height. The value of n was, however, foundto be dependent upon deposition velocity and inversion height. For noble gases, which do not deposit, avalue of n equal to 1.2 is retained. This value is also retained for tritium because, although tritium
rapidly exchanges with water in the soil and vegetation, it is quickly re-emitted to the atmosphere.
Carbon-14 also exchanges with carbon in soil and vegetation and partially returns to atmospherethrough plant and soil respiration. A value for n of 1.4 is therefore used for this radionuclide.
36. As part of the development of the current methodology, the activity concentrations estimatedusing this simple relationship were compared with the results obtained using a more complex Gaussianmodel, for a uniform wind rose (for which activity concentrations were averaged over 12 sectors) thatallowed for depletion of the plume but not radioactive decay. The results were found to be very similar(less than a factor of two difference).
37. A reduction factor is now included in the methodology to allow for the radioactive decay of shortlived radionuclides, notably radon, during dispersion over distances of hundreds of kilometres. Thisfactor is based on a mean wind speed, ua, of 2 m/s [I2].
38. The deposition rate of radionuclides at a specified distance (in Bq/(m2 s) is determined by theapplication of a simple “effective deposition velocity” that allows for both wet and dry deposition onvegetation, soil and other surfaces. An annual-average deposition rate of 0.002 m/s [U6] is used for all38 UNSCEAR 2016 REPORT

radionuclides, with the exception of the noble gases, tritium and 14C for which a value of zero isapplied. As noted above, tritium and 14C exchange with the soil and vegetation; however, this ismodelled using a specific activity approach (as discussed below). Because these radionuclides are
returned to atmosphere, the deposition rate is not used and the effective deposition velocity is taken to
be zero, even for 14C which is partly retained on the ground. This is consistent with the use of thegeneric approach to the modelling of dispersion applied in the UNSCEAR 2000 Report [U6] and, withthe application of a value of n of 1.42, 1.4 or 1.2 in equation 1. The use of these parameters ensures that
there is a balance between the activity discharged and the activity deposited within a distance of around2,000 km, thereby ensuring that the relevant exposures are included in the estimation of collective dose.
39. The calculated activity concentrations in air at ground level depend to some extent on the heightfrom which the discharge takes place and whether the discharge is from a point source or a wide area.
The heights of stacks from which discharges take place (i.e. point sources) vary considerably. The
discharges from uranium mine and mill tailings, on the other hand, take place at ground level over awide area. Electronic attachment 4 discusses the variation in estimated activity concentrations in air
with stack height and the differences for point and area sources. The differences in the concentration
profiles with stack height lessen with distance from the stack. For the case illustrated in electronic
attachment 4, where a uniform wind rose is assumed together with Pasquill Category D conditions for
65% of the time and all other categories for the rest of the time, the estimated concentration in air for aground-level discharge is about a factor of three higher than that from a 100 m stack at a distance of
5 km, and less than a factor of two higher than from a 30 m stack. Larger and smaller differences would
be seen for the individual Pasquill categories but this comparison applies to the annual average that isneeded for assessing routine continuous discharges. At a downwind distance of 100 km, the differencein concentrations for a ground-level discharge and one from a 100 m stack is about a factor of two. For
point and area sources, the differences in estimated activity concentrations at these distances are small.
The ratio of the time-integrated activity concentrations in air for an area source to that for a point sourceis 0.9 at 5 km and is 1 beyond 10 to 20 km (see electronic attachment 4). Given these findings, theCommittee agreed that, for its generic methodology, it was reasonable to adopt a single stack height of
30 m and to consider all discharges to atmosphere as point sources.

3.

Behaviour of radionuclides in a terrestrial environment

40. Radionuclides discharged to atmosphere may be transferred to plants by a number of processes,
primarily direct deposition onto the surface of plants and uptake by their roots from material depositedon the soil. The most important processes for the transfer of radionuclides through the terrestrial
environment are illustrated in figure III.

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 39

Figure III. The important processes for transfer of radionuclides through the terrestrial environment

41. There have been significant developments in the modelling of the transfer of radionuclides
through the terrestrial environment in recent years, partly as a result of studies following the Chernobyl
accident, see for example [C1]. The standard compilations of the concentration factors, linking different
parts of the environment, e.g. soil and plants, and other relevant parameters have been reviewed andupdated as part of international coordinated research activities [C1, I5].
42. For the purposes of this methodology, the Committee used the dynamic food-chain model
FARMLAND [B6], as implemented in PC-CREAM [S8], to derive integrated activity concentrations infood (fresh weight) per unit deposition rate; these are the activity concentrations integrated to 100 years
for continuous deposition at a rate of 1 Bq/(m2 s) for one year. These estimated concentrations takeaccount of the various environmental transfer, migration and loss processes illustrated in figure III andradioactive decay. Values of radionuclide transfer parameters for terrestrial environments were selectedfrom international compilations of data [I1, I5], supplemented by additional data [S8]. Expert
judgement was used to select values appropriate for the Committee’s purposes which are not overly
cautious. The values selected are intended to be appropriate for a range of environments and climates.
The FARMLAND model uses a standard approach to estimate the transfer of radionuclides throughterrestrial food chains and has been found to be in good agreement with measured data [S8]. In a
comparison exercise carried out as part of the IAEA EMRAS II (Environmental Modelling for40 UNSCEAR 2016 REPORT

Radiation Safety) programme [I6], the results obtained from FARMLAND were found to be in good
agreement with those results from other models. The derived activity concentrations applied in theCommittee’s methodology are summarized in table 5 and full details of the model and parameter values
can be found in Brown and Simmonds [B6] and Smith and Simmonds [S8].
Table 5. Activity concentrations in foodstuffs integrated to 100 years for continuous deposition at
a rate of 1 Bq/(m2 s) for one year [S8]Integrated activity concentration in food (fresh weight) (Bq/kg per Bq/(m2 s))
Radionuclidea

Cereals(Ccereal,unit)

Vegetables & fruits(Cveg,unit)

Milk & dairy products(Cmilk,unit)

Meat & offal(Cmeat,unit)

35

S

3.7 × 105

1.2 × 105

1.4 × 106

6.0 × 10654

Mn

8.6 × 104

1.3 × 105

9.0 × 104

1.9 × 10558

Co

5.3 × 104

9.3 × 104

3.9 × 105

7.4 × 10560

Co

1.1 × 105

1.5 × 105

2.3 × 106

1.3 × 10765

Zn

1.8 × 105

2.2 × 105

7.0 × 105

5.8 × 106

90

Sr

7.1 × 105

7.3 × 105

1.5 × 106

3.3 × 105

106

Ru

9.8 × 103

1.0 × 105

1.1 × 103

9.7 × 105129

I

6.0 × 105

2.3 × 105

3.1 × 105

2.1 × 105131

I

4.2 × 104

4.1 × 104

5.8 × 104

2.5 × 104

134

Cs

4.9 × 105

1.4 × 105

2.5 × 105

1.2 × 106

135

Csa

7.0 × 105

3.3 × 105

3.0 × 105

1.6 × 106

137

Cs

5.9 × 105

2.2 × 105

3.0 × 105

1.5 × 106

138

Csa

8.5 × 100

1.7 × 102

1.3 × 10−2

1.3 × 10−3

210

Pb

4.1 × 104

1.4 × 105

1.2 × 104

2.4 × 104

212

Pba

1.6 × 102

3.2 × 103

2.1 × 102

8.5 × 10−1

214

Pba

7.0 × 100

1.4 × 102

6.6 × 10−1

1.1 × 10−4

214

Bia

5.4 × 100

1.0 × 102

1.4 × 100

1.7 × 10−2

210

Po

4.9 × 103

9.5 × 104

8.2 × 104

1.1 × 105226

Ra

2.6 × 105

3.0 × 105

1.7 × 105

3.3 × 104

228

Raa

8.8 × 104

1.3 × 105

7.4 × 104

1.4 × 104

228

Tha

5.4 × 103

1.0 × 105

1.5 × 102

1.2 × 103

230

Th

7.9 × 103

1.1 × 105

1.7 × 102

2.2 × 103232

Th

7.9 × 103

1.1 × 105

1.7 × 102

2.2 × 103

234

Tha

3.4 × 103

6.6 × 104

1.1 × 102

6.6 × 101

234

U

5.6 × 104

1.5 × 105

5.7 × 104

2.8 × 105238

U

5.6 × 104

1.5 × 105

5.7 × 104

2.8 × 105

239

Pu

1.0 × 104

1.1 × 105

1.6 × 102

8.8 × 103

240Pu

1.0 × 104

1.1 × 105

1.6 × 102

8.8 × 103

241

Am

1.5 × 104

1.1 × 105

8.5 × 101

4.7 × 103

Included as progeny only.

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 41

43. The dispersion of 3H and 14C in the environment is more complex than that of other radionuclides
owing to the role of hydrogen and carbon in biological systems. These radionuclides therefore need to
be treated differently from other radionuclides. As previously, the Committee has assumed theseradionuclides reach equilibrium rapidly with their corresponding stable element, and used an approachbased on their specific activity.
44. The specific-activity model for tritium is based on the assumption that tritium behaves ashydrogen in the environment so that tritium discharged to atmosphere exchanges with the hydrogen inwater in air, soil, plants and animals. The concentration of tritiated water (HTO) and organically boundtritium (OBT) in plants is determined on the basis of the water content in plants and a partition factor
that takes account of the presence of exchangeable hydrogen in the dry weight of the plant [I5]. Fulldetails of the approach are given in the appendix. The concentrations of HTO and OBT in animal
products are determined using concentration factors that relate the concentrations in these products to
those in feed, drinking water and inhaled air.
45. The specific activity (or activity concentration) of 14C in elemental or stable carbon is determinedfrom the discharge by taking account of dispersion in the atmosphere and the concentration of stablecarbon in the atmosphere. The concentration of 14C, expressed in Bq/kg of stable carbon, is assumed tobe the same in plants as in air. The following equation is used to calculate the concentration of 14C interrestrial foods, Cf,C-14(x), at distance x from the discharge point:C=( x ) Cair,14C ( x ) ⋅f ,14 C

Sp
Sair

(2)

where Cair,C-14 (x) is the activity concentration of 14C in air at distance x (Bq/m3), Sp is the concentrationof stable carbon in the crop of interest (grams of carbon per kg fresh weight of crop) and Sair is theconcentration of stable carbon in air (grams of carbon per cubic metre of air).
Similarly the concentration of 14C in animal products, Cf,C-14 at distance x from the discharge point isgiven by:

C f ,14 C ( x ) =f C ⋅ Cpasture,14C ( x ) ⋅ S aSp(3)

where fc is the fraction of feed containing 14C (assumed to be 1 in the methodology), Cpasture,C-14 (x) is theconcentration of 14C in pasture at distance x from the discharge point (derived as for crops using
equation (2)), Sa is the concentration of stable carbon in the animal product (grams of carbon per
kilogram fresh weight) and Sp is the concentration of stable carbon in the pasture (grams of carbon per
kilogram fresh weight of pasture).
46. The values of the model parameters forming part of the specific-activity model for 3H and 14C aresummarized in table 6 (see the appendix for more information). The symbols used here and in the
workbooks are those used in the IAEA publication describing this model for ease of reference.

42 UNSCEAR 2016 REPORT

Table 6. Parameters and their values associated with the specific-activity model for 3H and 14C(taken from [I5])
Parameter

Symbol

Value

Unit

Stable carbon concentration in air

Sair

0.2

g of C/m3

Stable carbon concentration in cereals (fresh weight)

Sp

390

g of C /kg

Stable carbon concentration in vegetables (fresh weight)

Sp30

g of C /kg

Stable carbon concentration in pasture (fresh weight)

Sp

100

g of C /kg

Stable carbon concentration in cow meat (fresh weight)

Sa

200

g of C /kg

Stable carbon concentration in cow milk (fresh weight)

Sa65

g of C /kg

Fraction of feed containing 14C

fc

1

—

Fractional water content of cereals (fresh weight)

WCp

0.12

L/kg

Fractional water content of vegetables (fresh weight)

WCp

0.92

L/kg

Fractional water content of pasture (fresh weight)

WCp

0.76

L/kg

HTO:H2O water vapour pressures

γ

0.909

—

Absolute humidityHa

6 × 10−3

L/m3

Relative humidity

RH

0.7

—

Empirical constant

CRs-a

0.3

—

Concentration ratio for HTO intake through milk

CRa,HTO

0.87(Bq/kg fresh weight)per (Bq/L)

Concentration ratio for HTO intake through meat

CRa,HTO

0.66(Bq/kg fresh weight)per (Bq/L)

Water equivalent factora, cereals

WEQp

0.56

L/kg

Water equivalent factor, vegetables

WEQp

0.51

L/kg

Water equivalent factor, pasture

WEQp

0.56

L/kg

Rp

0.54

—

Concentration ratio for OBT intake, milk

CRa,OBT

0.24(Bq/kg fresh weight) per(Bq/kg dry weight)

Concentration ratio for OBT intake, meat

CRa,OBT

0.4(Bq/kg fresh weight) per(Bq/kg dry weight)

Partition factorba

The water equivalent factor is the mass (kg) of water produced per unit mass (kg) of dry matter combusted.

b

The ratio of the concentration of non-exchangeable organically-bound tritium in combustion water to that in tissues (e.g. leaves).

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 43

4.

External and internal exposure(a) External exposure
47. The Committee considers that its previous methodology [U6] for estimating the dose fromexternal exposure due to immersion by assuming a semi-infinite cloud of radionuclides is still
appropriate. It acknowledges some limitations with this approach, most notably where activity
concentrations in air are likely to be non-uniform over a distance of a few hundred metres to one
kilometre from the point of discharge. However, given the distances over which the Committee’s
assessments are made, the application of dose coefficients for external exposure used previously, whichwere based on [E2], is unlikely to be a source of significant error. The dose coefficients applied in theCommittee’s methodology are presented in table 1.
48. The annual characteristic individual dose from immersion in the plume, HE(ex,cloud),i forradionuclide, i (Sv/a) is estimated at a series of distances, x (m), from the discharge point by assuming
100% occupancy (the fraction of the time that is spent at a particular location) at locations at thosedistances (Oann (s/a)) and applying the relevant dose coefficients for external exposure due to immersion
in the plume, Dex, cloud,i from table 1 (Sv per (Bq s/m3)). Account is then taken of the fraction of timespent outdoors, Oout (dimensionless) and the shielding effect of buildings, Lcloud (dimensionless) as
shown in equation (4) and discussed below.

H E (ex,cloud),i=( x ) Cair,i ( x ) ⋅ Dex,cloud,i ⋅ Oann ⋅ ( Oout + (1 − Oout ) ⋅ Lcloud )(4)

Where Cair,i (x) is the activity concentration of radionuclide i in air (Bq/m3) at location x. The annual
individual dose from external exposure due to deposited material, HE(ex,deposit),i (x) (Sv/a) at a distancex (m) from the discharge point is calculated from the time-integrated activity concentration (to100 years) of the radionuclide i on soil at the location of interest, di ( x) (Bq s/m2), the length of thedischarge, tdischarge (s/a) the relevant dose coefficient for external exposure due to deposition, Dex,deposit,i(Sv/ Bq s/m2), the fraction of time spent outdoors, Oout (dimensionless) and the dimensionless location
factor that takes account of the shielding effect of buildings, as shown in equation (5) and discussed
below.

H E ( ex,deposit ),i ( x=) di ( x ) ⋅ tdischarge ⋅ Dex,deposit,i ⋅ ( Oout + (1 − Oout ) ⋅ Ldeposit )(5)

49. The Committee’s previous methodology applied the dose coefficients for external exposure fromradionuclides deposited on soil provided by Beck [B1]. An alternative approach was used for theCommittee’s 2013 Report of the levels of radiation exposure due to the nuclear accident at the
Fukushima-Daiichi nuclear power station [U12] based on the model published by Petoussi-Henss et al.
[P2]. The effective dose coefficients calculated using this more recent model [P2] were similar to orwithin a factor of two of those used previously. For the sake of simplicity and consistency with its
2013 Report [U12], the Committee decided to apply dose rate coefficients for external exposure, edep(nSv/h per kBq/m2) from Petoussi-Henss et al. These values are based on calculations using a voxel
phantom and the latest definition of effective dose [I13] (the use of the previous definition of effective
dose in the methodology would make little difference to the calculated effective doses for external
exposure). Where data for particular radionuclides were not available, the data set was supplemented byfactors derived from the United States Federal Guidance Report No. 12 [E2].

44 UNSCEAR 2016 REPORT

50. Petoussi-Henss et al. [P2] modelled the geometry as an infinite mono-energetic plane sourceshielded by a soil layer of depth 0.5 g/cm2. Such an assumption gives a good description of theradiation field after wet deposition on the ground and after dry deposition following the first rainfall[I15, J1]. With this geometry, the air kerma from a 137Cs (137mBa) source is reduced by a factor of 0.67
compared to a plane source on the ground–air interface. The effective dose coefficients were derived
assuming a constant density of air of 1.2 × 10−3 g/cm3 and a soil density of 1 g/cm3, which is considered
to be representative of the upper 2 cm of soil [P2].
51. In order to model the reduction of gamma dose rates because of migration into the soil, theCommittee adopted an attenuation function r(t) derived empirically from data on the migration of 137Cs[G2]:

 ln2  ln2 r ( t )= p1 ⋅ exp  −t  + p2 ⋅ exp  −t T1 
 T2 

(6)

where p1, and p2 are dimensionless parameters with values derived empirically, T1 and T2 are the initialand final times, respectively, following deposition and t is the time after the deposition of interest. The
parameter values in this function were determined from gamma-spectrometric analyses of over 400 soil
samples taken during 1986–2003 in the areas of Germany (specifically Bavaria), the RussianFederation, Sweden and Ukraine affected by deposition from the accident at the Chernobyl nuclear
power plant in 1986 (e.g. [G2, J3, L2, S2]). For the purposes of the Committee’s methodology,
parameter values derived for areas distant from the Chernobyl plant were applied, namely p1 was taken
to be 0.5 and p2 to be 0.5; T1 was assumed to be 1.5 years while T2 was assumed to be 50 years. Thus
the relationship below with T = 100 years was applied to the value of the dose coefficient from
deposition ( edep ) given in [P2]:
t

t 


ln 2 
− li +t
T1 

edep = =edep ⋅ er ( t ) dt edep ⋅  p1e 

0
0


∫

( − lit )∫+ p2 eln 2  − li +
t
T2  





dt

(7)

Substituting for p1, p2, T1, T2 in the above equations and changing from a decay constant to a half-life,
the effective dose (Sv) from external exposure in year 100, edep, is given by: ln 2 ln 2   ln 2 ln 2  


−
+
−
+
T
T 



0.5
0.5
1 − e  t1/2 1.5   +1 − e  t1/2 50   =edep edep ⋅ 
ln
2
ln
2
ln
2
ln
2





+
 t + 50  t1/2 1.5 
1/2


(8)

This can be simplified to ln 2 ln 2   ln 2 ln 2   



−
+
−
+
T
T
0.75t1/2
25t1/2 t1/2 1.5   t1/2 50   



=
+
1− e
1− e
edep edep ⋅ (t1/2 + 1.5) ln 2  (t1/2 + 50) ln 2 




 


As an example, for caesium-137 and caesium-134 the following values can be calculated:(9)

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 45()

edep 137 Cs+137m Ba = 1.55∙× 10−7 Sv/(Bq/m2)()

edep 134 Cs = 6.23∙× 10−8 Sv/(Bq/m2)

(10)

(11)

52. Despite differences in climate, little variation in indoor occupancies has been found between
countries and there is no evidence of a significant difference due to climate [A1]. Studies show thatindoor occupancies range from around 84 to 91% [G1, O2]. The Committee considered that values ofbetween 70% (for an outdoor worker) and 90% (for pensioners and indoor workers) were appropriate
for a Japanese population [U12]. In accordance with its previous methodology, the Committee assumed
an indoor occupancy factor of 80% [U6]. In equations (4) and (5) above the outdoor occupancy Oout isused; this is simply a fractional value and is 1 − 0.8 = 0.2.
53. People indoors receive some protection from external exposure due to gamma-emitting
radionuclides in the plume as it passes overhead. The reduction in ambient dose equivalent rate depends
on the nature and structure of the building as well as on the energy of the radiation. A number of studiesof the shielding effect of various types of buildings have been published, particularly following theChernobyl accident [E2]. For example, the work of Le Grand [L1] and Jacob and Meckbach [J1, J2]
indicated that shielding factors (i.e. the ratio of the ambient dose equivalent rate indoors to that
outdoors) for radiation from airborne radionuclides ranged from 0.01 to 0.7 for single-family houses invarious European countries. For multi-storey buildings the shielding factor varies with storey, with thelowest values in the basement and first storey and the highest value for the top storey. Le Grand [L1]estimated that shielding factors appropriate for an airborne plume with photon energy of 0.68 MeV
ranged from 0.01 to 0.1 for the ground floor and from 0.09 to 0.4 for the upper floors of buildings. For
its methodology, the Committee has selected a shielding factor of 0.2 for radionuclides in the plume,
being within the range applicable to single and multi-storey buildings.
54. The UNSCEAR 1982 Report [U4] provided an overview of shielding factors appropriate fordeposited radionuclides, with values ranging from 0.05 in office buildings, 0.2 in masonry homes, and
0.4 in wooden buildings. This information was based on work by Burson and Profio, which showed arange of shielding factors depending on the nature of the building, with the lowest factors being forbasements of large multi-storey buildings [B8]. The United States Nuclear Regulatory Commission[N6] and others have suggested a generic shielding factor of 0.5 for regulatory purposes, which it
considers appropriate for photon energies above a few hundred kiloelectronvolts [E2]; for photons oflower energy, use of such a value may considerably overestimate the dose equivalent [K6]. In theirassessment of the radiation exposure of the population of the United States of America [N3, S1], theNational Council on Radiation Protection and Measurements (NCRP) applied a shielding factor of
0.59 (±6%) for terrestrial gamma radiation, which it considered to be appropriate for photon energies ofgreater than a few hundred kiloelectronvolts.
55. Location factors are used to express the ratio of external exposures in terms of ambient dose
equivalent rate at a specific location to the ambient dose equivalent at the location for which calculated
or measured ambient dose equivalent was obtained (for example, this might be between the exposures
outdoors in an urban and a rural environment). This is a broader term than shielding factor which justallows for the reduction of ambient dose equivalent from being indoors. For the purposes of thismethodology, the two terms can be considered to be equivalent, because it is the reduction in external
exposures from being indoors that is relevant. Golikov et al. evaluated the location factors for gamma
radiation in air in rural and urban environments in the Russian Federation five years after the Chernobyl
accident [G2]. In urban areas, they derived location factors for living areas that varied from 0.01 for a
multi-storey house to 0.09 for a wooden house, while they estimated the factor for buildings where

46 UNSCEAR 2016 REPORT

people work to be around 0.02. In rural areas, location factors for living areas ranged from 0.02 (for a
multi-storey house) to 0.13 (for a single-storey wooden building). This work formed the basis for thetime-dependent location factors used in the Committee’s 2013 Report on the levels and effects ofradiation exposure due to the nuclear accident at the Fukushima-Daiichi nuclear power station [U12].
Three different time-dependent location factors were used for that assessment: (a) for paved externalsurfaces, (b) for non-paved external surfaces, and (c) for inside buildings. This time dependence isimportant when assessing exposures due to an accident. However, for routine releases it is the dose in
the 100th year of continuous discharge that is of interest and the effect of using a time-dependent
location factor is small compared to differences in location factors for different building types. These
results also show a greater degree of shielding from deposited radionuclides than from radionuclides in
the plume, and this is reflected in the Committee’s choice of values for its methodology. In accordancewith its previous methodology [U6], the Committee has used a shielding (location) factor of 0.1 toallow for the reduction in ambient dose equivalent from external exposure due to deposited material
while indoors. This value was chosen as representative of occupancy in a single-storey building andwas considered to be reasonably consistent with the result that would be obtained if an approach basedon time-dependent location factors had been used.

(b) Internal exposure from inhalation
56. As previously, the Committee uses a standard approach to assess the internal exposure frominhalation of radionuclides in the air following discharges to atmosphere. This approach uses theestimated activity concentration of a radionuclide in air (section III.A.2 above), an appropriate
breathing rate and the relevant dose coefficient for intake by inhalation (table 1).
57. The Committee continues to use a nominal adult breathing rate of 20 m3/d [U7] in its
methodology to maintain consistency with previous assessments, recognizing differences between thisvalue and that used by ICRP to derive dose coefficients. As discussed earlier, the Committee considers
that the dose coefficients for inhalation given in ICRP Publication 119 [I14] using the ICRP model ofthe respiratory tract [I11] continue to be appropriate (table 1). No account is taken of any reduction in
activity concentrations in air when people are indoors because the Committee considers this to be asecond-order effect that will not have a significant effect on the estimated doses.
58. A different approach is needed for the inhalation of isotopes of radon. In the UNSCEAR 2006
Report, the Committee considered a number of issues related to the assessment of doses from theinhalation of radon and its short-lived progeny [U9]. It concluded that it should continue to use a value
for the radon dose coefficient of 9 nSv per (Bq h)/m3 for this methodology, with equilibrium factors(the ratio between the activity of the short-lived radon progeny and the activity that would be atequilibrium with the radon parent) of 0.6 for radon outdoors and 0.4 for radon indoors (indoor
occupancy being 80% as discussed earlier) to calculate annual effective doses from inhalation of radonand its short-lived progeny (see the appendix for more details).

(c) Internal exposure from ingestion
59. Members of the public may be exposed through ingestion of food that contains radionuclides
resulting from discharges to atmosphere. As indicated in section III.A.3, radionuclides discharged toatmosphere may be transferred to human food by a number of routes. The annual effective dose fromthe ingestion of food, f, containing radionuclide, i, at distance x (m) from the discharge point in regionr, (HE(ing),f,r,i (x)) (Sv/a) is given by equation (12):ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 47

H E ( ing ), f ,r ,i ( x=
) C f ,i ( x ) ⋅ Ding,i ⋅ Flocal ⋅ I f ,r(12)

where Ding,i (Sv/Bq) is the effective dose coefficient for ingestion of radionuclide i as given in table 1;
Flocal is the fraction of food that is locally produced (dimensionless); If,r is the average consumption rateof food f in region r (kg or L per year) as discussed below; Cf,i is the activity concentration ofradionuclide i in terrestrial food f at distance x from the discharge point (Bq/kg) given by:C=
C f ,unit,i ⋅ di ( x )f ,i ( x )(13)where Cf,unit,i is the activity concentrations in foodstuffs integrated to 100 years for continuous deposition
at a rate of 1 Bq/(m2 s) for one year (Bq/kg per Bq/(m2 s)) given in table 5 and di ( x) is the deposition rate
at distance x of radionuclide i (Bq /(m2 s)). The food contamination monitoring and assessment
programme (GEMS/Food) database for 2012 — part of the WHO Global Environment Monitoring
System — formed the basis for the population-weighted consumption rates used by the Committee tocalculate doses from ingestion for each UNEP region [W2]. For the Committee’s methodology, the World
Health Organization (WHO) combined the data for specified food groups and reconfigured them for the
UNEP regions and subregions (figure II). The resulting values of If,r for terrestrial foods are presented intable 7. World-average values have also been derived; these are the averages of the values for each region
weighted by the population densities given in table 3. As discussed in paragraph 11, for the estimation ofcharacteristic individual doses, it is assumed that only 25% of foods consumed are produced locally (Flocal
is 0.25). However in estimating collective doses, people are assumed to obtain 100% of their food fromthe area where they live (Flocal is 1.0), this assumption is considered reasonable because collective doses
are estimated for areas of 30,000 km2 or more.
Table 7. Annual average per caput consumption rates of terrestrial foods by UNEP region
Annual average per caput consumption ratea (kg) for each food type
UNEP region

Cereals

Vegetables andfruit

Milk and dairy
products

Meat and offal

Africa

130

220

31

17

Asia and Pacific

140

240

45

30

Europe

110

280

120

65

Latin America and Caribbean

110

210

77

63

North America

88

260

120

100

West Asia

140

180

43

34

World averageb

130

230

65

44

a

Values are rounded to two significant figures in order not to imply great precision.

b

World-average values are the averages of the values for each region weighted by the population densities given in table 3.

60. The Committee considered that the inclusion of all food groups and all 13 regions in the WHO
GEMS/Food [W2] would imply a level of accuracy that was inconsistent with other aspects of therevised methodology. Although the dietary data in GEMS/Food were based on clusters of countries
with similar dietary habits, this clustering would not have been consistent with other information for
UNEP regions used to calculate doses (e.g. population distribution and energy production). Therefore, asimplified approach was used based on the UNEP regions shown in figure II; any differences in intakes
and hence doses because of this simplification were deemed commensurate with those due to othersimplifying assumptions made in the methodology.

48 UNSCEAR 2016 REPORT

61. As discussed in section III.A.3, activity concentrations in foods integrated to 100 years forcontinuous deposition at a rate of 1 Bq/(m2 s) for one year have been determined using combined foodand radionuclide-specific transfer parameters. These estimated concentrations take account of a rangeof physical processes that do not depend on the chemical element (such as interception by plantsurfaces) and biochemical factors that do depend on the chemical element, such as root uptake andtransfers from animal feed to meat and milk. For tritium and carbon-14, combined transfer parameters
are defined per unit activity concentration in air based on a specific-activity approach. The intake ofradionuclides is then determined using the population-weighted food consumption rates for the different
geographical regions. This is a slightly different approach from that used previously by the Committee,
where simple global average intakes were used, (see electronic attachments 1 and 2); however, theoverall difference in the resulting dose estimates is not significant.
62. A limited sensitivity analysis was carried out to determine the applicability of the generic foodchain model for different regions of the world and whether the foods and related consumption rates
were appropriate. This analysis is outlined in electronic attachment 4 and the Committee concluded that
it was reasonable to use the food-chain model and the consumption data by geographical region for the
purposes of assessing worldwide exposures.

B. Assessment of doses from discharges to a freshwater environment
63. The extent of dispersion of radionuclides in freshwater bodies varies significantly depending on
characteristics of the water body, particularly the volume of the body into which the discharge occurs
and the water flows. However, the Committee decided that there would be value in developing genericdose-calculation factors that provide an estimate of the integrated activity concentrations in water froma discharge of each radionuclide for a year, allowing for dispersion within different types of water
body. An approach for estimating individual doses and for calculating regional components ofcollective dose was developed on this basis. (As discussed previously, for discharges to rivers using thesimple, generic approach adopted in the methodology, it is not possible to distinguish between local andregional components of collective dose).
64. The following exposure pathways are considered for discharges to freshwater bodies (seefigure I):(a) Internal exposure from ingestion of radionuclides in drinking water;(b) External exposure (beta and gamma emitters) from radionuclides in sediments deposited on
the riverbank;(c) Internal exposure from ingestion of radionuclides incorporated into freshwater fish;(d) Internal exposure from ingestion of radionuclides incorporated into terrestrial foods because of
irrigation.
65. Other exposure pathways such as internal exposure from inadvertent ingestion of water andsediment, external exposure when swimming or boating, and external exposure from radionuclides
deposited on irrigated land, are also possible. However, the exposure pathways above have been foundto be the largest contributors to individual and collective doses following continuous discharges towater bodies [J5]. The external exposure pathway, (b) above, is included only in the calculation ofcharacteristic individual dose. This pathway is unlikely to be a significant contributor to collective
doses from aquatic discharges from nuclear power stations and reprocessing plants, as demonstrated by

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 49

past assessments (e.g. [J6]). Individual doses are estimated from individual consumption rates andoccupancy data, while collective doses from ingestion are calculated using region-specific data on total
fish catches, water abstraction and agricultural production. Thus, the resulting individual and collective
dose estimates are not directly linked in the way that they are for discharges to atmosphere. In effect,collective doses from discharges to a freshwater body are based on the total usage of water or total
amount of produce from the generic water body, irrespective of the location of people.
66. The Committee’s methodology initially considered three types of environment for assessing doses
arising from aquatic discharges to freshwater bodies: (a) lakes, (b) small rivers and (c) large rivers.
However, the Committee agreed that the dispersion of aquatic discharges from facilities on the shoresof a large lake can be assessed using the same approach as for large rivers. Given the differences incharacteristics and hence dispersion in different rivers and lakes throughout the world, the Committee
agreed that the distinction between lakes and large rivers was a second-order effect and did not need tobe considered further for its purposes. It should be noted that the Committee’s methodology does not
apply to closed lake systems, i.e. those without rivers feeding into or out of the lake.
67. Two types of river are defined to allow for the range of discharge conditions that may occur. For
example, inland nuclear facilities (with the exception of mines and mills) are assumed to discharge intolarge rivers, and non-nuclear facilities and mines into small rivers. The dimensions and flow rates
associated with a range of rivers into which nuclear and other facilities discharge were reviewed inorder to select appropriate parameter values for the Committee’s methodology.

1.

Dispersion of radionuclides in freshwater bodies

68. The Committee’s methodology for deriving activity concentrations of radionuclides following
their discharge into freshwater bodies assumes that complete mixing occurs immediately. This is a
simplification because for rivers complete dilution would occur over some tens of kilometres
downstream, the actual distance depending on the width of the river, which is in turn related to thevolumetric flow rate. Nevertheless, significant mixing does occur over relatively short distances. As an
illustration of this, table 8 shows results from using the NCRP screening model, which is designed forthe purpose of demonstrating compliance with environmental standards for discharges of radionuclides[N2]. The table indicates that the activity concentration in unfiltered water assuming complete mixingof the radionuclide across the total width of the river essentially equals the actual activity concentrationin unfiltered water in the plume (i.e. the value of the partial mixing correction factor approaches one)for rivers up to a width of around 100 m at a downstream distance of around 5 km. For wider rivers, thevalue of the correction factor would be of the order of one to five at this distance; complete mixing
would, however, occur over a distance of around 100 km for such rivers. Specific information isavailable for the Techa–Iset–Tobol–Irtysh–Ob river system in the Russian Federation [N5]. This is acomplex river system and complete mixing does not occur until some distance downstream. (At 27 km
downstream there was not complete mixing with nearly a factor of 5 variation in measured
concentrations at different locations across the river.) However, the abstraction point for drinking water
is likely to be at some distance downstream from the discharge point and fish will be caught over large
sections of the river. Given this and the generic nature of the revised methodology, the distances overwhich collective doses are calculated, and other uncertainties associated with the approach, theCommittee considered it unlikely that assuming complete mixing would significantly affect thecollective dose estimate. Although the dose to an individual obtaining their drinking water from theimmediate vicinity of a source would depend on the location of the abstraction point relative to thepoint of discharge, because of the generic nature of the approach, assuming complete mixing is also
appropriate for individual dose estimation.

50 UNSCEAR 2016 REPORT

Table 8. Correction factors to complete mixing in a river plume as a function of downstream
distance from the point of discharge for different river widths [N2]
The correction factor for partial mixing represents the ratio of the activity concentration in water in the plume to the activity
concentration in water if, at the same point, complete mixing in the total width of the river were assumed. Complete mixing
would give a value of 1aPartial mixing correction factor at various river widthsa

Distance downstream
(km)

10 m

50 m

100m

200 m

400 m

0.1

3.2

6

8

10

12

1

1

2.4

3.2

5

8

5

1

1

1.6

2.6

3.6

10

1

1

1

1.9

3

50

1

1

1

1

1.4

100

1

1

1

1

1

Values are rounded to one decimal place.

69. For the purposes of the Committee’s methodology, the activity concentrations of radionuclides inriver water are therefore derived assuming that the discharged radionuclides are dispersed uniformly
across the river. Under these conditions, the activity concentrations in water depend upon thevolumetric flow rate of the river alone, which, in turn, is primarily determined by the width of the river.
Thus, it is assumed that, for all freshwater pathways, the activity concentration in unfiltered freshwater
is given by:

Cuw,i = Qi / F(14)

where Cuw,i is the activity concentration of radionuclide, i, in unfiltered water at the discharge point(Bq/m3); Qi is the discharge rate of radionuclide, i, (Bq/s) and F is the volumetric flow rate of the riverat the point of discharge (m3/s). The activity concentrations of radionuclides in filtered water are
determined from the suspended sediment load and the radionuclide-specific partition factors for
freshwater environments. These activity concentrations are then used, for example, to derive activity
concentrations of radionuclides in freshwater fish. To derive activity concentrations of radionuclides indrinking water, a water treatment factor is applied.
70. For the large river (and as noted earlier, for lakes), a width of 240 m and flow rate of 1,000 m3/s
are assumed for the purposes of the methodology. Table 9 gives some illustrative flow rates and otherdimensions for a range of rivers of different sizes throughout the world. The values assumed for a large
river are consistent with the data for major sections of the Rhône, Loire, Danube and Rhine; these arerivers into which a number of nuclear installations discharge radionuclides. They are also similar toother major rivers throughout the world, as shown in table 9, although some rivers have greater flows,
notably the Amazon.

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 51

Table 9. Flow rates and dimensions of some typical rivers and those assumed in the methodology
River

Country

Flow rate(m3/s)

Length in
specifiedcountry (km)

Width
(m)

Depth
(m)

5 × 102

6

AmazonaPeru, Columbia, Brazil

2.0 × 105

6.9 × 103Danubeb

Germany, Austria, Slovakia,
Hungary, Croatia, Serbia,
Bulgaria, Romania,
Moldova, Ukraine

3 × 103

3 × 103

Gangesa

India, Bangladesh

1.6 × 104

2.5 × 103

Kennet (tributary of Thames)

United Kingdom

1 × 101

3.3 × 101

17

1

LoirecFrance

7.8 × 102

5.6 × 102

2.6 × 102

3

Mississippia

United States of America

1.6 × 104

3.7 × 103

Niled,eEgypt

1.8 × 103

1.5 x 103

Paraná f

Brazil, Paraguay, Argentina

1.7 × 104

4.8 × 103

Pearl

Chinag

2.2 × 103

2.2 × 103

RhônecFrance1.2 × 103

4.0 × 102

2 × 102

7

ThamesUnited Kingdom

5 × 10

2.3 × 1050

2

Yangtze

g

China

4

3.2 × 10

6.3 × 10

Yellow

Chinag2.1 × 103

5.5 × 103h12

1.8 × 104

3

ASSUMED VALUES IN THE METHODOLOGY
Methodology—small riveri

1 × 101

1 × 102

30

1

Methodology—large riveri

1 × 103

5 × 102240

4aMilliman and Farnsworth (2011) [M2].

bMaringer (2000) [M1].

cSmith and Simmonds (2009) [S8].

dWahaab and Badawy (2004) [W1].

eHamza (2014) [H1].

fEncyclopaedia Britannica (Parana-River) [E3].

g

Information provided by the Chinese delegation.

hHilton et al. (2003) [H4].

i

The dimensions of the small river are similar to those of the River Kennet in the United Kingdom; those of the large river are
similar to those of major sections of the Rhône and Loire in France and other rivers.

71. For the small river, a flow rate of 10 m3/s (width of 30 m and depth of 1 m) is assumed in themethodology. This is similar to that of the River Kennet in the United Kingdom.
72. The measured suspended sediment loads of the Colorado and Mississippi Rivers have been shownto vary by three or more orders of magnitude [H2, S10]. However, a value of 5 × 10−4 t/m3 is considered
to be appropriately representative within this range and is assumed for large rivers; this also agrees withother data presented for the Colorado River [V1].

52 UNSCEAR 2016 REPORT

73. The suspended sediment loads of small rivers also vary by several orders of magnitude,
depending on flow rate. For low flow rates, values in the range 2 to 5 × 10−5 t/m3 may be considered tobe typical (see, for example, Hejduk and Banasik [H3]). A value of 2 × 10−5 t/m3 was adopted for the
purposes of this methodology.

2.

Behaviour of radionuclides in a freshwater environment(a) Transfers in freshwater environments
74. The key element-specific parameters that account for the transfer of radionuclides in an aquatic
environment are concentration factors that relate the concentrations of elements in fish to theirconcentrations in water, and partition coefficients that relate the distribution of elements between
sediment and water. Standard compilations of these data [I3, I5] have also been updated since theCommittee’s assessments were issued in the UNSCEAR 2000 and 2008 Reports [U6, U10, U11].
75. The activity concentration of a radionuclide in an aquatic food is derived by applying anequilibrium concentration factor, defined as the ratio of the concentration of the radionuclide in theaquatic food (fresh weight) at equilibrium to its concentration in water (Bq/kg per Bq/L) (note in somecompilations of data the term “bioaccumulation factor” is used for these values). The most recent
international compilation of concentration factors for freshwater environments [I5] provides transfer
parameter values appropriate for equilibrium conditions. Where the available data permit, ranges ofobserved values are presented together with mean values. Those mean values have been used for the
purposes of this methodology supplemented by data from other sources, where necessary; the values
adopted by the Committee for this methodology are given in table 10.
76. The transfer of radionuclides between the water column and suspended and bottom sediments
depends on both the characteristics of the water body and the chemical and physical characteristics ofthe radionuclides themselves. The partition coefficient, Kd, relates to the partitioning of radionuclides
between the solid and aqueous phases and is expressed as the ratio of the activity of a radionuclide perunit dry weight of sediment at equilibrium to the activity of that radionuclide per unit volume of water(Bq/kg per Bq/L) (this can also be referred to as a distribution coefficient). The most recent
international compilation of Kd values for freshwater environments is presented in [I5], although theCommittee also used values from other sources, where data were not available in [I5]. The data used
are given in table 10 and are intended to represent generic best estimates, recognizing that Kd values
vary depending on the characteristics of the water body.

Table 10. Values of radionuclide-specific parameters used in the model for freshwater environments
Radionuclide3

H (HTO and OBT)

Concentration factor for freshwater fish,
Bfish (L/kg) (fresh weight)

1.0 × 100 [N2]

Activity concentrations in sediment (dry weight) in the
100th year of continuous discharge at 1 Bq/s (Bq/kg)
Small river[S8]

Large river
[S8]

0

0

C

4.0 × 10 [I5]5.0 × 1035

S

8.0 × 102 [N2]

54

Mn

58

5

−4

−6

Partition coefficient, Kd(m3/t)

1.0 × 1000 [I2]

5.0 × 100

1.0 × 10

5.0 × 100 [I2]2.0 × 10−2

1.8 × 10−4

5.4 × 10−1

2.0 × 102 [B4]2.4 × 102 [I5]3.1 × 100

2.0 × 10−3

3.6 × 10−1

7.9 × 104 [I5]Co

7.6 × 101 [I5]2.3 × 100

1.9 × 10−3

5.4 × 10−1

4.4 × 104 [I5]60

Co

7.6 × 101 [I5]2.3 × 100

1.9 × 10−3

5.4 × 10−1

4.4 × 104 [I5]65Zn

3.4 × 103 [I5]4.9 × 10−2

4.0 × 10−4

5.4 × 10−1

5.0 × 102 [I2]90Sr

2.9 × 100 [I5]1.2 × 10−1

7.5 × 10−4

8.1 × 10−1

1.2 × 103 [I5]

106

Ru

5.5 × 101 [I5]2.0 × 100

1.9 × 10−3

5.4 × 10−1

3.2 × 104 [I5]129I

3.0 × 101 [I5]4.0 × 10−1

1.4 × 10−3

8.1 × 10−1

4.4 × 103 [I5]131I

3.0 × 101 [I5]4.0 × 10−1

1.4 × 10−3

8.1 × 10−1

4.4 × 103 [I5]

134

Cs

2.5 × 103 [I5]1.8 × 100

1.9 × 10−3

8.1 × 10−1

2.9 × 104 [I5]

137

Cs

2.5 × 103 [I5]1.8 × 100

1.9 × 10−3

8.1 × 10−1

2.9 × 104 [I5]

210

Pb

2.5 × 101 [I5]8.3 × 10−1

1.7 × 10−3

5.4 × 10−1

1.0 × 104 [K1]

212

Pb

2.5 × 101 [I5]6.3 × 10−1

1.5 × 10−3

5.4 × 10−1

1.0 × 104 [K1]

214

Pb

2.5 × 101 [I5]1.3 × 10−3

2.1 × 10−4

5.4 × 10−1

1.0 × 104 [K1]

210

Po

3.6 × 101 [I5]2.2 × 100

1.9 × 10−3

5.4 × 10−1

4.0 × 104 [S3]

226

Ra

4.0 × 100 [I5]6.4 × 10−1

1.6 × 10−3

5.4 × 10−1

7.4 × 103 [I5]

228

Th

6.0 × 100 [I5]4.0 × 100

2.0 × 10−3

3.0 × 10−1

1.9 × 105 [I5]

230

Th

6.0 × 100 [I5]4.0 × 100

2.0 × 10−3

3.0 × 10−1

1.9 × 105 [I5]

232

Th

6.0 × 100 [I5]4.0 × 100

2.0 × 10−3

3.0 × 10−1

1.9 × 105 [I5]

234

Th

6.0 × 100 [I5]3.9 × 100

2.0 × 10−3

3.0 × 10−1

1.9 × 105 [I5]234U

9.6 × 10−1 [I5]5.0 × 10−3

4.9 × 10−5

3.0 × 10−1

5.0 × 101 [I2]238U

9.6 × 10−1 [I5]5.0 × 10−3

4.9 × 10−5

3.0 × 10−1

5.0 × 101 [I2]ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 53

14

Water treatment factor
FWT
[B7]a

a

Concentration factor for freshwater fish,
Bfish (L/kg) (fresh weight)

Activity concentrations in sediment (dry weight) in the
100th year of continuous discharge at 1 Bq/s (Bq/kg)
Small river[S8]

Large river[S8]

Water treatment factor
FWT
[B7]a

Partition coefficient, Kd(m3/t)

239

Pu

3.0 × 101 [I2]4.1 × 100

2.0 × 10−3

2.7 × 10−1

2.4 × 105 [I5]

240

Pu

3.0 × 101 [I2]4.1 × 100

2.0 × 10−3

2.7 × 10−1

2.4 × 105 [I5]

241

Am

2.4 × 102 [I5]3.5 × 100

2.0 × 10−3

2.7 × 10−1

1.2 × 105 [I5]

The water treatment factors used are based on flocculation, coagulation, clarification and rapid sand filtration. Where data were not given for a specific element in Brown et al. [B7], an analogue approach was used.

54 UNSCEAR 2016 REPORTRadionuclide

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 55(b) Transfers from fresh water to terrestrial environments
77. The irrigation of crops with fresh water may lead to the transfer of radionuclides from a
freshwater to a terrestrial environment. This exposure pathway was not included in the Committee’s
previous methodology but, given the importance of irrigation of crops throughout the world, theCommittee decided that it should be considered. In general terms, the rate at which radionuclides
present in fresh water are deposited on crops and soil is estimated from information on the rate ofirrigation and the length of the crop-growing season. This information is then used to estimate theresultant activity concentrations of radionuclides in terrestrial crops from their activity concentrations infresh water and the derived calculation factors for their activity concentrations in food per unit
deposition rate.
78. There are many different types of irrigation. The most important for the transfer of radionuclides
to crops is sprinkler or spray irrigation because radionuclides in the water will be deposited directly on
the surface of the plants. Other types of irrigation where water is deposited on the surface of the soil
give rise to lower transfers and so were not included in the Committee’s methodology. Although there
are obvious differences in the behaviour of radionuclides that are deposited on the surface of plants
during sprinkler irrigation and that of radionuclides deposited by wet and dry deposition from theatmosphere, the Committee considered that the differences in the overall transfers to plants are asecond-order effect in the context of the objectives of the methodology. Therefore, the values of thetransfer factors for deposition from the atmosphere are assumed to apply also to irrigation. These
parameters include the transfers from plant surfaces and from uptake from the soil. The Committee
recognizes that there may be some overestimation in assuming that the rate of uptake from irrigationwater is equivalent to that from deposition from the atmosphere, given that greater losses from plantsurfaces may be anticipated for the higher deposition rates characteristic of irrigation.
79. The Food and Agriculture Organization of the United Nations (FAO) publishes statistics on water
resources in its AQUASTAT database [F4]. The information indicates that 40% of global foodproduction involves irrigation. Data are available on the overall rates of abstraction of water by source(whether from surface or groundwater). Information is available on the overall rate of water abstracted
for irrigation, but not on the fraction of irrigation water that is drawn from each source. In the absenceof this information, it is assumed that the fraction of water used for irrigation of the total amountabstracted is the same for both sources. Using data from the AQUASTAT database, the fraction of
surface water withdrawn for irrigation and the area irrigated per unit volume of water was obtained foreach region as shown in table 11, which also gives world-average values. The appendix describes how
these values are used in the methodology.
80. From this table, it is clear that, with the exception of West Asia, there is only limited abstractionof water from surface water.
81. The International Commission on Irrigation and Drainage provides data on the area of land under
different types of irrigation for 45 countries [I7]. The Committee has used these data to derive thefraction of land under sprinkler/spray irrigation for each region and an average for the world (seetable 12). The values in this table are for various times between 1999 and 2012.
82. The fractions given in table 12 are used in the methodology together with the calculated activity
concentrations of radionuclides in water and the abstraction rate for irrigation to derive concentrationsof radionuclides in foods and hence to estimate individual and collective doses. The Committee agreed
that irrigation should only be considered for vegetables and grain and only for large rivers. For small

56 UNSCEAR 2016 REPORT

rivers with low abstraction rates, it is unlikely that there is significant irrigation. Although other crops
are likely to be irrigated, the most important are leafy vegetables and grain [F4]. The irrigation of
pasture is important in some parts of the world (such as parts of the United States) but it is rare in othermajor areas with significant cattle production, such as Brazil and Argentina, and so it is not included in
the methodology.
Table 11. The fraction of water for irrigation withdrawn from surface waters and the total area
irrigated per unit volume of surface water available
Parameter
RegionFraction of water for irrigation withdrawn
from surface waters (dimensionless)Area irrigated per unit volume ofwithdrawn water (m2 a /m3)

Africa2.7 × 10−2

6.0 × 10−1

Asia and Pacific9.4 × 10−2

7.9 × 10−1

Europe1.2 × 10−2

1.9 × 100

Latin America and Caribbean

1.0 × 10−2

8.2 × 10−1

North America

2.4 × 10−2

1.3 × 100

West Asia

5.7 × 10−1

4.7 × 10−1

World average

3.9 × 10−2

8.5 × 10−1

Table 12. Fraction of irrigated land that is spray-irrigated (based on information from [I7]
RegionFraction

Africa0.2

Asia and Pacific

0.1

Europe0.4

Latin America

0.2North America

0.5West Asia

0.3

World average

0.2(c) Specific activity models for tritium and carbon-14 in fresh water
83. The Committee has adopted a specific-activity approach for assessing the transfer of tritium andcarbon-14 in a freshwater environment. This is the same approach as it used previously and as
described for a terrestrial environment. The assumption that equilibrium conditions exist, which is
implicit in the specific-activity model, is considered to be a good approximation for most aquaticcompartments in the model [E1, K5]. The model adopted is that described in [I5] in which activity
concentrations of tritium in fish are determined on the basis of the HTO concentration in the watercolumn and the fractional water content of fish. The fractional water content is found to be 0.78 L/kg
for most fish that form part of the human diet [I5]. The concentration of organically bound tritium(OBT) is also taken into account. The relevant equations and parameter values used in the methodology
are given in the appendix.

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 57

84. A specific activity approach is also applied to estimate the total tritium concentration in soil waterby taking account of the activity concentration of tritium in irrigation water, the rate of irrigation andthe effect of mixing with water reaching the soil from precipitation [C1]. The equations and dataapplied are presented in the appendix.
85. For discharges to aquatic bodies, the specific activity of dissolved inorganic carbon is assumed to
be in equilibrium within the part of the environment of interest. The methodology includes a simplified
approach based on a dynamic model developed by Sheppard et al. [S4, S5] with some modification to
apply it to irrigation based on the approach outlined in [C1]. The equations and parameter values
applied are given in the appendix.

3.

External and internal exposure(a) External exposure from radionuclides in riverbank sediments
86. Annual individual doses from external exposure during occupancy of riverbanks are calculatedusing the time-integrated activity concentrations of radionuclides in sediments, the dose coefficients forexternal exposure from surface deposits and the amount of time spent on riverbanks. This exposure
pathway is only considered for the calculation of the characteristic individual dose.
87. For discharges to fresh water bodies, activity concentrations of radionuclides in sediments at a
distance of 5 km downstream from the discharge point are used for the purpose of calculating
characteristic individual doses from external exposure due to radionuclides in riverbank sediments. It is
assumed that there will be continual cycling of radionuclides from the aquatic environment intoriverbank sediments taking account of build-up over the 100 year integration period. Dose coefficientsfor external radiation exposure from surface deposits (Sv/s per Bq/m2) from the United States Federal
Guidance Report No. 12 [E2] are used in the Committee’s revised methodology. However, these dosecoefficients apply to an infinite surface (assuming a surface deposit only); for river banks, a geometry
factor is applied to allow for the finite size of the source of external exposure. Apostoaei et al. [A2]have discussed appropriate geometry factors (these depend upon the surface area of the sediment on theriver bank, as represented by the width of the sediment on the bank, and upon the radionuclide). They
recommend that factors for 137Cs are applicable in general. The Committee agreed to apply a geometry
factor of 0.2 based on the values given by Apostoaei et al. for riverbank sediments with a width of afew metres. In addition, the Committee agreed that the methodology should use an annual occupancy
factor of 50 hours for the calculation of the characteristic individual dose.

(b) Internal exposure from ingestion
88. Three sources of internal exposure via ingestion are considered in the methodology: drinking
water; freshwater fish and irrigated terrestrial foods. In estimating the characteristic individual doses, it
is assumed that only 25% of the food is locally produced (see section II, paragraph 11).

58 UNSCEAR 2016 REPORTDrinking water
89. As described above, the activity concentrations of radionuclides in unfiltered fresh water are
assumed to be a function of the volumetric flow rate of the river, which is, in turn, directly related to thetype of river assumed. In order to derive activity concentrations in drinking water, element-specificwater treatment factors are applied. These are similar to those presented in the UNSCEAR 2000 Report[U6], updated with information included in the most recent edition of the WHO Drinking Water
Guidelines [W3]. The WHO provides indicators of the performance of a number of common watertreatment methods for a range of radionuclides.
90. Different treatment techniques are applied across the world and have different effectiveness.
Based on information used for the WHO Drinking Water Guidelines [W3], the Committee assumed
that—for most countries with high human development indices—flocculation, coagulation, clarification
and rapid sand filtration are used. As there are insufficient data available on the techniques employed inless developed countries, the Committee decided to apply a single set of water-treatment factors basedon the data for countries with high human development indices. The fractions of each radionuclide
removed by standard water-treatment processes assumed in the methodology are shown in table 10.
91. Previously, the Committee used a value of 500 litres per year (1.4 litres per day) [U6, U7, U9] forthe worldwide average individual rate of ingestion of drinking water. The Committee has reconfirmedthat this value should be retained because it is appropriate for average individual consumption rather
than the somewhat higher values typically used for protection purposes.
Freshwater fish
92. Individual doses from the ingestion of freshwater fish are calculated using the activity
concentrations of radionuclides in fish and the per caput consumption rate; collective doses arecalculated using the activity concentrations and fish-catch data. Information on fish-catch data for eachregion in 2007 was obtained from [F3]. The total fish catch for each region was divided by the size ofthe population for the same region to derive per caput consumption rates, which were used to estimatecharacteristic individual doses. Table 13 shows the resulting annual per caput consumption rates offreshwater fish for each region and a world-average value.
Table 13. Annual per caput consumption of freshwater fish by region
These values represent the total intakes of freshwater fish; the factor of 25% to account for local consumption of food isapplied subsequently when calculating characteristic individual doses

RegionAnnual consumption (kg)Africa2.7

Asia and Pacific

7.8

Europe3.4

Latin America and Caribbean

1.6

North America

4.5West Asia

1.3

World average

5.7

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 59

93. For the calculation of collective doses, the Committee has derived generalized fish-catch
information for a range of rivers using data from FAO [F4]. Aquaculture (e.g. the intensive farming of
freshwater fish) is excluded from consideration because a large proportion of fish reared in aquaculture
come from lakes and ponds [F4].
94. Typical values for the quantity of freshwater fish caught per unit distance and volume of water fortypical small and large rivers have been derived for use in the Committee’s methodology for eachregion separately and as a world average. From the FAO data, information was derived on the catch offreshwater fish including and excluding that obtained from aquaculture, plus the annual fish catch foreach river in a country. Using information on the length and volume of the relevant stretch of eachriver, annual fish catches per unit length and per unit volume were calculated and the results are shownin table 14. Based on this information, generic values for small and large rivers were adopted for use in
the methodology. It is therefore assumed that one tonne of freshwater fish is caught annually perkilometre from small rivers (of length less than 500 km) and that ten tonnes are caught annually perkilometre from large rivers (length greater than 500 km). Because of the volumes of water assumed forthe two river sizes, which also differ by a factor of 10, the freshwater catch per unit volume isindependent of river size.

60 UNSCEAR 2016 REPORT

Table 14. Freshwater fish-catch data for a range of rivers
Data taken from FAO [F1] unless otherwise stated

CountryCountry-wide annual fish catchRiver

Annual fish catchfor the river(s)Excludingaquaculture (t)RiverAnnual fish catch
Per unit length, excluding
aquaculture (t/km)

Includingaquaculture (t)Excludingaquaculture (t)Australia

4.0 × 103

1.2 × 103

All rivers

China

1.9 × 107

1.6 × 106Pearl

2.8 × 104

1.4 × 103

Egypta9.4 × 105

2.3 × 105

Nile

2.3 × 105

1.5 × 103

France4.4 × 104

2.6 × 103

Loire

1.0 × 103

5.6 × 102

4.3 × 108

2.0 × 10−6

1.9 × 100Rhône

1.5 × 103

4.0 × 102

6.3 × 108

2.0 × 10−6

3.9 × 100

All rivers

2.6 × 103
7.9 × 103

5.6 × 104

2.1 × 104

DanubeJapan

6.7 × 104

2.5 × 104

All rivers

Romania

1.6 × 104

5.7 × 103Danube5.7 × 103

Sudanb

8.1 × 104

7.9 × 104

Nile

5.4 × 104

Nile tributaries(incl. Blue andWhite Nile)

2.5 × 104

ThamesUnited States

1.6 × 104

2.5 × 105

2.5 × 103

1.4 × 104

Volume (m3)Per unit volume, excluding
aquaculture (t/m3)

6.2 × 10−1

Germany

United Kingdom

Length in specified
country (km)

2.3 × 109

1.0 × 10−5

2.0 × 101
1.5 × 102

3.0 × 10-1
5.0 × 102

2.1 × 108

4.0 × 10−5

1.6 × 101
1.4 × 101

4.0 × 102

3.2 × 109

2.0 × 10−6

1.4 × 101

2.3 × 102

2.3 × 102

2.5 × 107

9.0 × 10−6

9.8 × 10−1

Kennet (tributaryof Thames)

6.0 × 100

3.3 × 101

6.1 × 105

9.0 × 10−6

1.7 × 10−1

All rivers

2.5 × 103

All riversaAdditional data obtained from Hamza [H1].

bData for 2012 from internal report published by the Sudan Ministry of Animal Resources and Fisheries [S11].

8.0 × 10−1
3.4 × 10−1

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 61

Irrigated foods
95. As discussed in section III.B.2(b), the ingestion of irrigated foods is included in the methodology
for the estimation of both characteristic individual and collective doses. In the calculation of individual
doses from this pathway, it is assumed that cereals and vegetables are irrigated. For the assessment ofcollective doses, it is assumed that only cereals are irrigated, because these represent the bulk of
irrigated crops intended for human consumption [F4].
96. As illustrated in table 15 [C1], there are variations in irrigation rate which are likely to depend tosome extent upon factors such as the type of crop and the climate. A single value for a daily irrigation
rate was considered to be appropriate for the purposes of this methodology. The United States NCRP in
its screening methodology uses a value of 5 L/m2 for the volume of irrigation water applied per unit
area for a period of 150 days annually[N2]. The FAO AQUASTAT database [F4] contains data for awide range of countries on irrigation water withdrawals and/or requirements, and on the total areas ofirrigated crops harvested. These data could then be used to indicate values of daily irrigation rates in
L/m2 assuming a total irrigation period of 150 days in a year. Table 16 illustrates the range of irrigation
rates thus obtained from the FAO data. The irrigation rate varies from country to country and althoughsome of the highest values are for arid areas of the world, some countries that are generally arid (e.g.
Peru) do not have particularly high estimated irrigation rates. From the data in tables 15 and 16 a single
value for the daily irrigation rate of 4 L/m2 was adopted for the methodology.
Table 15. Irrigation rates averaged over the growing season for various crops and climates [C1]Country or area

Crop type

Daily irrigation rate (L/m2)
Min

CanadaFrance(Loire Valley)

India

Republic of Korea

ForageAverage

Reference
Max

0

[C6]

Garden vegetables

0

1.8

Fruit

2

6

Garden vegetables

2

6

Grain

0

2.4

Maize

0.8

2.8

Banana

2.6

Gram

1.8

Ground nut

6.3

Improved jowar

2.0

Pigeon pea

1.6

Rice

0.9

Sugar cane

4.1

Turmeric

2.1

Wheat

4.3

Rice

5

[C4]

[P1]

8

[J4]

62 UNSCEAR 2016 REPORTCountry or area

Crop type

Daily irrigation rate (L/m2)
Min

United States(California)

Average

ReferenceMax

Alfalfa

4.9

7.1

Barley2.4

3.2

Fruit orchards

3.9

4.9

Garden vegetables

4.2

5.2

Hay

3.2

5.4

Oats, rye

2.0

2.5

Pasture3.7

4.4

Grapes

1.7

[U14]

[B5]

Table 16. Effective irrigation rates obtained from data in the FAO AQUASTAT database [F4]Country

Daily irrigation rate (L/m2)

Algeria

5.2

Argentina

3.5

Australia

3.3

Bangladesh

2.8

Brazil1.9

Canada

4.1

China

1.8

Egypt

5

France1.4

Greece

2.8

India

2.8

Indonesia

2.3

Japan

8.6

Kenya

3.2

Peru

3.5

Qatar

7.3

Russian Federation

6.5

Spain

4.5

United States

5.1

97. Collective doses from irrigation are determined by assuming that, on average, the fraction ofwater abstracted for irrigation from the type of river of interest, in a given geographical region, is thesame as the fraction of water abstracted for irrigation purposes from all surface waters (rivers, lakes andso on). The fraction of water abstracted for irrigation and the amount of land irrigated per unit volume
of water, was derived from information in the FAO AQUASTAT database [F4]. This information,
together with the annual yields of cereals [F1] (given in table 17), the integrated activity concentrationsof radionuclides in cereals for unit deposition rate (given in table 5), and the percentage of irrigation

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 63

that is sprinkler irrigation (see table 12 and section III.B.2(b)), is used to calculate collective doses from
irrigation. Table 17 shows the parameter values used in the methodology for the estimation of collective
doses in different regions from the irrigation of cereals; world-average values are also given.
Table 17. Parameter values used in the model for the irrigation of cereals
RegionAnnual cereal yield(kg/m2)aFraction of surface waterused for irrigationb
Firr,surface

Area irrigated per unit
volume of water
withdrawnb
Airr,unit (m2 a/m3)

Africa2.3 × 10−1

2.7 × 10−2

6.0 × 10−1

Asia and Pacific

2.9 × 10−1

9.4 × 10−2

7.9 × 10−1

Europe

3.6 × 10−1

1.2 × 10−2

1.9 × 100

Latin America and Caribbean

3.3 × 10−1

1.0 × 10−2

8.2 × 10−1

North America

3.0 × 10−1

2.4 × 10−2

1.3 × 100

West Asia

2.9 × 10−1

5.7 × 10−1

4.7 × 10−1

World average

3.0 × 10−1

3.9 × 10−2

8.5 × 10−1aFrom FAO 2010 [F1].

bFrom FAO 2014a [F3].

4.

Individual and collective doses

98. Annual characteristic individual doses arising from the ingestion of drinking water and freshwater
fish are calculated using the derived activity concentrations of radionuclides in water and fish, therelevant consumption rates, and the dose coefficients for ingestion (see table 1). Because the activity
concentrations in unfiltered fresh water are assumed to be a function of the river flow rate alone, these
individual dose estimates are directly related to the type of river assumed.
99. Collective doses are calculated using the total abstraction rate of drinking water and the total catchof fish associated with the freshwater body. The abstraction rate of drinking water is assumed to be afraction of that abstracted for municipal use. The abstraction rate was derived from FAO statistics
which indicate that around 1–10% of domestic water is used for drinking [F1]. Since not all municipal
water will be for domestic use, the lower end of this range is applied in this methodology. The total
amount of water abstracted depends on the size of the water body with smaller amounts removed fromsmall rivers than from large rivers; the water flow is also inversely related to this size. This means that,
although for small rivers the activity concentrations per unit discharge are higher than for large rivers,
the amount of water abstracted is lower and therefore—using the Committee’s methodology—thecollective doses from drinking water are estimated to be essentially independent of river size.
100. The irrigation pathway is considered using generic irrigation rates and transfer factors appropriate
for a terrestrial environment. Characteristic individual doses from this pathway are determined using
region-specific population-weighted consumption rates and the assumption that 25% of the food islocally produced. Collective doses are determined from information on the region-specific proportionof water abstracted from surface waters for irrigation purposes, the area irrigated and the annual yieldof cereal products.

64 UNSCEAR 2016 REPORT101. Collective doses from the consumption of water and freshwater fish are calculated on the basis of
the total abstraction of water and fish-catch data, respectively, and are thus related to the river as awhole and cannot be divided into local and regional components.

C. Assessment of doses from discharges to a marine environment
102. As indicated above, the Committee decided to develop generic dose calculation factors that takeaccount of dispersion within different types of water bodies. The approach used for assessing doses
from discharges into a marine environment is based on a simple two-box compartment model, thecharacteristics of which are broadly representative of areas where discharges into a coastal environment
occur (see figure IV, which is based on [C2]). The larger compartment (referred to as the regional box)
could also be used to assess the dispersion of radionuclides discharged into the deep sea from oil andgas platforms, if required.
103. The following exposure pathways are considered for discharges to a marine environment (seefigure I):(a) External exposure from radionuclides (beta and gamma emitters) in sediments;(b) Internal exposure from ingestion of radionuclides incorporated into marine foods.
The first pathway is considered for the calculation of characteristic individual doses only; a review ofprevious studies assessing collective dose indicated that this pathway provided less than 1% of thecollective dose. Similar to discharges to atmosphere and freshwater environments, other exposure
pathways (e.g. inadvertent ingestion of water or sediments, and exposures during swimming) are
possible; however, the ingestion of marine foods has been found to be the most important exposure
pathway in published assessments of collective doses (for example, see [J5]).
Figure IV. Simple two-box compartment model representing a marine environment
Q is the discharge rate (Bq/a); λl,r is the effective rate of transfer between the local and regional compartments (a−1), taking
into account exchange between the compartments; λr,g is the effective rate of transfer between the regional and global
oceans (a−1), taking into account exchange between the compartments; λs,l and λs,r are the rates of transfer to sediment for
the local and regional compartments, respectively (a−1), and li is the radioactive decay constant for radionuclide, i, (a−1)

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 65

104. Figure IV illustrates the simplified approach used to model the movements of radionuclides between
marine compartments. The model considers inputs (such as discharges and incoming transfers ofradionuclides) and losses (such as radioactive decay, sedimentation and outgoing transfers ofradionuclides). This allows the activity concentrations in the local and regional compartments to be
determined. The equations and parameter values associated with this approach are presented in the
appendix. The model includes transfer to other marine areas but any contribution to collective doses fromthe global oceans is only considered for the long-lived radionuclides (3H, 14C and 129I) in a separate global
circulation model as discussed below. The size of the regional compartment and the related transfer
parameters have been chosen to ensure that the model is a reasonable representation. A comparison with amore complex multi-compartmental marine model [S7, S8], which has been validated against
measurement data, has shown that the simple model is adequate for the purpose of this methodology.
105. The dimensions and characteristics applied in the methodology are presented in table 18.
Table 18. Characteristics of the marine compartment model
Marine compartment

Characteristic
Volume (m )
3

Local

Regional

1 × 101 × 1015

9Effective rate of transfer between compartments (a−1):
Local to regional, λ l,r2 × 101

Regional to global, λr, g1

Depth (m)

10

1 000

Suspended sediment load, SSL (t/m3)

2 × 10−4

1 × 10−7

Sedimentation rate, SR (t m−2 a−1)

1 × 10−4

1 × 10−5

10

1 000

Length of coastline (km)OTHER PARAMETERS
Volume of global oceans, Vg (m3)

1 × 1018

Volumetric exchange between local and regional compartments,∆Vl,r (m3/a)

2 × 1010

Volumetric exchange between regional compartment and global oceans,∆Vr,g (m3/a)

1 × 1015

106. The transfer of radionuclides between water and sediments is dependent on a combination offactors that relate to the characteristics of the water body and of individual radionuclides. The key
radionuclide-independent parameters are the suspended sediment load and sedimentation rate, which
are determined by the nature of the coastal or marine compartment. Values have been chosen from therange of those associated with European waters [S7]. Sedimentation is also dependent upon theradionuclide-dependent partition factor (Kd) for a coastal or marine environment; these values havebeen derived from [I3].
107. The purpose of the local marine compartment is to allow doses to a characteristic individual livingin the area and collective doses to the local population to be calculated. The selected dimensions of thelocal marine compartment allow tidal, bathymetric and sedimentary conditions to be broadly
homogenous throughout the compartment and have been selected so that the compartment issufficiently large to represent a source of shellfish for the local population. The dimensions of the local

66 UNSCEAR 2016 REPORT

marine compartment are typical of those used in the MARINA II project for European waters [S7]. The
dimensions of the regional marine compartment are significantly larger than those generally adopted forsite-specific modelling: approximately equivalent to the size of the north-east Atlantic. This again
reflects the purpose of the methodology; the regional marine compartment is intended to represent thesource of exposure for the regional component of collective dose, which is assumed to extend for a
distance of around 1,500 km from the point of discharge.
108. The depth of the sea in the regional marine compartment is assumed to be 1,000 m, which is thedepth down to which fish are assumed to be caught.
109. The sensitivity of the calculated activity concentrations of radionuclides in water to changes in thevolume of the local compartment and movements between the local and regional compartments wasinvestigated. The calculated activity concentrations in water in the local compartment (assuming
constant transfer rates between the local and regional compartments) decrease proportionally as thevolume of the local compartment increases. A water transfer rate of 20–40 a−1 is typical of those foundin European waters based on the volume exchange rates and water volumes given in [S8]. The
Committee adopted a volume of 1 × 109 m3 with a water transfer rate of 20 a−1 for the local marine
compartment as being reasonably representative of coastal conditions into which nuclear sites are
known to discharge.
110. The values of the sedimentation rate and suspended sediment load adopted for the local
compartment are typical of values for local compartments around the United Kingdom (e.g. the southwestern part of the North Sea) [S7]. The values adopted for the regional compartment correspond to thenorth-east Atlantic [S7], considered to be typical of large deep oceans.

1.

Radionuclide transfers in a marine environment

111. The key element-specific parameters that take account of the transfer of radionuclides from waterto other parts of a marine environment are (a) concentration factors, which relate concentrations ofelements in fish and shellfish to their concentrations in water, and (b) partition coefficients, which
express the distribution of elements between sediment and water. The most significant international
compilation of concentration factors (also referred to as bioaccumulation factors) and partition
coefficients, Kd, for coastal and deep-sea environments is presented in [I3]. Following a critical reviewof the data, these values have been adopted and applied in the present methodology, and are given intable 19.

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 67

Table 19. Radionuclide-specific parameters used in the marine model[I3]

Dose coefficient forexternal exposure
from surface
deposita,
Dex,deposit(Sv/s per Bq/m2) [E2]H

0

1 × 100

1 × 101 × 100

0

1 × 100

1 × 100

14

C

2 × 104

2 × 104

2 × 104

1.3 × 10−20

1 × 103

2 × 10335

S

1 × 100

1 × 100

3 × 100

1.3 × 10−20

5 × 10−1

1 × 100

54

Mn

1 × 103

5 × 103

5 × 104

7.9 × 10−16

2 × 106

2 × 108

58

Co

7 × 102

7 × 103

2 × 104

9.3 × 10−16

3 × 105

5 × 10760

Co

7 × 102

7 × 103

2 × 104

2.3 × 10−15

3 × 105

5 × 107

65

Zn

1 × 103

3 × 105

8 × 104

5.4 × 10−16

7 × 104

2 × 105

90

Sr+ 90Y

3 × 100

5 × 100

1 × 101

1.6 × 10−18

8 × 100

2 × 102106

Ru + 106Rh

2 × 100

1 × 102

5 × 102

3.5 × 10−16

4 × 104

1 × 103

129

I

9 × 100

3 × 100

1 × 101

2.0 × 10−17

7 × 101

2 × 102131

I

9 × 100

3 × 100

1 × 101

3.6 × 10−16

7 × 101

2 × 102

134

Cs

1 × 102

5 × 101

6 × 101

1.5 × 10−15

4 × 103

2 × 103137

Cs+ 137mBa

1 × 102

5 × 101

6 × 101

5.8 × 10−16

4 × 103

2 × 103210

Pb

2 × 102

9 × 104

5 × 104

2.1 × 10−18

1 × 105

1 × 107212

Pb

2 × 102

9 × 104

5 × 104

1.4 × 10−16

1 × 105

1 × 107

214

Pb

2 × 102

9 × 104

5 × 104

2.4 × 10−16

1 × 105

1 × 107

210

Po

2 × 103

2 × 104

2 × 104

8.1 × 10−21

2 × 107

2 × 107

226

Ra

1 × 1021 × 1021 × 102

6.1 × 10−18

2 × 103

4 × 103228

Ra

1 × 1021 × 1021 × 1020

2 × 103

4 × 103

228

Th

6 × 1021 × 103

1 × 103

2.1 × 10−18

3 × 106

5 × 106

230

Th

6 × 1021 × 103

1 × 103

6.4 × 10−19

3 × 106

5 × 106

232

Th

6 × 1021 × 103

1 × 103

4.6 × 10−19

3 × 106

5 × 106

234

Th

6 × 1021 × 103

1 × 103

7.5 × 10−18

3 × 106

5 × 106

234

U

1 × 100

1 × 101

3 × 101

5.9 × 10−19

1 × 103

5 × 102

238

U

1 × 100

1 × 101

3 × 101

4.2 × 10−19

1 × 103

5 × 102239

Pu

1 × 1022 × 102

3 × 103

2.8 × 10−19

1 × 105

1 × 105

240Pu

1 × 1022 × 102

3 × 103

6.0 × 10−19

1 × 105

1 × 105

241

Am

1 × 102

4 × 1021 × 103

2.3 × 10−17

2 × 106

2 × 106

Concentration factor (L/kg) (fresh weight)
Radionuclide

Fish
Bfish
[I3]

3

a

Crustaceans
BcrustMolluscs
Bmolluscs[I3]

Partition coefficient (L/kg)(dry weight sediment)
Ocean
marginKd,local [I3]Open ocean
Kd,regional[I3]

These dose coefficients are for the parent only; the contribution from short-lived progeny is added separately (see appendix).

68 UNSCEAR 2016 REPORT

2.

External and internal exposure(a) External exposure from radionuclides in beach sediments
112. Annual characteristic individual doses due to external exposure from occupancy of beaches are
estimated on the basis of the modelled activity concentrations of each radionuclide in sediments, therelevant dose coefficients for external exposure from surface deposits and the amount of time spent on
the beach.
113. The activity concentration of each radionuclide in sediments is estimated using the activity
concentration in water, the relevant partition coefficient Kd, an assumed average thickness of sedimentof 5 cm and an assumed density of sediment of 1.2 t/m3 [I2]. The dose coefficients for external
exposure from surface deposits are taken from [E2] for each radionuclide. A factor of 0.5 is applied tothe dose coefficients to account for the geometry of the radionuclide distribution on a marine shoreline[E2]. The estimates of effective dose include contributions from beta and gamma irradiation of the skinfrom radionuclides in the sediment. Any additional dose from irradiation of the skin due to directcontact with sediment is not included in the methodology because it is deemed not a major contributor
to the overall characteristic individual doses.
114. An average individual beach occupancy rate of around four days per year is assumed, based ondata for the Nord-Cotentin area of France [R1]. Because this exposure pathway is unlikely to be asignificant contributor to collective doses from nuclear installations [J6], it is not included in theestimation of collective dose.

(b) Ingestion of marine foods
115. Annual individual doses from ingestion of radionuclides in marine foods are estimated from themodelled activity concentration of each radionuclide in fish and shellfish (comprising crustaceans andmolluscs) harvested from the relevant marine compartment, the region-specific or world-average annual
consumption of the food (for adults) and the relevant dose coefficient for ingestion of eachradionuclide. Information on the annual consumption of marine fish, crustaceans and molluscs was
obtained from WHO [W2] and the values used are given in table 20.
Table 20. Per caput annual consumption of marine foods by region
Based on data provided by WHO [W2]Regiona

Per caput annual consumption of food (kg)Fish

Crustaceans

Molluscs

Africa

6.6

0.1

0a

Asia and Pacific

6.9

1.4

2.4

Europe

13

0.9

1.0

Latin America and Caribbean

5.9

0.6

0.4

North America

8.2

2.8

1.4

West Asia

4.5

0.3

0a

World average

7.5

1.1

1.6

Consumption rate is less than 2× 10−2 kg/a and treated as 0.

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 69

116. Activity concentrations of radionuclides in aquatic foods are estimated from the product of themodelled activity concentrations in water and concentration factors that relate the activity
concentrations in sea food to those in water (see table 19). In order to apply this approach, it wasnecessary to make some assumptions about the origin of the food and therefore the activity
concentrations in water that are appropriate to adopt. Thus, it was necessary to make an assumption
about the proportions of fish and shellfish that are derived from each compartment in the model.
117. One approach applied to estimate doses to individuals is to assume that a certain proportion offish and shellfish is derived from the local and regional marine compartments. The proportions assumed
would depend upon the purpose of the assessment. In a similar manner to that used for terrestrial andfreshwater foods, it is assumed in the Committee’s methodology that the characteristic individual
obtains 25% of their fish consumption from the local compartment, with the remaining fraction
obtained from the regional compartment. All of the crustaceans and molluscs are assumed to beobtained from the local compartment.
118. The respective local and regional components of collective dose from marine discharges are
estimated using (a) data on the average crustacean-catch and mollusc-catch per unit length of coastline
appropriate for the local compartment, and (b) data on the average fish-catch per unit volume
appropriate for the regional compartment. Catch data for fish, crustaceans and molluscs were derived
from the FAO FishStatJ software [F2], which comprises the FAO databases on capture and aquaculture
for major fishing areas from 1955 to 2012. The area of sea and length of coastline associated with eachof the major fishing areas were determined using a geographical information system (ArcGIS,
version 10) and then used to estimate the relevant values for the UNEP geographical regions employedin the methodology; world-average values were also derived. The edible fraction of the catch of fish,
crustaceans and molluscs is assumed to be 0.5, 0.35 and 0.15, respectively [S6]. Table 21 shows thecatch data for estimating collective doses for the different regions.
Table 21. Annual catch of fish, crustaceans and molluscs used for estimating collective doses fromdischarges to marine environments [F2]Annual catch
Mass of fish
per unit area of sea(kg/km2)

Mass of crustaceans
per unit length of coastline(kg/km)Mass of molluscs
per unit length of coastline(kg/km)Africa2 × 102

4 × 103

5 × 103

Asia and Pacific

2 × 1021 × 104

5 × 104

Europe2 × 1021 × 103

5 × 103

Latin America and Caribbean

9 × 101

7 × 103

1 × 104

North America

6 × 101

7 × 103

7 × 103

West Asia

2 × 102

5 × 103

5 × 103

World average

1 × 102

8 × 103

3 × 104

Region

70 UNSCEAR 2016 REPORT(c) Individual and collective doses
119. Annual individual doses arising from the ingestion of marine fish and shellfish are estimatedusing the modelled activity concentrations of radionuclides in water and fish, the per caput annual
consumptions given in table 20 and the dose coefficients for ingestion of the relevant radionuclides
given in table 1.
120. To estimate collective doses, region-specific information on the catch of fish and shellfish per unit
area of sea or length of coastline are applied from table 21.

IV. GLOBALLY DISPERSED RADIONUCLIDES
121. The Committee has periodically assessed public exposure to long-lived globally dispersedradionuclides discharged from nuclear power and reprocessing plants since 1982. The radionuclides ofparticular interest are 3H, 14C, 85Kr and 129I. The features of the approaches used and the evolution of theCommittee’s approach are summarized in electronic attachment 2. The estimated collective doses per unit
discharge integrated to various times after discharge are summarized in table 22 at the end of this section.

A. Tritium
122. The Committee’s previous approach to estimating the doses from the global circulation of tritium
discharged from nuclear installations [U6] was based on a comparison of models developed by Kelly et
al. [K2], NCRP [N1], Bergman et al. [B2] and Killough and Kocher [K4]. The relevant concentration oftritium for each compartment of the Committee’s model was determined from the total amount oftritium in the compartment divided by the volume of water represented in the compartment. The
concentration in humans was then estimated from the modelled concentration of tritium and therelevant fractional intake from each compartment of the model.
123. The global collective dose from discharges to the near-surface atmosphere within the 30°–50°
latitude band of the northern hemisphere was estimated by the Committee using the model developedby Killough and Kocher [K4]. This gave a global collective dose per unit discharge of tritium of
2.3 man Sv/PBq [U6]. The global collective dose arising from discharges to the ocean estimated by
NCRP [N1] and Bergman et al. [B2] was around one tenth of that arising from discharges toatmosphere. A value of 0.2 man Sv/PBq was therefore adopted in the UNSCEAR 2000 Report [U6] forthe collective dose per unit discharge of tritium to the ocean. The Committee still considers theseestimates to be appropriate for a world population of 10 billion.
124. The approach used by the Committee to estimate the collective dose commitment is summarized
as follows [S8]:S (t ) =∑ c ∑ p I c (t ) ⋅ f p ,c ⋅ Rc ⋅ U p ⋅ P(15)

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 71

where S(t) is the collective dose commitment truncated at time t (man Sv); Ic(t) is the time-integrated
activity concentration in compartment c at time t ((Bq a)/kg); fp,c is the fraction of the individual’s
intake of water due to pathway p; Rc is the total consumption rate by an individual of water fromcompartment c (kg/a); Up is the dose per unit intake from pathway p (Sv/Bq); and P is the number ofpeople in the exposed population. This model was implemented in the PC-CREAM 08 computer
system [S8] and has been used to derive values for the global collective doses from unit discharges oftritium for the Committee’s methodology. An important factor in the collective dose estimation is theassumed number of people in the world population. The estimate of the United Nations for this for
1 December 2014 was 6.4 billion (6.4 × 109) [U2]. The value for the global population of 10 billion
used in the Committee’s methodology is equivalent to the United Nations’ median estimate of the
projected population for 2060 [U1] and is deemed reasonable as a rounded value for collective doses
integrated into the future.

B. Carbon-14
125. Carbon-14 is the largest contributor to the collective dose from global dispersion of long-lived
radionuclides discharged from reprocessing of nuclear fuel and is a significant contributor to that resulting
from operation of nuclear reactors. A 23-compartment model [T1] (see figure VI, annex A of the
UNSCEAR 2000 Report [U6]) was applied by the Committee to estimate the activity per unit mass ofcarbon in each environmental compartment over time. Once mixing had been achieved, a specific-activity
approach was used to estimate the collective dose commitments from 14C, assuming that the specificactivity of 14C in the carbon ingested by humans was the same as that in the compartments that directly
related to the intake of food (i.e. ground vegetation for terrestrial foods and surface water compartments
for marine foods). The collective dose commitment per unit discharge to atmosphere, truncated at
10,000 years, was estimated to be 109,000 man Sv/PBq. The model can be used to estimate collective
doses for discharges to any compartment; the collective doses from unit discharge to the surface waters ofoceans were found to be about the same as those from unit discharges to atmosphere, but doses from unit
discharges to deep oceans were around 20% lower [U6]. Killough and Rohwer [K3] found that theestimates from six models differed by a factor of only 1.5, suggesting a remarkable level of agreement(although it has not been possible to fully validate any of the models against global measurements). Thishas been attributed to the long half-life of 14C relative to its rate of movement in the environment, whichmakes calculated dose commitments insensitive to the detailed structure of the models [K3].
126. The Committee has therefore not modified its approach to the estimation of collective dose
commitment from 14C from that used in 2000. The approach to derive the collective dose commitment
is implemented in PC-CREAM 08 and is summarized as follows [S8]:=S (t )∑∑ I (t ) ⋅ f
c

c

p

p ,c

⋅ Rc ⋅ U p ⋅ P(16)

where S(t) is the collective dose commitment truncated at time t (man Sv); Ic(t) is the time-integrated
activity concentration in compartment c at time t (Bq a/kg); fp,c is the fraction of the individual’s intake
of carbon from pathway p; Rc is the total intake rate of stable carbon (kg/a); Up is the dose per unit
intake from pathway p (Sv/Bq); and P is the size of the exposed population. As before for tritium, theworld population is assumed to be 10 billion.

72 UNSCEAR 2016 REPORT

C. Iodine-129
127. The Committee used the global circulation compartment model for 129I developed by Titley et al.
[T1] for the assessment it presented in the UNSCEAR 2000 Report (see figure VII, annex A [U6]). The
inventories of stable iodine in the compartments of the model and the transfers between thecompartments were estimated from environmental measurements and the requirement that the total
mass of iodine in the environment was balanced. Iodine intakes by humans from each modelcompartment were estimated from the average inhalation and food consumption rates combined withthe concentrations of stable iodine in the atmosphere and foods respectively, or using a specific-activity
approach. Five exposure pathways were considered as follows, with the values of individual intakes ofstable iodine used given in brackets: inhalation (0.29 µg/d); deposition from the atmosphere onto crops
followed by ingestion by humans or by dairy or beef cattle subsequently ingested by humans (6.6 µg/d);
ingestion of surface water (5.3 µg/d); ingestion of marine fish and shellfish (11 µg/d); and root uptake
from soils and surface waters followed by ingestion of crops and animal products (200 µg/d).
128. The PC-CREAM computer code [S8] also implements the model by Titley et al. The approach
used to estimate the collective dose is similar to that used for tritium and 14C, except that individual
intakes were estimated using transfer factors to relate the activity concentrations in food and air to theactivity concentrations in the various compartments, rather than to the intake of stable iodine. The
Committee still considers this approach to be appropriate.

D. Krypton-85
129. The Committee’s approach to estimating the global dispersion of 85Kr is described in detail in the
UNSCEAR 1988 Report [U5]. It was based on the approach presented in the UNSCEAR 1982 Report[U4], where a simple two-compartment model was used for the discharge of 85Kr [C2]. The twocompartments represent the tropospheres of the northern and southern hemispheres. A transfer
coefficient of 0.5 a−1 between the compartments was used. The time-integrated activity concentration in
air per unit discharge was 10−10 (Bq s)/m3 per Bq [U4]. The dose coefficients to convert from activity
concentration in air to absorbed dose rate in air and absorbed dose rate in skin from the emitted beta
radiation were derived from [C2].
130. The value of the collective effective dose equivalent per unit discharge used in the UNSCEAR
1982 Report [U4] was 0.17 man Sv/PBq (based on a world population in 1982 of 4 billion). This wasscaled to 0.2 man Sv/PBq for an assumed population in 1988 of 4.6 billion [U5]. It was noted that thiscollective dose commitment would be delivered during the first 50 years after discharge.
131. No changes have been made to the previous approach to modelling the dispersion of 85Kr. Around
50% of the effective dose arises from gamma irradiation of the whole body, and 50% arises from beta
irradiation of the skin [I8, Z1]. The value of dose rate for unit activity per mass of air is 8 × 10−9 Sv/aper Bq/kg (the value is given per unit mass of air because this expression is required for the global
circulation model and was obtained using a density of air of 1.225 kg/m3), based on an assumption of
immersion in a semi-infinite cloud (with no shielding) [S8]. The collective dose commitment is thenestimated as follows:

S (t )= I (t ) ⋅ F ⋅ P(17)

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 73

where S (t) is the collective dose commitment truncated at time t (man Sv); I(t) is the time-integrated
activity concentration in air of the appropriate hemisphere at time t ((Bq a)/kg); F is the dose rate perunit concentration in air (Sv/a per Bq/kg); and P is the number of people in the exposed population. As
for the other globally circulating radionuclides, the Committee has taken for its methodology values ofcollective doses derived by PC-CREAM 08, which implements the model used by the Committee anduses a world population of 10 billion.
Table 22. Collective dose commitments from globally dispersed radionuclides
Collective dose commitment truncated after a given timefrom a radionuclide discharge of 1 Bq/s over a year (man Sv)
Radionuclide3

To atmosphere

To a marine environment

100 years

500 years

10 000 years

100 years

500 years

10 000 years

H

1.0 × 10−8

1.0 × 10−8

1.0 × 10−8

1.1 × 10−9

1.1 × 10−9

1.1 × 10−9C

2.8 × 10−4

5.6 × 10−4

2.7 × 10−3

1.1 × 10−4

3.5 × 10−4

2.5 × 10−3Kr

8.0 × 10−9

8.0 × 10−9

8.0 × 10−9

—

—

—

I

2.7 × 10−3

3.1 × 10−3

7.4 × 10−3

3.3 × 10−6

6.7 × 10−6

1.4 × 10−4

14
85

129

V. LIMITATIONS OF THE MODELS AND DATA USED132. The methodology is generic and intended for use with discharges of radionuclides from nuclear
energy generating technologies, and a number of non-nuclear electrical energy production sources,
throughout the world. It builds on previous work that has provided the Committee with robust resultsthat have been suitable for its purposes. In developing the methodology further, no intentional bias has
been introduced to either under- or overestimate radiation exposures; the aim has been to be generic andas realistic as possible. It is difficult to quantify the uncertainties, because of the generic nature of themethodology. Nevertheless, this section outlines the limitations and uncertainties associated with themethodology. The Committee notes the following generic limitations:(a) The input to the models is assumed to be a continuous discharge, and annual average
parameter values are used. The models therefore do not apply for short-duration planned or
accidental releases of radionuclides to the environment;(b) Although some parameter values are used that are specific to geographical region (such as forfood consumption), most of the models are generic and are not intended for detailed site-specificdose assessments, risk assessments or demonstrating regulatory compliance;(c) The assumptions and data used for estimating individual doses are intended to apply to acharacteristic individual living local to the discharge point with typical habits and behaviour;(d) In order to be realistic rather than cautious in estimating characteristic individual doses, it isassumed that only 25% of food intake (both terrestrial and aquatic) is locally produced. The
resulting ingestion doses are sensitive to this assumption, being directly proportional to the foodintake, and the overall dose will be similarly sensitive where ingestion is a major component.
However, it should be noted that this assumption is not used in the assessment of collective doses;

74 UNSCEAR 2016 REPORT(e) The assessed collective doses are directly proportional to the assumed population densities forreleases to atmosphere or to the assumed total amount of drinking water or aquatic foods forreleases to water bodies. The methodology gives results for releases to atmosphere for four
different population distributions with values provided by geographical region for three of them.
The effect of using different distributions can be seen in the results presented in the following
section but are generally only relatively small. However, the use of a very low population density,
suitable for remote sites, leads to collective doses around two orders of magnitude lower than when
the default population densities are used. Therefore, the value for the very low population density
should be used with caution.
(f) The models and data are thought to be the most appropriate for use currently and are also
assumed to apply for representing the future. No account is taken of possible future changes, suchas to population distributions in different regions or the effects of climate change. The uncertainties
in the results of the methodology increase with time; this is particularly the case for global
circulation models of long-lived radionuclides when integrated for 10,000 years or more.
133. Specific limitations and uncertainties relating to the different areas of the methodology are notedin the following sections.

A. Discharges to atmosphere134. A standard Gaussian-plume model is used to estimate the dispersion of radionuclides following
discharges to atmosphere. As discussed in section III.A, the Gaussian model can be implemented
generically for assessing exposures from different sources of discharges throughout the world and it
produces results that agree reasonably well with measurements [C5]. However, as implemented in themethodology, the model makes no allowance for local topographical features (such as hills, buildings or
the site being on the coast) which can strongly influence dispersion, particularly close to the discharge
point. The height of the discharge, which is related to both the physical height of the discharge stack
and any buoyancy of the discharge caused, for example, by heat (referred to as plume rise), also has asignificant effect on the subsequent dispersion of the radionuclides. However, this effect reduces with
distance from the discharge point (see electronic attachment 4), and so the Committee agreed to adopt asingle value for stack height of 30 m for its generic methodology.
135. A simplified approach has also been retained to account for the deposition of radionuclides in theplume on the ground. An effective deposition velocity of 0.002 m/s is used to represent both dry andwet deposition for all radionuclides (apart from noble gases, tritium and carbon-14, for which, a valueof zero is used because noble gases do not deposit, and because deposition is not explicitly modelled fortritium and carbon-14, as they exchange quickly between the atmosphere and ground and a specificactivity approach is used instead and). If, for example, a higher deposition velocity were used, it would
clearly lead to higher deposition close to the discharge point but lower deposition at greater distances
because of depletion of radionuclides in the plume. Thus, overall, for the estimation of collective doses,
the effects tend to counteract each other (except for very short-lived radionuclides). However, theestimation of characteristic individual doses is more sensitive to the assumed deposition velocity.
Nevertheless, because this methodology is intended for comparative purposes, the Committee
considered that the simplified approach was justified.
136. The approach used to estimate the transfer of radionuclides through the terrestrial environment tofood is again a standard one. Although the model used, FARMLAND [B6], was developed in the

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 75

United Kingdom, it has been validated against measured data [S8] and is in reasonable agreement withmodels used elsewhere in the world [I6]. Also, to some extent, parameter values that are considered to
be widely applicable are used, as discussed in section III. The FARMLAND model was also used in a
modified form for the Committee’s 2013 assessment of the levels and effects of the nuclear accident
following the 2011 great east-Japan earthquake and tsunami [U12]—agricultural practices andparameter values specific to Japan were used. A comparison of the results of the use of two versions of
FARMLAND—the one using parameter values that are considered to be widely applicable and theother parameter values that are specific to Japan—are compared in electronic attachment 4. Although
there are some differences, these are not significant for the purposes of the methodology.
137. The methodology makes use of consumption rates of terrestrial foods that differ by geographicalregion (see table 7). This is a compromise between the use of a single diet for the world and morecountry-specific consumption rates. From table 7, it is clear that there are differences in the total
amount of food that is eaten annually, which will have a direct influence on both collective andcharacteristic individual doses but these are generally less than a factor of two (differences of up to afactor of 5 are seen for individual food types). There are also differences in the composition of the diet,and it is less clear what impact this will have on estimated doses. For discharges from nuclear sites, oneof the most important radionuclides in terms of both collective and individual doses is carbon-14 [U8].
A specific-activity model is included in the methodology, so that the intake of carbon-14 is related tothe intake of stable carbon in the diet. As carbon is an essential component of the diet and is found in all
foods, it is likely that the intake of carbon-14 is insensitive to changes in the dietary composition. Thiswas investigated in a sensitivity study (described in electronic attachment 4) that considered the intakesof carbon for diets typical of Japan and the United Kingdom. This study showed that the total amountof carbon ingested was similar and therefore the dose estimates would be similar even though the foodsfrom which the carbon was derived were different.
138. The estimation of doses due to external exposure from both radionuclides in the plume and those
deposited on the ground takes account of the shielding effects of buildings when people are indoors.
The methodology uses a single indoor occupancy factor and single shielding/location factors forradionuclides in the plume and those on the ground. The indoor occupancy factor is 0.8 and any
variations in this for much of the world are likely to be less than a factor of two and so the effect would
be relatively minor. The shielding factors, 0.1 and 0.2 for radionuclides deposited on the ground and in
the plume, respectively, are typical of those for standard single-storey buildings. As discussed in
section III.A, there are differences between the shielding factors for different building types (shielding
factors might be a factor of 5 or 6 higher than used here for wooden buildings and a factor of up to
10 times lower for multi-storey buildings). There is also evidence for location factors changing as afunction of time. The empirical models used are based on measurements and there can be good
agreement between the results of such models and personal dosimeter measurements where local
factors are taken into account. The factors used here are considered to be appropriate for use in theCommittee’s assessments, recognizing the variations that occur globally

B. Discharges to freshwater and marine environments
139. The methodology includes consideration of two generic freshwater environments for aquatic
discharges: (a) a small river and (b) a large river; the results for a large river are also considered toapply to a large lake. In all cases, instantaneous mixing is assumed within a single body of water, which
is obviously a significant simplification. In reality, lakes and rivers are complex environments andrivers have many tributaries and sections with different flow rates, volumes and behaviours. Similarly,

76 UNSCEAR 2016 REPORTthe methodology uses a generic approach for discharges to a marine environment with a simple
compartmental model consisting of two compartments representing local and regional marine waters.
Again in reality, there are significant differences between marine environments depending on local andregional currents, and the nature of the coastline and local environment. Although there have beenvalidation studies showing good agreement between models and predictions, these are for specific sites
and situations and are not necessarily applicable to the simplified generic approach adopted here.
However, the Committee considers that its generic approach is adequate for the purpose of assessing
worldwide exposures from radioactive discharges and for comparative studies, but notes the limitationsof the approach for other purposes.
140. Similar to terrestrial foods, the values for consumption rates of freshwater and marine fish are
based on data by geographical region and are suitable for estimating the characteristic individual dose.
There are differences in the values of consumption rates for freshwater fish and some marine fish.
Table 13 for freshwater fish shows differences of up to a factor of 6 between regions, and table 20
shows differences of a factor of 10 or more for the consumption of crustaceans and even greater
differences for consumption rates of molluscs. These differences would be directly reflected in thevalues of the calculated individual doses.
141. The methodology now considers the transfer of radionuclides from freshwater to terrestrial foodsby means of irrigation. There are many different types of irrigation systems and significant variations
across the world in their usage, the source of water (surface or groundwater), and the type of crops thatare irrigated. The estimation of the transfer of radionuclides to terrestrial foods through irrigation istherefore particularly uncertain. The simplifying assumptions in the methodology however are only
intended to be suitable for generic assessments and the Committee considers that its adopted approachis robust enough and suitable for this purpose. Nevertheless, the limitations should be recognized; theimportance of these limitations will depend on the relative importance of the irrigation exposure
pathway compared to exposures from intakes of drinking water and freshwater fish.
142. The estimation of collective dose from discharges of radionuclides to rivers, lakes and the sea is
not based on the sum of individual doses as is the case for discharges to atmosphere. Instead, thecollective dose is based on estimates of the total amount of fish caught in the part of the environment ofinterest, and the amount of water that is abstracted for drinking water and for irrigation. As such, themethodology assumes that the fish, drinking water and irrigated foods are consumed but no account istaken of who consumes them. The Committee considers that this approach is adequate for estimatingcollective doses for its purposes, but notes that it is not possible with its methodology to break down thecollective dose into individual doses or to distinguish between local and regional components ofcollective dose for any aquatic discharges.

C. Global modelling
143. The approach used to estimate collective doses from the four most important radionuclides thatcan be globally dispersed (3H, 14C, 85Kr and 129I) is one that is well accepted and has been used by theCommittee for a number of years. The globally dispersed radionuclides, particularly 14C, are the biggest
contributors to the overall collective doses due to discharges from the nuclear industry when thesedoses are integrated over 500 years or more. One of the factors that influences the estimated collective
doses is the size of the global population that is assumed, because the global collective dose is
proportional to the size of the global population. A rounded value of 10 billion people is included in themethodology; the uncertainty associated with this value increases with time. The global models areANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 77

based on movements of water, carbon and other natural elements through the environment. Suchmovements are likely to be affected by changes in the climate, sea level, ocean currents andatmospheric conditions. The model for carbon is a specific activity approach based on the global carbon
cycle and the amounts of stable carbon in different parts of the environment. Increases in stable carbon
due to the burning of fossil fuels will influence the transfers in the future. These possible future changes
are not taken into account in the methodology, which leads to increasing uncertainties in estimation ofdoses in the longer term. It is also not possible to validate this type of model whose scope covers wide
areas and time frames.

D. Comparison of the current and previous methodologies
144. The methodology described in this annex has been developed from that used previously by theCommittee [U6, U7, U10]. The main changes that have been made are discussed in electronic
attachment 1, with further information given in electronic attachment 2.
145. The main changes that have been introduced are:(a) The results of the methodology are now given in terms of the dose per unit discharge rather
than being normalized to the amount of electricity generated (dose factors normalized toelectricity generated can be calculated subsequently as required);(b) The methodology has been extended for application to the estimation of exposures fromdischarges of radionuclides from non-nuclear electrical energy production;(c) The inclusion of factors that depend on geographical region, notably for population
distributions, consumption rates and fish-catch data;(d) The inclusion of a more detailed approach for modelling the transfer of radionuclides from
fresh water to irrigated crops;(e) The results are provided for a range of integration times;(f) Some updates have been made to the various models included in the methodology.
As noted in electronic attachment 1, many parts of the methodology remain unchanged, because theCommittee considered, following review, that they were still appropriate for its purposes.
146. The results of the various workbooks implementing the methodology were compared with data
derived from the previous methodology [U6, U7]. The estimates of the annual collective effective doses
from the following pathways were compared:(a) External exposure due to deposited radionuclides from discharges to atmosphere;(b) Inhalation of radionuclides discharged to atmosphere;(c) Ingestion of radionuclides discharged to atmosphere;(d) All pathways from radionuclides discharged to both freshwater and marine environments.
In general, the results are within about an order of magnitude, despite the changes that have been
introduced to parts of the methodology (see electronic attachment 1, tables 2–6).

78 UNSCEAR 2016 REPORT

VI. APPLICATION OF THE METHODOLOGY147. The Committee decided that its calculations should be transparent and that the methodology
should be able to be relatively easily applied and updated in the future. It agreed that the methodology
be implemented in a series of Excel® workbooks. The appendix describes the end points considered, theworkbook design, the mathematical equations to implement the methodology in the workbooks, thetreatment of progeny, the quality assurance carried out to ensure that the workbooks were implemented
correctly, and details of how the workbooks could be used for assessing the exposures from different
source terms.
148. The results of the methodology, as given by the workbooks, are for specific discharges todifferent environments. The results are in the form of dose calculation factors for individual
radionuclides; for radionuclides that are in secular equilibrium with their short-lived progeny, thefactors represent the sums of the contributions from the parent and progeny (see appendix). The factors
are for characteristic individual doses (in units of sieverts) in the 100th year of a continuous dischargeof a radionuclide at a rate of 1 Bq/s, and collective doses (in units of man–sieverts) integrated to 100,
500 and 10,000 years for a continuous discharge of a radionuclide at a rate of 1 Bq/s for 1 year (seesection II.C). For most radionuclides the collective doses to 100 years only are provided because thiswas deemed to be sufficient. However collective doses integrated to other times are given for 3H, 14C,
85
Kr and 129I.
149. The resulting characteristic individual and collective dose calculation factors for unit discharge ofradionuclides to atmosphere, to the different freshwater environments (the results for a large river also
apply to a lake) and to a marine environment—derived using the methodology implemented in the
workbooks—are presented in tables 23–31. The doses can then be scaled by the actual rates ofdischarge of each radionuclide from a particular site and then summed over all relevant radionuclides togive the overall characteristic individual and collective doses. Results are given for sites located indifferent regions of the world, with additional results for coastal and inland nuclear sites and for sites
situated in areas of low population density.
150. For future assessments, where discharges are reported to the Committee for broad groups ofradionuclides (e.g. gross alpha, noble gases, radioiodines or particulates), some assumptions would
need to be made about the proportion of the different radionuclides constituting the group (seesection II.D.2).

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 79

Table 23. Estimated characteristic individual doses in the 100th year from a continuous dischargeof a radionuclide to atmosphere
The characteristic individual dose is to a person living 5 km from the discharge point obtaining 25% of their food locally. The
regions are as shown in figure II

Individual effective dose in the 100th year of a continuous discharge at a rate of 1 Bq/s (Sv)Radionuclide

AfricaAsia & Pacific

EuropeLatin
America

North
AmericaWest Asia

Ha

4.2 × 10−14

4.4 × 10−14

4.8 × 10−14

4.3 × 10−14

4.7 × 10−14

4.1 × 10−14

14

C

3.3 × 10−12

3.7 × 10−12

3.7 × 10−12

3.4 × 10−12

3.6 × 10−12

3.6 × 10−1235

S

5.1 × 10−12

7.2 × 10−12

1.4 × 10−11

1.2 × 10−11

1.8 × 10−11

7.6 × 10−1241Ar

4.2 × 10−14

4.2 × 10−14

4.2 × 10−14

4.2 × 10−14

4.2 × 10−14

4.2 × 10−1454

Mn

1.5 × 10−12

1.6 × 10−12

1.9 × 10−12

1.7 × 10−12

2.0 × 10−12

1.5 × 10−1258

Co

6.4 × 10−12

6.8 × 10−12

7.9 × 10−12

7.4 × 10−12

8.4 × 10−12

6.7 × 10−1260

Co

2.3 × 10−10

2.5 × 10−10

3.0 × 10−10

2.9 × 10−10

3.5 × 10−10

2.5 × 10−1065Zn

3.0 × 10−11

3.9 × 10−11

6.7 × 10−11

6.1 × 10−11

8.8 × 10−11

4.0 × 10−1185

Kr

8.7 × 10−17

8.7 × 10−17

8.7 × 10−17

8.7 × 10−17

8.7 × 10−17

8.7 × 10−17

90

Srb

2.4 × 10−10

2.8 × 10−10

3.8 × 10−10

2.9 × 10−10

3.7 × 10−10

2.5 × 10−10

106

Rub

1.9 × 10−11

2.1 × 10−11

2.9 × 10−11

2.7 × 10−11

3.5 × 10−11

2.1 × 10−11

129

I

4.3 × 10−10

4.9 × 10−10

5.6 × 10−10

4.6 × 10−10

5.3 × 10−10

4.5 × 10−10131I

1.3 × 10−11

1.5 × 10−11

1.8 × 10−11

1.4 × 10−11

1.7 × 10−11

1.3 × 10−11

133

Xeb

1.0 × 10−15

1.0 × 10−15

1.0 × 10−15

1.0 × 10−15

1.0 × 10−15

1.0 × 10−15

135

Xe

8.6 × 10−15

8.6 × 10−15

8.6 × 10−15

8.6 × 10−15

8.6 × 10−15

8.6 × 10−15

138

Xeb

3.2 × 10−14

3.2 × 10−14

3.2 × 10−14

3.2 × 10−14

3.2 × 10−14

3.2 × 10−14

134

Cs

1.2 × 10−10

1.4 × 10−10

1.7 × 10−10

1.5 × 10−10

1.8 × 10−10

1.4 × 10−10

137

Csb

2.1 × 10−10

2.2 × 10−10

2.4 × 10−10

2.3 × 10−10

2.6 × 10−10

2.2 × 10−10

210

Pb

1.1 × 10−9

1.2 × 10−9

1.3 × 10−9

1.1 × 10−9

1.2 × 10−9

1.0 × 10−9

210

Po

2.1 × 10−9

2.3 × 10−9

2.7 × 10−9

2.4 × 10−9

2.8 × 10−9

2.1 × 10−9

222

Rnb

2.7 × 10−12

2.7 × 10−12

2.7 × 10−12

2.7 × 10−12

2.7 × 10−12

2.7 × 10−12

226

Ra

2.2 × 10−9

2.3 × 10−9

2.4 × 10−9

2.2 × 10−9

2.3 × 10−9

2.1 × 10−9

230

Th

5.7 × 10−9

5.7 × 10−9

5.7 × 10−9

5.6 × 10−9

5.7 × 10−9

5.6 × 10−9

232

Thb

2.8 × 10−8

2.8 × 10−8

2.8 × 10−8

2.8 × 10−8

2.8 × 10−8

2.8 × 10−8234

U

1.4 × 10−9

1.5 × 10−9

1.5 × 10−9

1.5 × 10−9

1.5 × 10−9

1.4 × 10−9

238

Ub

1.2 × 10−9

1.2 × 10−9

1.2 × 10−9

1.2 × 10−9

1.2 × 10−9

1.2 × 10−9

239

Pu

2.0 × 10−8

2.0 × 10−8

2.0 × 10−8

2.0 × 10−8

2.0 × 10−8

2.0 × 10−8

240

Pu

2.0 × 10−8

2.0 × 10−8

2.0 × 10−8

2.0 × 10−8

2.0 × 10−8

2.0 × 10−8

241

Am

1.7 × 10−8

1.7 × 10−8

1.7 × 10−8

1.7 × 10−8

1.7 × 10−8

1.7 × 10−83aFor 3H the calculated doses include a contribution from HTO and OBT as discussed in section III.A.3.

b

For these radionuclides the calculated doses include a contribution from the ingrowth of progeny (see appendix for details).

Characteristic individual doses are to people living 5 km downstream of the discharge point obtaining 25% of their food locally. The regions are as shown in figure II

Individual effective dose in the 100th year of a continuous discharge at a rate of 1 Bq/s (Sv)Radionuclide

AfricaAsia & Pacific

Small

Large

Small

Large

Hb

3.0 × 10−12

3.0 × 10−14

3.0 × 10−12

14C

1.6 × 10−8

1.6 × 10−10

35

S

6.2 × 10−11

54

Mn

58

EuropeLatin America

Small

Large

Small

Large

3.0 × 10−14

3.0 × 10−12

3.0 × 10−14

3.0 × 10−12

4.5 × 10−8

4.5 × 10−10

2.0 × 10−8

2.0 × 10−10

6.5 × 10−13

1.4 × 10−10

1.3 × 10−12

7.3 × 10−11

5.2 × 10−9

3.5 × 10−12

5.3 × 10−9

3.5 × 10−12Co

4.7 × 10−9

4.0 × 10−12

4.7 × 10−960

Co

1.2 × 10−8

1.1 × 10−1165Zn

1.0 × 10−9

90

Src

106

North America
Small

Large

3.0 × 10−14

3.0 × 10−12

9.3 × 10−9

9.4 × 10−11

8.0 × 10−13

4.5 × 10−11

5.2 × 10−9

3.5 × 10−12

4.0 × 10−12

4.7 × 10−9

1.2 × 10−8

1.1 × 10−11

9.0 × 10−12

2.7 × 10−9

1.2 × 10−9

1.9 × 10−11

Ruc

1.7 × 10−9

129

I

131

West Asia
Small

Largea3.0 × 10−14

3.0 × 10−12

3.0 × 10−14

2.6 × 10−8

2.6 × 10−10

7.6 × 10−9

7.8 × 10−11

4.9 × 10−13

9.0 × 10−11

9.6 × 10−13

4.1 × 10−11

4.7 × 10−13

5.2 × 10−9

3.5 × 10−12

5.2 × 10−9

3.5 × 10−12

5.2 × 10−9

3.5 × 10−12

4.1 × 10−12

4.7 × 10−9

4.0 × 10−12

4.7 × 10−9

4.1 × 10−12

4.7 × 10−9

4.0 × 10−12

1.2 × 10−8

1.1 × 10−11

1.2 × 10−8

1.1 × 10−11

1.2 × 10−8

1.1 × 10−11

1.2 × 10−8

1.1 × 10−11

2.2 × 10−11

1.3 × 10−9

1.1 × 10−11

6.9 × 10−10

6.1 × 10−12

1.6 × 10−9

1.4 × 10−11

5.9 × 10−10

5.3 × 10−12

1.2 × 10−9

1.5 × 10−11

1.2 × 10−9

2.7 × 10−11

1.2 × 10−9

1.9 × 10−11

1.2 × 10−9

2.8 × 10−11

1.2 × 10−9

2.0 × 10−11

3.5 × 10−12

1.7 × 10−9

3.4 × 10−12

1.7 × 10−9

3.7 × 10−12

1.7 × 10−9

3.5 × 10−12

1.7 × 10−9

3.8 × 10−12

1.6 × 10−9

3.5 × 10−12

4.7 × 10−9

6.1 × 10−11

5.1 × 10−9

5.3 × 10−11

4.7 × 10−9

7.4 × 10−11

4.6 × 10−9

5.9 × 10−11

4.8 × 10−9

7.6 × 10−11

4.6 × 10−9

6.3 × 10−11I

4.0 × 10−10

1.9 × 10−12

4.0 × 10−10

1.9 × 10−12

4.0 × 10−10

2.0 × 10−12

4.0 × 10−10

1.9 × 10−12

4.0 × 10−10

2.0 × 10−12

4.0 × 10−10

1.9 × 10−12

134

Cs

8.6 × 10−9

1.8 × 10−11

1.2 × 10−8

2.1 × 10−11

9.2 × 10−9

2.0 × 10−11

7.8 × 10−9

1.7 × 10−11

1.0 × 10−8

2.1 × 10−11

7.6 × 10−9

1.7 × 10−11

137

Csc

4.1 × 10−9

1.1 × 10−11

6.7 × 10−9

1.2 × 10−11

4.5 × 10−9

1.2 × 10−11

3.5 × 10−9

9.9 × 10−12

5.0 × 10−9

1.3 × 10−11

3.4 × 10−9

1.0 × 10−11

210

Pb

2.0 × 10−8

2.2 × 10−10

2.1 × 10−8

2.0 × 10−10

2.0 × 10−8

2.5 × 10−10

1.9 × 10−8

2.1 × 10−10

2.0 × 10−8

2.6 × 10−10

1.9 × 10−8

2.2 × 10−10

210

Po

3.4 × 10−8

3.5 × 10−10

3.7 × 10−8

3.4 × 10−10

3.4 × 10−8

3.9 × 10−10

3.3 × 10−8

3.5 × 10−10

3.5 × 10−8

4.0 × 10−10

3.3 × 10−8

3.5 × 10−10

226

Ra

7.6 × 10−9

1.1 × 10−10

7.8 × 10−9

9.0 × 10−11

7.7 × 10−9

1.4 × 10−10

7.6 × 10−9

1.0 × 10−10

7.7 × 10−9

1.4 × 10−10

7.6 × 10−9

1.1 × 10−10

230

Th

3.2 × 10−9

3.7 × 10−11

3.2 × 10−9

3.4 × 10−11

3.2 × 10−9

4.4 × 10−11

3.2 × 10−9

3.7 × 10−11

3.2 × 10−9

4.6 × 10−11

3.2 × 10−9

3.7 × 10−11

232

Thc

3.3 × 10−8

2.8 × 10−10

3.3 × 10−8

2.5 × 10−10

3.3 × 10−8

3.2 × 10−10

3.3 × 10−8

2.7 × 10−10

3.3 × 10−8

3.3 × 10−10

3.3 × 10−8

2.8 × 10−10234

U

7.4 × 10−10

9.6 × 10−12

7.4 × 10−10

8.4 × 10−12

7.4 × 10−10

1.2 × 10−11

7.4 × 10−10

9.5 × 10−12

7.4 × 10−10

1.3 × 10−11

7.4 × 10−10

9.7 × 10−12

238

Uc

7.3 × 10−10

9.4 × 10−12

7.4 × 10−10

8.3 × 10−12

7.3 × 10−10

1.2 × 10−11

7.3 × 10−10

9.3 × 10−12

7.3 × 10−10

1.2 × 10−11

7.3 × 10−10

9.5 × 10−12

239

Pu

3.5 × 10−9

4.1 × 10−11

3.6 × 10−9

3.7 × 10−11

3.5 × 10−9

4.9 × 10−11

3.4 × 10−9

4.1 × 10−11

3.5 × 10−9

5.1 × 10−11

3.4 × 10−9

4.1 × 10−11

240

Pu

3.5 × 10−9

4.1 × 10−11

3.6 × 10−9

3.7 × 10−11

3.5 × 10−9

4.9 × 10−11

3.4 × 10−9

4.1 × 10−11

3.5 × 10−9

5.1 × 10−11

3.4 × 10−9

4.1 × 10−11

241

Am

3.8 × 10−9

3.4 × 10−11

5.6 × 10−9

3.1 × 10−11

4.1 × 10−9

4.1 × 10−11

3.4 × 10−9

3.3 × 10−11

4.5 × 10−9

4.3 × 10−11

3.3 × 10−9

3.3 × 10−11

3

a

a

a

a

The dose estimates for large rivers are also assumed to apply to lakes.

b

For 3H the calculated doses include a contribution from HTO and OBT as discussed in section III.A.3.

c

For these radionuclides the calculated doses include a contribution from the ingrowth of progeny (see appendix for details).

aa

80 UNSCEAR 2016 REPORT

Table 24. Estimated characteristic individual dose in the 100th year of a continuous discharge of a radionuclide to rivers

Table 25. Estimated characteristic individual dose in the 100th year of a continuous discharge of a radionuclide to a marine environment
The characteristic individual dose is for people ingesting crustaceans and molluscs from the local marine compartment plus 25% of their marine fish consumption from the local compartment and 75% from theregional compartment. The regions are as shown in figure II

Individual effective dose in the 100th year of a continuous discharge at a rate of 1 Bq/s (Sv)Radionuclide
Asia and Pacific

EuropeLatin America

North America

West Asia

H

5.0 × 10−17

1.6 × 10−16

1.5 × 10−16

7.0 × 10−17

1.7 × 10−16

4.0 × 10−17

14C

3.2 × 10−11

1.0 × 10−10

9.4 × 10−11

4.5 × 10−11

1.1 × 10−10

2.6 × 10−1135

S

1.9 × 10−15

1.1 × 10−14

7.6 × 10−15

3.5 × 10−15

9.5 × 10−15

1.5 × 10−1554Mn

2.6 × 10−9

2.7 × 10−9

2.6 × 10−9

2.6 × 10−9

2.7 × 10−9

2.6 × 10−958

Co

4.0 × 10−10

4.6 × 10−10

4.3 × 10−10

4.1 × 10−10

4.5 × 10−10

4.0 × 10−1060

Co

1.2 × 10−9

1.5 × 10−9

1.3 × 10−9

1.2 × 10−9

1.4 × 10−9

1.2 × 10−965Zn

2.8 × 10−10

3.6 × 10−9

2.0 × 10−9

1.3 × 10−9

5.5 × 10−9

5.3 × 10−10

90

Sra

2.5 × 10−13

1.6 × 10−12

1.1 × 10−12

5.0 × 10−13

1.5 × 10−12

2.1 × 10−13

3Rua

2.3 × 10−11

3.7 × 10−11

2.9 × 10−11

2.5 × 10−11

3.3 × 10−11

2.3 × 10−11129

I

2.7 × 10−12

7.6 × 10−12

7.3 × 10−12

3.3 × 10−12

7.0 × 10−12

1.9 × 10−12131I

3.6 × 10−14

7.1 × 10−14

6.8 × 10−14

4.0 × 10−14

6.6 × 10−14

3.0 × 10−14

134

Cs

1.5 × 10−11

2.1 × 10−11

2.3 × 10−11

1.6 × 10−11

2.2 × 10−11

1.4 × 10−11

137

Csa

7.2 × 10−12

1.2 × 10−11

1.3 × 10−11

7.9 × 10−12

1.2 × 10−11

6.3 × 10−12

210

Pba

2.3 × 10−8

4.2 × 10−7

2.2 × 10−7

1.2 × 10−7

5.1 × 10−7

4.1 × 10−8

210

Po

1.0 × 10−8

1.4 × 10−7

7.5 × 10−8

4.0 × 10−8

1.5 × 10−7

1.3 × 10−8

226

Raa

2.4 × 10−8

4.2 × 10−7

2.2 × 10−7

1.3 × 10−7

5.1 × 10−7

4.1 × 10−8

230

Th

3.7 × 10−10

1.6 × 10−9

1.3 × 10−9

6.3 × 10−10

1.8 × 10−9

3.2 × 10−10

232

Tha

6.3 × 10−9

1.1 × 10−8

9.1 × 10−9

7.4 × 10−9

1.2 × 10−8

6.3 × 10−9234

U

2.5 × 10−13

6.8 × 10−12

3.2 × 10−12

1.5 × 10−12

5.5 × 10−12

3.3 × 10−13

238

Ua

6.4 × 10−12

3.2 × 10−11

2.4 × 10−11

1.2 × 10−11

3.4 × 10−11

5.6 × 10−12

239

Pu

9.1 × 10−11

3.0 × 10−9

1.4 × 10−9

5.7 × 10−10

1.9 × 10−9

8.5 × 10−11

240

Pu

9.1 × 10−11

3.0 × 10−9

1.4 × 10−9

5.7 × 10−10

1.9 × 10−9

8.5 × 10−11

241

Am

1.5 × 10−10

1.1 × 10−9

6.0 × 10−10

3.3 × 10−10

9.2 × 10−10

1.5 × 10−10

For these radionuclides the calculated doses include a contribution from the ingrowth of progeny (see appendix for details).

81

a

106

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES

Africa

The collective doses are out to 1,500 km based on the population distributions given in table 3 for different regions of the world as shown in figure II. They can be used for discharges from any type of source forcomparative purposes and are the only values appropriate for non-nuclear sites

Collective doses integrated to 100 years from a year’s continuous discharge at a rate of 1 Bq/s (man Sv)

Radionuclide

AfricaLocal

3

Asia and PacificRegional

Local

EuropeRegional

Local

Latin America

Regional

Local

North America

Regional

Local

West Asia

Regional

Local

Regionalb

1.4 × 10

1.2 × 10

5.6 × 10

4.7 × 10

3.0 × 10

2.5 × 10

2.6 × 10

2.2 × 10

7.3 × 10

6.0 × 10

1.8 × 10C

1.1 × 10−6

5.3 × 10−6

4.3 × 10−6

2.2 × 10−5

2.0 × 10−6

1.0 × 10−5

1.9 × 10−6

9.7 × 10−6

4.9 × 10−7

2.5 × 10−6

H

14

−8

−7

−8

−7

−8

−7

−8

−7

−9

−8

World averagea

−8

−7

Local

Regional

1.5 × 10−8

3.2 × 10

2.7 × 10−7

1.5 × 10−6

7.8 × 10−6

2.4 × 10−6

1.2 × 10−535

S

1.8 × 10−6

8.2 × 10−6

9.1 × 10−6

4.2 × 10−5

8.2 × 10−6

3.8 × 10−5

7.5 × 10−6

3.5 × 10−5

2.7 × 10−6

1.3 × 10−5

3.5 × 10−6

1.6 × 10−5

6.9 × 10−6

3.2 × 10−541Ar

6.1 × 10−10

3.2 × 10−15

2.2 × 10−9

1.1 × 10−14

9.8 × 10−10

5.2 × 10−15

1.0 × 10−9

5.5 × 10−15

2.5 × 10−10

1.3 × 10−15

7.9 × 10−10

4.2 × 10−15

1.2 × 10−9

6.5 × 10−15

54Mn

3.8 × 10−7

1.8 × 10−6

1.6 × 10−6

7.4 × 10−6

9.0 × 10−7

4.3 × 10−6

7.8 × 10−7

3.7 × 10−6

2.3 × 10−7

1.1 × 10−6

5.0 × 10−7

2.4 × 10−6

9.2 × 10−7

4.4 × 10−658

8.9 × 10−7

4.1 × 10−6

3.6 × 10−6

1.7 × 10−5

2.4 × 10−6

1.1 × 10−5

2.2 × 10−6

1.0 × 10−5

6.7 × 10−7

3.1 × 10−6

1.3 × 10−6

6.0 × 10−6

2.3 × 10−6

1.1 × 10−560

3.0 × 10−5

1.4 × 10−4

1.3 × 10−4

6.3 × 10−4

9.6 × 10−5

4.6 × 10−4

9.4 × 10−5

4.5 × 10−4

3.1 × 10−5

1.5 × 10−4

5.0 × 10−5

2.4 × 10−4

9.1 × 10−5

4.3 × 10−4

Co
Co

65

Zn

8.5 × 10−6

4.0 × 10−5

4.3 × 10−5

2.0 × 10−4

3.7 × 10−5

1.7 × 10−4

3.5 × 10−5

1.6 × 10−4

1.2 × 10−5

5.9 × 10−5

1.6 × 10−5

7.6 × 10−5

3.2 × 10−5

1.5 × 10−485

Kr

1.4 × 10−11

1.1 × 10−10

4.8 × 10−11

4.0 × 10−10

2.2 × 10−11

1.8 × 10−10

2.3 × 10−11

1.9 × 10−10

5.5 × 10−12

4.6 × 10−11

1.8 × 10−11

1.5 × 10−10

2.7 × 10−11

2.3 × 10−10

90

Src

8.7 × 10−5

4.2 × 10−4

3.6 × 10−4

1.7 × 10−3

2.3 × 10−4

1.1 × 10−3

1.8 × 10−4

8.6 × 10−4

5.5 × 10−5

2.6 × 10−4

1.2 × 10−4

5.5 × 10−4

2.2 × 10−4

1.0 × 10−3

106

Ruc

3.9 × 10−6

1.9 × 10−5

1.8 × 10−5

8.4 × 10−5

1.3 × 10−5

5.9 × 10−5

1.2 × 10−5

5.7 × 10−5

4.1 × 10−6

1.9 × 10−5

6.2 × 10−6

3.0 × 10−5

1.2 × 10−5

5.6 × 10−5129I

1.6 × 10−4

7.6 × 10−4

6.4 × 10−4

3.1 × 10−3

3.3 × 10−4

1.6 × 10−3

2.9 × 10−4

1.4 × 10−3

7.9 × 10−5

3.8 × 10−4

2.1 × 10−4

1.0 × 10−3

3.6 × 10−4

1.7 × 10−3131I

3.9 × 10−6

1.4 × 10−5

1.6 × 10−5

5.6 × 10−5

9.3 × 10−6

3.2 × 10−5

7.6 × 10−6

2.7 × 10−5

2.2 × 10−6

7.9 × 10−6

5.1 × 10−6

1.8 × 10−5

9.3 × 10−6

3.2 × 10−5

133

Xec

1.6 × 10−10

8.0 × 10−10

5.6 × 10−10

2.9 × 10−9

2.6 × 10−10

1.3 × 10−9

2.7 × 10−10

1.4 × 10−9

6.5 × 10−11

3.3 × 10−10

2.1 × 10−10

1.0 × 10−9

3.2 × 10−10

1.6 × 10−9

135

Xe

8.3 × 10−10

1.7 × 10−10

3.0 × 10−9

5.9 × 10−10

1.4 × 10−9

2.7 × 10−10

1.4 × 10−9

2.9 × 10−10

3.4 × 10−10

6.8 × 10−11

1.1 × 10−9

2.2 × 10−10

1.7 × 10−9

3.4 × 10−10

138

Xec

1.9 × 10−12

1.7 × 10−31

6.7 × 10−12

6.2 × 10−31

3.0 × 10−12

2.8 × 10−31

3.2 × 10−12

3.0 × 10−31

7.7 × 10−13

7.1 × 10−32

2.4 × 10−12

2.3 × 10−31

3.8 × 10−12

3.5 × 10−31

134

Cs

2.9 × 10−5

1.4 × 10−4

1.2 × 10−4

5.9 × 10−4

7.3 × 10−5

3.5 × 10−4

6.9 × 10−5

3.3 × 10−4

2.1 × 10−5

9.9 × 10−5

4.4 × 10−5

2.1 × 10−4

7.5 × 10−5

3.6 × 10−4

137

Csc

3.5 × 10−5

1.7 × 10−4

1.4 × 10−4

6.8 × 10−4

7.9 × 10−5

3.8 × 10−4

7.6 × 10−5

3.6 × 10−4

2.2 × 10−5

1.0 × 10−4

5.0 × 10−5

2.4 × 10−4

8.5 × 10−5

4.1 × 10−4

210

Pb

3.0 × 10−4

1.4 × 10−3

1.2 × 10−3

5.5 × 10−3

6.0 × 10−4

2.9 × 10−3

5.0 × 10−4

2.4 × 10−3

1.4 × 10−4

6.7 × 10−4

3.5 × 10−4

1.7 × 10−3

6.4 × 10−4

3.1 × 10−3

210

Po

4.4 × 10−4

2.1 × 10−3

1.7 × 10−3

8.2 × 10−3

1.1 × 10−3

5.0 × 10−3

9.0 × 10−4

4.2 × 10−3

2.8 × 10−4

1.3 × 10−3

5.5 × 10−4

2.6 × 10−3

1.0 × 10−3

4.9 × 10−3

222

Rnc

4.0 × 10−7

1.7 × 10−6

1.4 × 10−6

6.0 × 10−6

6.4 × 10−7

2.8 × 10−6

6.8 × 10−7

2.9 × 10−6

1.6 × 10−7

6.9 × 10−7

5.1 × 10−7

2.2 × 10−6

8.0 × 10−7

3.4 × 10−6

226

Ra

4.3 × 10−4

2.1 × 10−3

1.7 × 10−3

8.0 × 10−3

8.4 × 10−4

4.0 × 10−3

7.4 × 10−4

3.5 × 10−3

2.0 × 10−4

9.5 × 10−4

5.4 × 10−4

2.6 × 10−3

9.3 × 10−4

4.4 × 10−3

82 UNSCEAR 2016 REPORT

Table 26. Local and regional components of collective dose for discharges to atmosphere from sites in different regions of the world

Collective doses integrated to 100 years from a year’s continuous discharge at a rate of 1 Bq/s (man Sv)

Radionuclide

AfricaAsia and Pacific

EuropeLatin America

North America

West Asia

World averagea

Local

Regional

Local

Regional

Local

Regional

Local

Regional

Local

Regional

Local

Regional

Local

RegionalTh

5.7 × 10−4

2.7 × 10−3

2.0 × 10−3

9.8 × 10−3

9.5 × 10−4

4.5 × 10−3

9.8 × 10−4

4.7 × 10−3

2.4 × 10−4

1.1 × 10−3

7.3 × 10−4

3.5 × 10−3

1.2 × 10−3

5.5 × 10−3

232

Thc

2.9 × 10−3

1.4 × 10−2

1.0 × 10−2

5.0 × 10−2

4.9 × 10−3

2.3 × 10−2

5.0 × 10−3

2.4 × 10−2

1.2 × 10−3

5.8 × 10−3

3.7 × 10−3

1.8 × 10−2

5.9 × 10−3

2.8 × 10−2

234

U

1.5 × 10−4

7.3 × 10−4

5.6 × 10−4

2.7 × 10−3

2.7 × 10−4

1.3 × 10−3

2.7 × 10−4

1.3 × 10−3

6.9 × 10−5

3.3 × 10−4

2.0 × 10−4

9.5 × 10−4

3.2 × 10−4

1.5 × 10−3

238

Uc

1.3 × 10−4

6.2 × 10−4

4.8 × 10−4

2.3 × 10−3

2.3 × 10−4

1.1 × 10−3

2.3 × 10−4

1.1 × 10−3

5.9 × 10−5

2.8 × 10−4

1.7 × 10−4

8.1 × 10−4

2.7 × 10−4

1.3 × 10−3

239

Pu

1.9 × 10−3

9.2 × 10−3

6.8 × 10−3

3.3 × 10−2

3.1 × 10−3

1.5 × 10−2

3.3 × 10−3

1.6 × 10−2

7.9 × 10−4

3.8 × 10−3

2.5 × 10−3

1.2 × 10−2

3.9 × 10−3

1.9 × 10−2

240

Pu

1.9 × 10−3

9.2 × 10−3

6.8 × 10−3

3.3 × 10−2

3.1 × 10−3

1.5 × 10−2

3.3 × 10−3

1.6 × 10−2

7.9 × 10−4

3.8 × 10−3

2.5 × 10−3

1.2 × 10−2

3.9 × 10−3

1.9 × 10−2

241

Am

1.6 × 10−3

7.7 × 10−3

5.7 × 10−3

2.7 × 10−2

2.6 × 10−3

1.3 × 10−2

2.8 × 10−3

1.3 × 10−2

6.6 × 10−4

3.2 × 10−3

2.1 × 10−3

1.0 × 10−2

3.3 × 10−3

1.6 × 10−2a

The world-average value is based on a world-average population density given in table 3 together with a world-average consumption rate for terrestrial foods (table 7).

b

For 3H the calculated doses include a contribution from HTO and OBT as discussed in section III.A.3.

c

For these radionuclides the calculated doses include a contribution from the ingrowth of progeny (see appendix for details of the progeny included).

Table 27. Local and regional components of collective dose for discharges to atmosphere for inland nuclear sites
These collective doses are based on the population distributions out to 1,500 km for inland nuclear sites given in table 4 for the regions shown in figure II. They apply to inland nuclear sites only. Because there were
no inland operating nuclear power plants in Africa or West Asia at the time of analysis, results are not given for these regions

Radionuclide

Collective doses integrated to 100 years from a year’s continuous discharge at a rate of 1 Bq/s (manSv)
Asia and Pacific
Local

EuropeRegional

Local

1.2 × 10

3.3 × 10C

9.2 × 10−6

35

S

41

Latin America
Regional

Local

3.4 × 10

1.6 × 10

1.6 × 10−5

2.3 × 10−6

2.0 × 10−5

3.1 × 10−5Ar

4.6 × 10−9

54

Mn

58

North America

Regional

Local

2.8 × 10

2.4 × 10

6.7 × 10−6

2.1 × 10−6

9.4 × 10−6

2.5 × 10−5

1.1 × 10−14

1.1 × 10−9

3.3 × 10−6

5.4 × 10−6Co

7.8 × 10−6

60

Co

65

Zn

b

H

14

3

World averagea

Regional

Local

2.2 × 10

1.1 × 10−6

8.1 × 10−6

4.8 × 10−15

1.0 × 10−6

1.2 × 10−5

2.8 × 10−4
9.2 × 10−5

−7

Regional

6.7 × 10−8

2.9 × 10

1.1 × 10−7

1.5 × 10−6

2.8 × 10−6

2.2 × 10−6

5.2 × 10−6

4.0 × 10−6

8.1 × 10−6

1.5 × 10−5

6.3 × 10−6

1.4 × 10−5

1.1 × 10−9

8.3 × 10−16

7.4 × 10−10

2.1 × 10−15

1.1 × 10−9

3.9 × 10−15

2.9 × 10−6

8.4 × 10−7

4.3 × 10−7

7.0 × 10−7

1.3 × 10−6

8.4 × 10−7

1.9 × 10−6

2.7 × 10−6

7.3 × 10−6

2.3 × 10−6

1.2 × 10−6

2.0 × 10−6

3.6 × 10−6

2.1 × 10−6

4.7 × 10−6

4.6 × 10−4

1.1 × 10−4

3.1 × 10−4

1.0 × 10−4

5.2 × 10−5

9.2 × 10−5

1.7 × 10−4

8.4 × 10−5

1.9 × 10−4

1.5 × 10−4

4.2 × 10−5

1.2 × 10−4

3.7 × 10−5

1.9 × 10−5

3.7 × 10−5

6.8 × 10−5

2.9 × 10−5

6.5 × 10−5

−7

−8

−7

−8

−8

−8

−8

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 83

230

Collective doses integrated to 100 years from a year’s continuous discharge at a rate of 1 Bq/s (manSv)
Asia and Pacific

EuropeLatin America

North America

World averagea

Local

Regional

Local

Regional

Local

Regional

Local

Regional

Local

Regional

85Kr

1.0 × 10−10

2.8 × 10−10

2.5 × 10−11

1.2 × 10−10

2.5 × 10−11

2.2 × 10−11

1.7 × 10−11

5.1 × 10−11

2.5 × 10−11

9.4 × 10−11

90

Src

7.7 × 10−4

1.3 × 10−3

2.6 × 10−4

7.2 × 10−4

1.9 × 10−4

9.9 × 10−5

1.7 × 10−4

3.0 × 10−4

2.0 × 10−4

4.5 × 10−4

106

Ruc

3.8 × 10−5

6.1 × 10−5

1.4 × 10−5

4.0 × 10−5

1.3 × 10−5

6.6 × 10−6

1.2 × 10−5

2.2 × 10−5

1.1 × 10−5

2.4 × 10−5129I

1.4 × 10−3

2.2 × 10−3

3.8 × 10−4

1.1 × 10−3

3.1 × 10−4

1.6 × 10−4

2.4 × 10−4

4.3 × 10−4

3.3 × 10−4

7.5 × 10−4131I

3.4 × 10−5

4.3 × 10−5

1.1 × 10−5

2.3 × 10−5

8.2 × 10−6

3.2 × 10−6

6.7 × 10−6

9.6 × 10−6

8.5 × 10−6

1.5 × 10−5

133

Xe

1.2 × 10−9

2.2 × 10−9

3.0 × 10−10

9.1 × 10−10

2.9 × 10−10

1.7 × 10−10

1.9 × 10−10

4.0 × 10−10

2.9 × 10−10

7.4 × 10−10

135

Xe

6.3 × 10−9

5.5 × 10−10

1.6 × 10−9

2.5 × 10−10

1.5 × 10−9

4.3 × 10−11

1.0 × 10−9

1.1 × 10−10

1.5 × 10−9

2.0 × 10−10

138

Xec

1.4 × 10−11

5.8 × 10−31

3.5 × 10−12

2.6 × 10−31

3.5 × 10−12

4.5 × 10−32

2.3 × 10−12

1.1 × 10−31

3.5 × 10−12

2.1 × 10−31

134

Cs

2.6 × 10−4

4.3 × 10−4

8.4 × 10−5

2.3 × 10−4

7.4 × 10−5

3.8 × 10−5

6.2 × 10−5

1.1 × 10−4

6.9 × 10−5

1.6 × 10−4

137

Csc

3.0 × 10−4

4.9 × 10−4

9.1 × 10−5

2.5 × 10−4

8.2 × 10−5

4.2 × 10−5

6.5 × 10−5

1.2 × 10−4

7.8 × 10−5

1.8 × 10−4

210

Pb

2.5 × 10−3

4.0 × 10−3

6.9 × 10−4

1.9 × 10−3

5.4 × 10−4

2.8 × 10−4

4.2 × 10−4

7.7 × 10−4

5.9 × 10−4

1.3 × 10−3

210

Po

3.7 × 10−3

6.0 × 10−3

1.2 × 10−3

3.3 × 10−3

9.7 × 10−4

4.9 × 10−4

8.2 × 10−4

1.5 × 10−3

9.6 × 10−4

2.1 × 10−3

222

Rnc

3.0 × 10−6

4.7 × 10−6

7.4 × 10−7

2.0 × 10−6

7.3 × 10−7

3.6 × 10−7

4.8 × 10−7

8.7 × 10−7

7.3 × 10−7

1.6 × 10−6

226

Ra

3.6 × 10−3

5.8 × 10−3

9.7 × 10−4

2.7 × 10−3

7.9 × 10−4

4.1 × 10−4

5.9 × 10−4

1.1 × 10−3

8.5 × 10−4

1.9 × 10−3

230

Th

4.4 × 10−3

7.1 × 10−3

1.1 × 10−3

3.0 × 10−3

1.0 × 10−3

5.4 × 10−4

7.1 × 10−4

1.3 × 10−3

1.1 × 10−3

2.4 × 10−3

232

Thc

2.2 × 10−2

3.6 × 10−2

5.6 × 10−3

1.6 × 10−2

5.3 × 10−3

2.7 × 10−3

3.6 × 10−3

6.7 × 10−3

5.4 × 10−3

1.2 × 10−2234

U

1.2 × 10−3

1.9 × 10−3

3.1 × 10−4

8.6 × 10−4

2.9 × 10−4

1.5 × 10−4

2.1 × 10−4

3.8 × 10−4

3.0 × 10−4

6.7 × 10−4

238

Uc

1.0 × 10−3

1.7 × 10−3

2.7 × 10−4

7.4 × 10−4

2.5 × 10−4

1.3 × 10−4

1.8 × 10−4

3.3 × 10−4

2.5 × 10−4

5.7 × 10−4

239

Pu

1.5 × 10−2

2.4 × 10−2

3.6 × 10−3

1.0 × 10−2

3.5 × 10−3

1.8 × 10−3

2.4 × 10−3

4.3 × 10−3

3.6 × 10−3

8.0 × 10−3

240

Pu

1.5 × 10−2

2.4 × 10−2

3.6 × 10−3

1.0 × 10−2

3.5 × 10−3

1.8 × 10−3

2.4 × 10−3

4.3 × 10−3

3.6 × 10−3

8.0 × 10−3

241

Am

1.2 × 10−2

2.0 × 10−2

3.0 × 10−3

8.4 × 10−3

3.0 × 10−3

1.5 × 10−3

2.0 × 10−3

3.6 × 10−3

3.0 × 10−3

6.8 × 10−3aThe world-average value is based on a world average population density given in table 4 together with a world-average consumption rate for terrestrial foods (table 7).

b

For 3H the calculated doses include a contribution from HTO and OBT as discussed in section III.A.3.

c

For these radionuclides the calculated doses include a contribution from the ingrowth of progeny (see appendix for details).

84 UNSCEAR 2016 REPORTRadionuclideTable 28. Local and regional components of collective dose for discharges to atmosphere for coastal nuclear sites
These collective doses are based on the population distributions out to 1 500 km for coastal nuclear sites given in table 4 for the regions shown in figure II. They apply to coastal nuclear sites only. Because there
were no coastal operating nuclear power plants in West Asia at the time of analysis, results are not given for this region

Collective doses integrated to 100 years from a year’s continuous discharge at a rate of 1 Bq/s (manSv)

Radionuclide

AfricaLocal

Asia and Pacific
Local

Regional

Local

2.3 × 10

6.9 × 10

5.3 × 10

1.6 × 10C

1.7 × 10−6

2.8 × 10−7

4.1 × 10−635

S

2.8 × 10−6

4.3 × 10−741Ar

9.6 × 10−10

54

Mn

58

Latin America

Regional

Local

2.5 × 10

1.2 × 10

7.7 × 10−6

1.6 × 10−6

8.6 × 10−6

1.5 × 10−5

6.1 × 10−17

2.0 × 10−9

6.1 × 10−7

9.7 × 10−8Co

1.4 × 10−6

60

Co

65

North America

Regional

Local

2.2 × 10

6.1 × 10

4.8 × 10−6

1.6 × 10−6

6.7 × 10−6

1.8 × 10−5

4.8 × 10−15

8.1 × 10−10

1.5 × 10−6

2.6 × 10−6

2.2 × 10−7

3.4 × 10−6

4.8 × 10−5

7.7 × 10−6Zn

1.3 × 10−5

85

Kr

90

World averagea

Regional

Local

3.4 × 10

3.0 × 10−6

6.3 × 10−6

3.5 × 10−15

7.4 × 10−7

5.9 × 10−6

1.2 × 10−4

2.1 × 10−6

2.1 × 10−11

Src

106

b

H

14

Regional

6.3 × 10−8

3.6 × 10

1.1 × 10−7

2.3 × 10−6

2.7 × 10−6

2.7 × 10−6

5.0 × 10−6

1.1 × 10−5

1.3 × 10−5

1.4 × 10−5

7.7 × 10−6

1.4 × 10−5

8.7 × 10−10

3.4 × 10−15

1.2 × 10−9

2.2 × 10−15

1.4 × 10−9

3.6 × 10−15

2.1 × 10−6

6.6 × 10−7

1.2 × 10−6

1.1 × 10−6

1.2 × 10−6

1.0 × 10−6

1.8 × 10−6

1.9 × 10−6

5.3 × 10−6

1.8 × 10−6

3.2 × 10−6

3.1 × 10−6

3.4 × 10−6

2.6 × 10−6

4.5 × 10−6

2.2 × 10−4

7.9 × 10−5

2.2 × 10−4

7.9 × 10−5

1.4 × 10−4

1.4 × 10−4

1.6 × 10−4

1.0 × 10−4

1.8 × 10−4

4.0 × 10−5

7.2 × 10−5

3.0 × 10−5

8.4 × 10−5

2.9 × 10−5

5.1 × 10−5

5.8 × 10−5

6.5 × 10−5

3.6 × 10−5

6.4 × 10−5

6.5 × 10−12

4.5 × 10−11

1.4 × 10−10

1.8 × 10−11

8.5 × 10−11

2.0 × 10−11

5.5 × 10−11

2.6 × 10−11

4.8 × 10−11

3.1 × 10−11

9.3 × 10−11

1.4 × 10−4

2.2 × 10−5

3.4 × 10−4

6.1 × 10−4

1.9 × 10−4

5.2 × 10−4

1.5 × 10−4

2.7 × 10−4

2.6 × 10−4

2.9 × 10−4

2.4 × 10−4

4.4 × 10−4

Ruc

6.2 × 10−6

9.9 × 10−7

1.7 × 10−5

3.0 × 10−5

1.0 × 10−5

2.9 × 10−5

1.0 × 10−5

1.8 × 10−5

1.9 × 10−5

2.1 × 10−5

1.3 × 10−5

2.4 × 10−5129

I

2.5 × 10−4

4.0 × 10−5

6.1 × 10−4

1.1 × 10−3

2.7 × 10−4

7.7 × 10−4

2.4 × 10−4

4.3 × 10−4

3.7 × 10−4

4.2 × 10−4

4.0 × 10−4

7.3 × 10−4131I

6.2 × 10−6

6.4 × 10−7

1.5 × 10−5

2.0 × 10−5

7.6 × 10−6

1.7 × 10−5

6.4 × 10−6

9.4 × 10−6

1.1 × 10−5

9.4 × 10−6

1.0 × 10−5

1.4 × 10−5

133

Xec

2.5 × 10−10

3.8 × 10−11

5.3 × 10−10

1.0 × 10−9

2.1 × 10−10

6.6 × 10−10

2.3 × 10−10

4.7 × 10−10

3.0 × 10−10

3.8 × 10−10

3.6 × 10−10

7.1 × 10−10

135

Xe

1.3 × 10−9

3.2 × 10−12

2.8 × 10−9

2.5 × 10−10

1.1 × 10−9

1.8 × 10−10

1.2 × 10−9

1.7 × 10−10

1.6 × 10−9

1.1 × 10−10

1.9 × 10−9

1.8 × 10−10

138

Xec

3.0 × 10−12

3.3 × 10−33

6.3 × 10−12

2.6 × 10−31

2.5 × 10−12

1.9 × 10−31

2.7 × 10−12

1.8 × 10−31

3.6 × 10−12

1.2 × 10−31

4.2 × 10−12

1.9 × 10−31

134

Cs

4.6 × 10−5

7.4 × 10−6

1.2 × 10−4

2.1 × 10−4

6.0 × 10−5

1.7 × 10−4

5.8 × 10−5

1.0 × 10−4

9.7 × 10−5

1.1 × 10−4

8.4 × 10−5

1.5 × 10−4

137

Csc

5.5 × 10−5

8.9 × 10−6

1.3 × 10−4

2.4 × 10−4

6.5 × 10−5

1.8 × 10−4

6.4 × 10−5

1.1 × 10−4

1.0 × 10−4

1.2 × 10−4

9.5 × 10−5

1.7 × 10−4

210

Pb

4.7 × 10−4

7.6 × 10−5

1.1 × 10−3

2.0 × 10−3

4.9 × 10−4

1.4 × 10−3

4.2 × 10−4

7.5 × 10−4

6.6 × 10−4

7.4 × 10−4

7.2 × 10−4

1.3 × 10−3

210

Po

6.9 × 10−4

1.1 × 10−4

1.6 × 10−3

2.9 × 10−3

8.7 × 10−4

2.4 × 10−3

7.6 × 10−4

1.3 × 10−3

1.3 × 10−3

1.4 × 10−3

1.2 × 10−3

2.1 × 10−3

222

Rnc

6.2 × 10−7

7.5 × 10−8

1.3 × 10−6

2.2 × 10−6

5.3 × 10−7

1.4 × 10−6

5.7 × 10−7

1.1 × 10−6

7.6 × 10−7

8.5 × 10−7

8.9 × 10−7

1.5 × 10−6

226

Ra

6.8 × 10−4

1.1 × 10−4

1.6 × 10−3

2.8 × 10−3

6.9 × 10−4

1.9 × 10−3

6.2 × 10−4

1.1 × 10−3

9.3 × 10−4

1.0 × 10−3

1.0 × 10−3

1.9 × 10−3

3

−8

−9

−8

−7

−8

−7

−8

−8

−8

−8

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 85

Regional

EuropeAfrica

Asia and Pacific

EuropeLatin America

North America

World averagea

Local

Regional

Local

Regional

Local

Regional

Local

Regional

Local

Regional

Local

Regional

230

Th

9.0 × 10−4

1.5 × 10−4

1.9 × 10−3

3.5 × 10−3

7.8 × 10−4

2.2 × 10−3

8.2 × 10−4

1.5 × 10−3

1.1 × 10−3

1.2 × 10−3

1.3 × 10−3

2.3 × 10−3

232

Thc

4.6 × 10−3

7.3 × 10−4

9.8 × 10−3

1.8 × 10−2

4.0 × 10−3

1.1 × 10−2

4.2 × 10−3

7.4 × 10−3

5.7 × 10−3

6.4 × 10−3

6.6 × 10−3

1.2 × 10−2234

U

2.4 × 10−4

3.9 × 10−5

5.3 × 10−4

9.5 × 10−4

2.2 × 10−4

6.3 × 10−4

2.3 × 10−4

4.1 × 10−4

3.2 × 10−4

3.7 × 10−4

3.6 × 10−4

6.5 × 10−4

238

Uc

2.1 × 10−4

3.3 × 10−5

4.5 × 10−4

8.1 × 10−4

1.9 × 10−4

5.3 × 10−4

2.0 × 10−4

3.5 × 10−4

2.8 × 10−4

3.1 × 10−4

3.1 × 10−4

5.5 × 10−4

239

Pu

3.0 × 10−3

4.9 × 10−4

6.4 × 10−3

1.2 × 10−2

2.6 × 10−3

7.3 × 10−3

2.8 × 10−3

4.9 × 10−3

3.7 × 10−3

4.2 × 10−3

4.3 × 10−3

7.9 × 10−3

240

Pu

3.0 × 10−3

4.9 × 10−4

6.4 × 10−3

1.2 × 10−2

2.6 × 10−3

7.3 × 10−3

2.8 × 10−3

4.9 × 10−3

3.7 × 10−3

4.2 × 10−3

4.3 × 10−3

7.9 × 10−3

241

Am

2.5 × 10−3

4.1 × 10−4

5.4 × 10−3

9.7 × 10−3

2.2 × 10−3

6.1 × 10−3

2.3 × 10−3

4.1 × 10−3

3.1 × 10−3

3.5 × 10−3

3.6 × 10−3

6.6 × 10−3aThe world-average value is based on a world average population density given in table 3 together with a world-average consumption rate for terrestrial foods (table 7).

b

For 3H the calculated doses include a contribution from HTO and OBT as discussed in section III.A.3.

c

For these radionuclides the calculated doses include a contribution from the ingrowth of progeny (see appendix for details).

86 UNSCEAR 2016 REPORT

Collective doses integrated to 100 years from a year’s continuous discharge at a rate of 1 Bq/s (manSv)

Radionuclide

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 87

Table 29. Local and regional components of collective dose for discharges to atmosphere for low
population density sites
These collective dose estimates are for locations with a low population density out to 1,500 km and are based on a population
density of 5 km−2. They apply to the world as a whole and are intended for use for uranium mines and mill tailings sites

Radionuclide

Collective dose integrated to 100 years from a year’s continuous discharge
at a rate of 1 Bq/s (man Sv)
Local component

Regional componentHa

1.0 × 10−9

8.4 × 10−914C

7.5 × 10−8

3.8 × 10−735

S

2.2 × 10−7

1.0 × 10−6

41Ar

3.8 × 10−11

2.0 × 10−1654Mn

2.9 × 10−8

1.4 × 10−758

Co

7.3 × 10−8

3.4 × 10−760

Co

2.9 × 10−6

1.4 × 10−565Zn

1.0 × 10−6

4.7 × 10−685

Kr

8.6 × 10−13

7.2 × 10−12

90

Srb

6.8 × 10−6

3.2 × 10−5

106

Rub

3.7 × 10−7

1.8 × 10−6129I

1.1 × 10−5

5.4 × 10−5131I

2.9 × 10−7

1.0 × 10−6

133

Xeb

1.0 × 10−11

5.1 × 10−11

135

Xe

5.3 × 10−11

1.1 × 10−11

138

Xeb

1.2 × 10−13

1.1 × 10−32

134

Cs

2.4 × 10−6

1.1 × 10−5

137

Csb

2.7 × 10−6

1.3 × 10−5

210

Pb

2.0 × 10−5

9.6 × 10−5

210

Po

3.3 × 10−5

1.5 × 10−4

222

Rnb

2.5 × 10−8

1.1 × 10−7

226

Ra

2.9 × 10−5

1.4 × 10−4

230

Th

3.6 × 10−5

1.7 × 10−4

232

Thb

1.9 × 10−4

8.9 × 10−4

234

U

1.0 × 10−5

4.8 × 10−5

238

Ub

8.6 × 10−6

4.1 × 10−5

239

Pu

1.2 × 10−4

5.8 × 10−4

240

Pu

1.2 × 10−4

5.8 × 10−4

241

Am

1.0 × 10−4

4.9 × 10−43aFor 3H the calculated doses include a contribution from HTO and OBT as discussed in section III.A.3.

b

For these radionuclides the calculated doses include a contribution from the ingrowth of progeny (see appendix for details).

Values are for small and large rivers situated in the regions shown in figure II. The results for large rivers are also assumed to apply for lakes

Collective doses integrated to 100 years from a year’s continuous discharge at a rate of 1 Bq/s (man Sv)

Radionuclide

Africa
Small

Large

Small

Europe

Large

Small

Latin America
Large

Small

North America

Large

Small

West Asia

Large

Small

World averagea

Large

b

H

1.9 × 10

1.9 × 10

1.9 × 10

1.9 × 10

1.9 × 10

1.9 × 10

1.9 × 10

1.9 × 10

1.9 × 10

1.9 × 10

1.9 × 1014

C

2.3 × 10−3

1.2 × 10−3

2.3 × 10−3

1.2 × 10−3

2.3 × 10−3

1.2 × 10−3

2.3 × 10−3

1.2 × 10−3

2.3 × 10−3

1.2 × 10−335

S

7.4 × 10−6

4.5 × 10−6

7.4 × 10−6

4.8 × 10−6

7.4 × 10−6

4.6 × 10−6

7.4 × 10−6

4.3 × 10−6

7.4 × 10−654

Mn

1.5 × 10−6

9.2 × 10−7

1.5 × 10−6

9.8 × 10−7

1.5 × 10−6

9.4 × 10−7

1.5 × 10−6

8.8 × 10−758

Co

1.6 × 10−6

1.3 × 10−6

1.6 × 10−6

1.4 × 10−6

1.6 × 10−6

1.3 × 10−6

1.6 × 10−660

Co

7.2 × 10−6

6.4 × 10−6

7.2 × 10−6

6.8 × 10−6

7.2 × 10−6

6.5 × 10−665Zn

1.4 × 10−4

6.1 × 10−5

1.4 × 10−4

6.1 × 10−5

1.4 × 10−4

90

Src

7.2 × 10−5

1.0 × 10−4

7.2 × 10−5

1.2 × 10−4

106

Ruc

1.4 × 10−5

1.2 × 10−5

1.4 × 10−5129I

3.1 × 10−4

3.8 × 10−4131I

5.7 × 10−6

134

Cs

137

Small

Large

1.9 × 10−7

1.9 × 10

1.9 × 10−7

2.3 × 10−3

1.4 × 10−3

2.3 × 10−3

1.2 × 10−3

5.2 × 10−6

7.4 × 10−6

1.7 × 10−5

7.4 × 10−6

4.8 × 10−6

1.5 × 10−6

1.1 × 10−6

1.5 × 10−6

3.6 × 10−6

1.5 × 10−6

9.7 × 10−7

1.3 × 10−6

1.6 × 10−6

1.4 × 10−6

1.6 × 10−6

3.1 × 10−6

1.6 × 10−6

1.4 × 10−6

7.2 × 10−6

6.1 × 10−6

7.2 × 10−6

7.2 × 10−6

7.2 × 10−6

2.2 × 10−5

7.2 × 10−6

6.7 × 10−6

6.1 × 10−5

1.4 × 10−4

6.0 × 10−5

1.4 × 10−4

6.2 × 10−5

1.4 × 10−4

9.2 × 10−5

1.4 × 10−4

6.1 × 10−5

7.2 × 10−5

1.1 × 10−4

7.2 × 10−5

8.8 × 10−5

7.2 × 10−5

1.4 × 10−4

7.2 × 10−5

9.8 × 10−4

7.2 × 10−5

1.2 × 10−4

1.2 × 10−5

1.4 × 10−5

1.2 × 10−5

1.4 × 10−5

1.2 × 10−5

1.4 × 10−5

1.2 × 10−5

1.4 × 10−5

1.5 × 10−5

1.4 × 10−5

1.2 × 10−5

3.1 × 10−4

4.5 × 10−4

3.1 × 10−4

4.0 × 10−4

3.1 × 10−4

3.4 × 10−4

3.1 × 10−4

5.3 × 10−4

3.1 × 10−4

3.3 × 10−3

3.1 × 10−4

4.4 × 10−4

5.3 × 10−6

5.7 × 10−6

5.4 × 10−6

5.7 × 10−6

5.4 × 10−6

5.7 × 10−6

5.3 × 10−6

5.7 × 10−6

5.5 × 10−6

5.7 × 10−6

9.1 × 10−6

5.7 × 10−6

5.4 × 10−6

3.5 × 10−4

7.8 × 10−5

3.5 × 10−4

8.7 × 10−5

3.5 × 10−4

8.1 × 10−5

3.5 × 10−4

7.1 × 10−5

3.5 × 10−4

9.8 × 10−5

3.5 × 10−4

4.9 × 10−4

3.5 × 10−4

8.5 × 10−5

Csc

2.4 × 10−4

5.5 × 10−5

2.4 × 10−4

6.3 × 10−5

2.4 × 10−4

5.7 × 10−5

2.4 × 10−4

5.0 × 10−5

2.4 × 10−4

7.2 × 10−5

2.4 × 10−4

3.9 × 10−4

2.4 × 10−4

6.1 × 10−5

210

Pb

1.3 × 10−3

1.2 × 10−3

1.3 × 10−3

1.3 × 10−3

1.3 × 10−3

1.2 × 10−3

1.3 × 10−3

1.2 × 10−3

1.3 × 10−3

1.3 × 10−3

1.3 × 10−3

2.5 × 10−3

1.3 × 10−3

1.3 × 10−3

210

Po

2.3 × 10−3

2.1 × 10−3

2.3 × 10−3

2.1 × 10−3

2.3 × 10−3

2.1 × 10−3

2.3 × 10−3

2.1 × 10−3

2.3 × 10−3

2.1 × 10−3

2.3 × 10−3

2.3 × 10−3

2.3 × 10−3

2.1 × 10−3

226

Ra

4.9 × 10−4

5.9 × 10−4

4.9 × 10−4

6.6 × 10−4

4.9 × 10−4

6.1 × 10−4

4.9 × 10−4

5.4 × 10−4

4.9 × 10−4

7.5 × 10−4

4.9 × 10−4

3.8 × 10−3

4.9 × 10−4

6.5 × 10−4

230

Th

2.0 × 10−4

2.0 × 10−4

2.0 × 10−4

2.0 × 10−4

2.0 × 10−4

2.0 × 10−4

2.0 × 10−4

2.0 × 10−4

2.0 × 10−4

2.0 × 10−4

2.0 × 10−4

2.7 × 10−4

2.0 × 10−4

2.0 × 10−4

232

Thc

1.5 × 10−3

1.6 × 10−3

1.5 × 10−3

1.6 × 10−3

1.5 × 10−3

1.6 × 10−3

1.5 × 10−3

1.5 × 10−3

1.5 × 10−3

1.7 × 10−3

1.5 × 10−3

4.3 × 10−3

1.5 × 10−3

1.6 × 10−3

234

U

4.7 × 10−5

5.1 × 10−5

4.7 × 10−5

5.3 × 10−5

4.7 × 10−5

5.1 × 10−5

4.7 × 10−5

4.9 × 10−5

4.7 × 10−5

5.7 × 10−5

4.7 × 10−5

1.7 × 10−4

4.7 × 10−5

5.3 × 10−5

238

Uc

4.6 × 10−5

5.0 × 10−5

4.6 × 10−5

5.2 × 10−5

4.6 × 10−5

5.1 × 10−5

4.6 × 10−5

4.8 × 10−5

4.6 × 10−5

5.5 × 10−5

4.6 × 10−5

1.6 × 10−4

4.6 × 10−5

5.2 × 10−5

239

Pu

2.3 × 10−4

2.2 × 10−4

2.3 × 10−4

2.2 × 10−4

2.3 × 10−4

2.2 × 10−4

2.3 × 10−4

2.2 × 10−4

2.3 × 10−4

2.2 × 10−4

2.3 × 10−4

3.3 × 10−4

2.3 × 10−4

2.2 × 10−4

240

Pu

2.3 × 10−4

2.2 × 10−4

2.3 × 10−4

2.2 × 10−4

2.3 × 10−4

2.2 × 10−4

2.3 × 10−4

2.2 × 10−4

2.3 × 10−4

2.2 × 10−4

2.3 × 10−4

3.3 × 10−4

2.3 × 10−4

2.2 × 10−4

241

Am

3.1 × 10−4

1.8 × 10−4

3.1 × 10−4

1.8 × 10−4

3.1 × 10−4

1.8 × 10−4

3.1 × 10−4

1.8 × 10−4

3.1 × 10−4

1.9 × 10−4

3.1 × 10−4

3.1 × 10−4

3.1 × 10−4

1.8 × 10−43aAsia and Pacific

−7

−7

−7

−7

−7

−7

−7

−7

−7

−7

−7

−7

The world-average collective dose is calculated using world-average values for irrigation (tables 11, 12 and 17) (note that the catch data for unit river length is the same for all regions and the world-average).
For 3H the calculated doses include a contribution from HTO and OBT as discussed in section III.A.3.
c
For these radionuclides the calculated doses include a contribution from the ingrowth of progeny (see appendix for details).
b

88 UNSCEAR 2016 REPORT

Table 30. Collective dose for discharges to small and large rivers

Table 31. Local and regional components of collective dose for discharges to a marine environment
These results are for discharges to a marine environment for the regions shown in figure II

Collective dose integrated to 100 years from a year’s continuous discharge at a rate of 1 Bq/s (man Sv)

Radionuclide

AfricaLocal

3

Regional

Local

Regional

Local
−13

Regional
−14

Local
−12

Regional
−14

Local
−12

West Asia

Regional
−14

Local
−12

World averagea

Regional
−14

Local
−12

Regional

5.8 × 10

6.0 × 10

4.7 × 10

1.4 × 10

2.8 × 10

1.1 × 10

1.8 × 10

1.0 × 10

5.7 × 10

2.2 × 10

3.1 × 10−14

14

5.5 × 10−7

3.3 × 10−8

2.2 × 10−6

4.0 × 10−8

3.9 × 10−7

3.2 × 10−8

9.2 × 10−7

1.9 × 10−8

6.9 × 10−7

1.2 × 10−8

6.5 × 10−7

3.9 × 10−8

1.4 × 10−6

2.1 × 10−8S

4.9 × 10−11

5.0 × 10−13

2.7 × 10−10

6.6 × 10−13

3.9 × 10−11

4.9 × 10−13

9.7 × 10−11

3.0 × 10−13

6.3 × 10−11

2.0 × 10−13

5.4 × 10−11

5.9 × 10−13

1.8 × 10−10

3.6 × 10−1354

Mn

5.1 × 10−7

1.5 × 10−9

3.9 × 10−6

5.2 × 10−9

4.5 × 10−7

1.4 × 10−9

1.2 × 10−6

1.8 × 10−9

7.2 × 10−7

1.1 × 10−9

5.1 × 10−7

1.7 × 10−9

2.6 × 10−6

3.3 × 10−958

2.5 × 10−7

3.7 × 10−10

1.6 × 10−6

9.9 × 10−10

1.9 × 10−7

3.4 × 10−10

5.8 × 10−7

3.9 × 10−10

3.8 × 10−7

2.5 × 10−10

2.8 × 10−7

4.2 × 10−10

1.1 × 10−6

6.1 × 10−1060

1.4 × 10−6

7.6 × 10−9

8.8 × 10−6

2.0 × 10−8

1.0 × 10−6

7.0 × 10−9

3.1 × 10−6

8.0 × 10−9

2.1 × 10−6

5.2 × 10−9

1.5 × 10−6

8.7 × 10−9

5.8 × 10−6

1.3 × 10−865Zn

2.7 × 10−5

3.1 × 10−8

1.0 × 10−4

1.1 × 10−7

1.2 × 10−5

1.7 × 10−8

5.3 × 10−5

5.5 × 10−8

4.7 × 10−5

4.8 × 10−8

3.5 × 10−5

4.0 × 10−8

7.0 × 10−5

7.2 × 10−8

90

Srb

7.6 × 10−9

2.4 × 10−10

4.1 × 10−8

3.3 × 10−10

5.6 × 10−9

2.3 × 10−10

1.5 × 10−8

1.5 × 10−10

1.0 × 10−8

9.8 × 10−11

8.6 × 10−9

2.8 × 10−10

2.7 × 10−8

1.8 × 10−10

106

Rub

5.7 × 10−8

8.9 × 10−11

4.1 × 10−7

5.1 × 10−10

4.7 × 10−8

7.7 × 10−11

1.4 × 10−7

1.7 × 10−10

8.4 × 10−8

1.1 × 10−10

5.9 × 10−8

9.6 × 10−11

2.7 × 10−7

3.3 × 10−10

129I

3.4 × 10−8

2.8 × 10−9

1.6 × 10−7

3.3 × 10−9

3.0 × 10−8

2.7 × 10−9

5.6 × 10−8

1.6 × 10−9

3.6 × 10−8

1.0 × 10−9

3.8 × 10−8

3.3 × 10−9

1.0 × 10−7

1.7 × 10−9131I

2.4 × 10−10

6.1 × 10−13

1.1 × 10−9

7.2 × 10−13

2.1 × 10−10

6.0 × 10−13

3.9 × 10−10

3.4 × 10−13

2.5 × 10−10

2.2 × 10−13

2.7 × 10−10

7.1 × 10−13

7.1 × 10−10

3.8 × 10−13

134

Cs

5.9 × 10−8

3.9 × 10−9

2.1 × 10−7

4.6 × 10−9

4.6 × 10−8

3.9 × 10−9

8.7 × 10−8

2.2 × 10−9

6.3 × 10−8

1.4 × 10−9

7.0 × 10−8

4.6 × 10−9

1.4 × 10−7

2.4 × 10−9

137

Csb

4.1 × 10−8

3.6 × 10−9

1.5 × 10−7

4.1 × 10−9

3.2 × 10−8

3.5 × 10−9

6.0 × 10−8

2.0 × 10−9

4.4 × 10−8

1.3 × 10−9

4.9 × 10−8

4.2 × 10−9

9.4 × 10−8

2.2 × 10−9

210

Pbb

2.5 × 10−3

1.1 × 10−5

1.2 × 10−2

2.9 × 10−5

1.4 × 10−3

8.9 × 10−6

5.3 × 10−3

1.3 × 10−5

4.2 × 10−3

1.0 × 10−5

3.2 × 10−3

1.3 × 10−5

7.9 × 10−3

1.8 × 10−5

210

Po

7.6 × 10−4

2.6 × 10−6

3.8 × 10−3

5.0 × 10−6

4.6 × 10−4

2.4 × 10−6

1.6 × 10−3

2.2 × 10−6

1.2 × 10−3

1.6 × 10−6

9.2 × 10−4

3.1 × 10−6

2.5 × 10−3

2.9 × 10−6

226

Rab

2.6 × 10−3

1.2 × 10−5

1.2 × 10−2

3.2 × 10−5

1.4 × 10−3

9.7 × 10−6

5.3 × 10−3

1.4 × 10−5

4.3 × 10−3

1.1 × 10−5

3.2 × 10−3

1.5 × 10−5

8.0 × 10−3

2.0 × 10−5

230

Th

8.7 × 10−6

3.5 × 10−7

3.8 × 10−5

4.4 × 10−7

5.8 × 10−6

3.4 × 10−7

1.6 × 10−5

2.1 × 10−7

1.2 × 10−5

1.4 × 10−7

1.0 × 10−5

4.1 × 10−7

2.5 × 10−5

2.4 × 10−7

232

Thb

3.1 × 10−5

7.3 × 10−7

1.4 × 10−4

1.0 × 10−6

1.8 × 10−5

7.0 × 10−7

5.9 × 10−5

4.7 × 10−7

4.7 × 10−5

3.2 × 10−7

3.8 × 10−5

8.6 × 10−7

9.0 × 10−5

5.5 × 10−7234

U

2.9 × 10−8

1.9 × 10−10

1.9 × 10−7

5.3 × 10−10

2.2 × 10−8

1.8 × 10−10

6.7 × 10−8

2.1 × 10−10

4.4 × 10−8

1.3 × 10−10

3.2 × 10−8

2.2 × 10−10

1.3 × 10−7

3.3 × 10−10

238

Ub

1.7 × 10−7

6.1 × 10−9

7.9 × 10−7

7.9 × 10−9

1.2 × 10−7

5.9 × 10−9

3.2 × 10−7

3.7 × 10−9

2.4 × 10−7

2.4 × 10−9

2.0 × 10−7

7.1 × 10−9

5.2 × 10−7

4.3 × 10−9

239

Pu

1.1 × 10−5

9.0 × 10−8

8.4 × 10−5

2.4 × 10−7

9.9 × 10−6

8.8 × 10−8

2.6 × 10−5

8.9 × 10−8

1.5 × 10−5

5.3 × 10−8

1.1 × 10−5

1.0 × 10−7

5.5 × 10−5

1.5 × 10−7

240

Pu

1.1 × 10−5

9.0 × 10−8

8.4 × 10−5

2.4 × 10−7

9.9 × 10−6

8.8 × 10−8

2.6 × 10−5

8.9 × 10−8

1.5 × 10−5

5.3 × 10−8

1.1 × 10−5

1.0 × 10−7

5.5 × 10−5

1.5 × 10−7

241

Am

4.4 × 10−6

6.3 × 10−8

2.7 × 10−5

1.1 × 10−7

3.3 × 10−6

6.0 × 10−8

9.7 × 10−6

4.8 × 10−8

6.5 × 10−6

3.1 × 10−8

4.9 × 10−6

7.3 × 10−8

1.8 × 10−5

6.6 × 10−8aThe world-average value is calculated using the world-average annual catch of fish, crustacea and molluscs from table 21.

b

For these radionuclides the calculated doses include a contribution from the ingrowth of progeny (see appendix for details).

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 89

3.4 × 10

Co

−14

North America

4.8 × 10Co

−12

Latin America

8.5 × 1035

−14EuropeHC−13Asia and Pacific90 UNSCEAR 2016 REPORTVII. CONCLUSIONS
151. An important aspect of the Committee’s work is the periodic assessment of global and regionalradiation exposures to the public from discharges of radionuclides to the environment. Such studies
require a robust, generic methodology that is defensible and takes into account developments in thefield. The methodology used by the Committee for many years has proved to meet its needs forworldwide exposure assessments but, as previously, before any major evaluation of exposures, themethodology has been reviewed by the Committee to ensure its continued validity. An important factor
for the current review was the decision by the Committee to update its evaluations of human radiation
exposures from all significant types of electrical energy production. Following the review, theCommittee decided to make some changes to parts of the methodology, while retaining other parts thatit felt were still appropriate for its purposes. Previously, the results had often been expressed in terms ofdose per unit of electricity generated but, for flexibility, it was agreed that the results be presented interms of dose per unit of activity discharged.
152. The updated methodology can be used to estimate characteristic individual doses (typical of theaverage person living in the area around the source) and collective doses due to discharges toatmosphere, to freshwater bodies (small and large rivers or a lake) and to the sea. A wide range ofradionuclides are included to enable exposures to be assessed for nuclear and non-nuclear forms ofelectrical energy production. The methodology applies to routine, continuous discharges only and it
cannot be used for accidental releases; nor is it suitable for detailed site-specific assessments of doses torepresentative persons for regulatory purposes. As previously mentioned, local, regional and global
components of collective dose are estimated, as appropriate. However, it is now possible to gain
additional insight for discharges to atmosphere on collective doses within different distance bands fromthe discharge point. Global components of collective dose are now also available for integration timesof 100, 500 and 10,000 years.
153. The methodology now includes factors for population densities and for food consumption rates
that vary with geographical region. The Committee added this detail because non-nuclear power
stations are found throughout the world where population densities and patterns of food consumption
are significantly different from a generic world average. However, the geographical regions adopted arestill very large and the inclusion of such region-specific data does not mean that the methodology issuitable for site-specific assessments. Another addition to the methodology is the inclusion of doses dueto the irrigation of terrestrial foods with fresh water into which radionuclides had been discharged. The
extent and nature of irrigation is very variable throughout the world but, for the purposes of thismethodology, a simplified generic approach has been adopted for estimating exposures due toirrigation.
154. This scientific annex describes the methodology, outlines changes from previous versions anddiscusses its limitations and particular areas of uncertainty. Where possible, the results from applying
the methodology have been compared with those from other methodologies and published studies.
Although the purpose of the other studies would have influenced the developers’ choice of models andparameter values and hence the estimated doses, the comparisons still helped to give confidence that theCommittee’s methodology is fit for its purposes.

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 91

155. The methodology has been implemented in a series of Excel® workbooks to ensure that it istransparent for use by the Committee in any future studies. It should also be relatively easy to update
the methodology using such a system. The workbooks enable the user to obtain information on theimportant exposure pathways and radionuclides to gain insight for the Committee’s work. The
workbooks have been checked and verified by people not involved in their development.
156. The Committee considers that the methodology should also be transparent to the wider community andtherefore, in addition to the information presented here, further detail is provided in electronic attachments
available on the UNSCEAR website (http://www.unscear.org/unscear/en/publications/2016.html). The
Committee is satisfied that the methodology, as implemented in the workbooks, is robust, builds on thestrong position of the previous versions of the methodology and is suitable for assessing worldwide
exposures due to routine discharges of radionuclides to the environment. It can be used to assess—in aconsistent way—exposures due to releases from non-nuclear and nuclear sites of electricity generation
including the related fuel-cycle sites (e.g. uranium mining and milling, nuclear fuel reprocessing and nuclear
power stations that are being decommissioned).

VIII. ACKNOWLEDGEMENTS
The Committee wishes to acknowledge with gratitude the experts directly involved in developing themethodology and the related workbooks, and those who carried out quality control checks, and support
staff in their institutes who assisted them. In particular, the Committee would like to thank: theconsultants, C. Robinson and J. Simmonds; the expert group, who provided valuable oversight andadvice for the work; staff, T. Anderson and K. Jones, from Public Health England in the United
Kingdom, who developed the workbook system and supported development of the methodology; andH. Grogan, who reviewed and checked the workbook system. The views expressed in the scientificannex and related material remain those of the Committee and do not necessarily represent the views ofindividual experts, the United Nations or its Member States.
Expert group
Co-chairs: E. Waller (Canada)
Members: T. Anderson (United Kingdom), L. Anspaugh (United States), H. Grogan (United States),
G. Hirth (Australia), K. Jones (United Kingdom), L. Hubbard (Sweden), R. Michel (Germany),
J. Simmonds (United Kingdom)
Observer: G. Proehl (IAEA)The Committee would also like to acknowledge with gratitude: the staff of the Metadata and SocioEconomics unit of UNEP/DEWA/GRID-GENEVA who analysed the population data around nuclear
power plants; P. Verger and colleagues from WHO for their analyses of food consumption data;
S. Golikov and M. Balonov for reviewing the model for external exposure from deposited material; andM. Tzivaki for conducting some specialized quality checks of the workbooks and their results.

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 93

APPENDIX A. IMPLEMENTATION OF THE METHODOLOGY
A1. The aim of this appendix is to provide more details of how the Committee has implemented themethodology as a set of Excel® workbooks by describing the workbook design (including the end
points considered, the mathematical equations to implement the methodology in the workbooks, andinformation on how the workbooks can be used to assess exposures from different types of discharge),
details of how radioactive progeny have been treated, and quality assurance.

I. WORKBOOK DESIGN
A2. The methodology has been implemented as a series of Excel® workbooks from which dose
calculation factors (i.e. the calculated individual and collective doses from unit discharge ofradionuclides under a set of defined conditions) can be derived. The series includes a workbook foreach environment into which a discharge can be made (i.e. atmosphere, freshwater bodies, and marine
environments); a further workbook summarizes the doses due to discharges into each of the threeenvironments, and another provides dose values for globally circulating radionuclides. Table A1summarizes the various end points that are produced by the workbooks, and how the relevant
populations and pathways are treated.
A3. The workbook for each environment into which a discharge is made comprises a number ofinterlinked worksheets. An initial status worksheet describes the purpose of the workbook and provides
key information about its format and contents, version and quality control. Other worksheets containvalues of the model input parameters (radionuclide-specific ones and others such as habit data);intermediate calculations (e.g. of activity concentrations of radionuclides as a function of distance fordischarges to atmosphere); and output values for the end points (i.e. dose calculation factors).
A4. Discharges to atmosphere. The workbook first calculates activity concentrations of radionuclides
in air at a number of distances from the point of discharge. Deposition rates and activity concentrationsof radionuclides on the ground surface, in soil and in crops, meat and milk are derived for eachdistance. These activity concentrations are used to calculate characteristic individual and collective
doses due to inhalation of the plume, and due to external exposure from radionuclides in the plume anddeposited on the ground. Consumption rates of terrestrial food for different geographical regions areused to calculate characteristic individual and collective doses from consumption of foods. Table A2
lists the equations that are implemented for discharges to an atmospheric environment that give rise toexposure through terrestrial pathways.
A5. Discharges to a freshwater environment. The workbook calculates activity concentrations ofradionuclides in water for two sizes of river (the results for a large river are also applied for a lake).
Concentration factors are used to derive activity concentrations in freshwater fish, and irrigation rates
are used to calculate activity concentrations in irrigated crops. Activity concentrations of radionuclides
in riverbank sediment are assumed to be the same as those in riverbed sediment. These values are used
to derive characteristic individual and collective doses from consumption of irrigated foods and water

94 UNSCEAR 2016 REPORT

abstracted from the river, consumption of fish caught from the river and individual doses from external
exposure to radioactive material in riverbank sediments. Table A3 lists the equations implemented fordischarges to a freshwater environment.
A6. Discharges to a marine environment. The workbook is based on the assumption that radionuclides
are discharged to a relatively small, local compartment and water in the local compartment is
exchanged with that in a larger, regional compartment. Activity concentrations in water are calculated
for each compartment, and activity concentrations in fish, crustaceans, molluscs and beach sediment arederived. The workbook calculates individual doses from external exposure to beach sediment, andindividual and collective doses from consumption of shellfish harvested in the local compartment andconsumption of fish caught in both compartments. Table A4 lists the relevant equations for discharges
to a marine environment.
A7. Summary workbook. This is linked to the above workbooks for each environment into which adischarge is made and, for this reason, must be saved in the same electronic folder as them. It presentscharacteristic individual and collective doses for each geographical region from discharges toatmospheric, freshwater and marine environments. The doses from discharges of radionuclides withvery short-lived progeny (e.g. 137Cs and 137mBa) include the contributions from radioactive progeny. In
other cases (i.e. radionuclides with longer-lived progeny), the contribution from the progeny are
considered separately (see section II below). If the degree of equilibrium between different members of
the radionuclide chain is significantly different from that assumed here, then the doses from the parent
and progeny have to be summed once the activity of radionuclides in the discharge and the degree of
equilibrium are specified.
A8. Globally circulating radionuclides. The final workbook includes calculation factors for collective
doses from the global circulation of the relevant radionuclides for 100, 500 and 10,000 years following
unit discharge. These factors have been derived using the PC-CREAM 08 computer system [S8], which
implements each of the global circulation models adopted by the Committee.
A9. The results can be scaled by the actual rates of discharge of each radionuclide from a particularsite and then summed over all relevant radionuclides to give the overall characteristic individual andcollective doses. Results are produced for sites located in different regions of the world, with additional
results for coastal and inland nuclear sites and sites situated in areas of low population density.

II. TREATMENT OF RADIOACTIVE PROGENY
A10. For most radionuclides the workbooks for atmospheric, freshwater and marine environments
provide the relevant dose factors separately for the parent and any progeny. (This permits the
workbooks to be used when non-equilibrium conditions apply). Tables A5, A6 and A7 contain further
details on how dose contributions from radioactive progeny have been treated for terrestrial, freshwaterand marine pathways respectively, following the general principles outlined in section II.D.2. The
summary workbook contains summed values based on the assumptions detailed in those tables.
However, for radionuclides that are in secular equilibrium with their short-lived progeny, the doses
presented and clearly marked in the workbooks are the sums of the doses from the parent and progeny.

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 95

III. QUALITY ASSURANCE
A11. A systematic review of the completed workbooks was carried out for quality assurance purposes,
firstly by the workbook developers and then by independent reviewers. This included a comparison of
the parameter values and variables with source data, and the workbook equations with the agreed
calculation methodology, defined by the lists of equations and data presented below in tables A2-A4.
This ensured that the workbooks had implemented the methodology agreed by the Committee.
A12. As further verification, a workbook was created that compared a sample of the results of the
workbooks with those published in the UNSCEAR 2000 Report [U6]. The results are discussed indetail in electronic attachment 1.

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 97

TABLES
Table A1. Summary of end points, populations and pathways considered(a) Characteristic individual dose
End point/pathway

Key features of methodology

Integration considerations

DISCHARGE TO ATMOSPHERE
Characteristic individual dose

External irradiation (beta + gamma
radiation from plume, and gamma
radiation from deposited radionuclides)

Total dose from all exposure pathways assuming
location 5 km distant from site of discharge

Time integration: 100 years
Individual dose in 100th year following
100 years continuous discharge

Activity concentrations of radionuclides in plumederived using approach in [U6]Time-integrated activity concentrations in soil from
[S8]
Dose factors (table A3 below)

Inhalation

Activity concentrations of radionuclides in plumederived using approach in [U6]

Ingestion of terrestrial foods

Activity concentrations of radionuclides interrestrial foods per unit deposition rate from [S8]Approaches for 14C and 3H based on [I5]

DISCHARGE TO FRESHWATER ENVIRONMENT
Characteristic individual dose

Total dose from all exposure pathways assuming
complete instantaneous mixing
For two situations: discharge into large and small
rivers (for nuclear and mines/non-nuclear
discharges, respectively)

External irradiation from occupancy on
freshwater sediments

Time-integrated activity concentrations ofradionuclides in sediment from [S8] 5 km distantfrom discharge point

Drinking water

Ingestion rate from [U6]

Fish ingestion

Region-specific consumption rates from [F3]Ingestion of irrigated terrestrial foods

Deposition rate related to generic irrigation rate
and growing period appropriate for grain [W2]Time integration: 100 years
Individual dose in 100th year following
100 years continuous discharge

Approaches for 14C and 3H derived from [C1]

DISCHARGE TO MARINE ENVIRONMENT
Characteristic individual dose

Total dose from exposure pathways calculatedusing a two-compartment model
Activity concentrations of radionuclides inseawater in local marine compartment

External irradiation from occupancy on
marine sediments

Time-integrated activity concentrations ofradionuclides in sediments within the local
marine box [S8]Ingestion of marine foods (fish,
crustaceans and molluscs)

Ingestion of crustaceans and molluscs from thelocal compartment (25%) and regional
compartment (75%)

Time integration: 100 years
Individual dose in 100th year following
100 years continuous discharge

98 UNSCEAR 2016 REPORT(b) Collective dose
Component

Key features of methodology

Integration considerations

DISCHARGE TO ATMOSPHERE
Local component

Regional component

Based on individual dose at midpoint and numberof people in annuli around discharge point
derived from population density informationfrom [F1] for nuclear sites or using a population
density of 5 km−2 for sites in areas of low
population (see electronic attachment 3)

Time integration: 100 years

Based on individual dose and number of people inannuli around discharge point for nuclear sites(see electronic attachment 3) or derived fromweighted population density information from
[F1]

Time integration: 100 years, 500 years

Distance integration: using activity
concentrations of radionuclides andindividual doses at centre point of annulararea, 0–100 km

Distance integration: using activity
concentrations of radionuclides andindividual doses at centre point offollowing annular areas: 100–500,
500−1 000 and 1 000−1 500 km
For discharge continuing for 100 years andintegration times of 100 years

Global component

Based on collective dose from globally dispersedradionuclides [S8], which implements models
referred to in [U6]Time integration: 100 years, 500 years and
10 000 years

DISCHARGE TO FRESHWATER ENVIRONMENT
Local and regional components from
freshwater pathways (external
irradiation from exposure on freshwater
sediment not included)

Based on total collective intakes from abstractionof drinking water and freshwater fish-catch datafrom specified lengths of small and large rivers

Global component

Based on collective dose from globally dispersedradionuclides [S8], which implements models
referred to in [U6]Time integration: 100 years

Consumption of irrigated terrestrial foods basedon generalized yield data and areas of land
equipped for irrigation as function of region(from [F1])DISCHARGE TO MARINE ENVIRONMENT
Regional component(External exposure on marine sediment
not included)
Global component

Based on total collective intake from seafoodcatch data derived from the representative
regional marine compartment
Based on collective dose from globally dispersedradionuclides [S8], which implements models
referred to in [U6]Time integration: 100 years

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES

Table A2. Parameter and equation list: Atmospheric discharge and terrestrial environment
Parameter

Symbol

Value

Units

Reference

ATMOSPHERIC DISPERSION
Activity concentration in air at a specified distance calculated using:

Cair ,i ( x ) = D1 ⋅ Q ⋅ x − n e − λi x ua

(T1)

Activity concentration of radionuclide, i, in air
at distance, x, from discharge point

Cair,i ( x )Equation (T1)

Bq/m3

—

Dilution factor at 1 km

D1

5.3 × 10−7

s/m3

[U6]

Discharge rate

Q

Unit discharge rate = 1

Bq/s

—

Downwind distance from discharge pointx

5, 50, 300, 750, 1 250

km—

Empirical indexnNoble gases,
tritium: 1.2
14
C: 1.4
All others: 1.42

—

[U6]

Radioactive decay constant of radionuclide, iλiRadionuclide-specific

s−1

[I9]

Geometric mean of wind speed at the heightof release representative of one yearua

2 × 10−3

km/s

[I2]

Activity concentration in air of 135Cs (progeny of 135Xe) and 138Cs (progeny of 138Xe) calculated using:
𝐶

,

𝑥 =𝐶

,

𝑥

progeny
progeny

parent1−𝑒progeny

parent(T2)(note this equation is valid for one progeny and does not include a branching fraction)
Activity concentration of progeny in air at
distance, x, from discharge point

Cair,progeny ( x )Equation (T2)

Bq/m3

—

Activity concentration of parent in air at
distance, x, from discharge point

Cair,parent ( x )Equation (T1)

Bq/m3

—

Radioactive decay constant of
parent/progeny

λparent , λprogeny

Radionuclide-specific

s−1

[I9]

Downwind distance from discharge pointx

5, 50, 300, 750, 1 250km—

Geometric mean of wind speed at the heightof release representative of one yearua

2 × 10−3

km/s

[I2]

Deposition rate at a specified distance calculated using:di ( x ) = VT Cair ,i ( x )Total daily-average deposition rate at
distance, x, of radionuclide, i, from both
wet and dry processesdi ( x )Equation (T3)

(T3)
Bq/(m2 s)—99

100 UNSCEAR 2016 REPORTParameter
Total deposition velocity (wet and dry)SymbolValue

VT

Noble and non-reactive gases: 0

Units

Reference

m/s

[U6]

Bq/m3

—

Tritium: 0
14

C: 0

All others: 0.002
Activity concentration of radionuclide, i, in
air at distance, x, from discharge point

Cair,i ( x )Equations (T1), (T2)ENVIRONMENTAL TRANSFER
Activity concentration in terrestrial food calculated using:C f ,i ( x ) = C f ,unit,i di ( x )(T4)Activity concentration of radionuclide, i, in
terrestrial food, f, at distance, x, fromdischarge pointC f ,i ( x )Equation (T4)

Bq/kg or Bq/L—

Activity concentration of radionuclide, i, inyear 100 in terrestrial food, f, per unit
deposition rate for continuous depositionC f ,unit,i

PC-CREAM output [S8]Bq/kg per Bq/(m2 s)orBq/L per Bq/(m2 s)—

Total daily average deposition rate at
distance, x, of radionuclide, i, from both
wet and dry processesdi ( x )Equation (T3)

Bq/(m2 s)—

14

C activity concentration in terrestrial food (cereal, vegetables and pasture) calculated using:C=( x ) Cair,14C ( x ) ⋅f ,14 C

Sp

(T5)

Sair

Activity concentration of 14C in food, andpasture, f, at distance, x, from discharge
point

C f ,14 C ( x )Equation (T5)

Bq/kg

[I5]

Activity concentration of 14C in air at distance,x, from discharge point

Cair,14C ( x )Equation (T1)

Bq/m3

—

Concentration of stable carbon in the plant(fresh weight)

Sp

Cereal: 3.9 × 102

g C/kg[I5]

g C/m3[I5]Vegetables: 30
Pasture: 1 × 102

Concentration of stable carbon in air
14

0.2

Sair

C activity concentration in terrestrial food (animal products) calculated using:C f ,14 C ( x ) =f C ⋅ Cpasture,14C ( x ) ⋅ S a

(T6)

Sp

Activity concentration of 14C in food, f, atdistance, x, from discharge pointC f ,14 C ( x )Equation (T6)

Bq/kg

[I5]

Fraction of animal feed that is contaminated

fC

1

—

[I5]

Activity concentration of 14C in pasture at
distance, x, from discharge point

Cpasture,14C ( x )Equation (T5)

Bq/kg

—

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 101

Parameter

SymbolConcentration of stable carbon in the animal
product (fresh weight)

Sa

Concentration of stable carbon in the plant(fresh weight)

Sp

Value

Units

Milk: 65

Reference

g/kg

[I5]

g/kg

[I5]

Beef: 200
Pasture: 1 × 102

Activity concentration of HTO in soil water calculated using:

Csw,HTO ( x ) =CR s-a ⋅ Cair,HTO ( x )
Ha

(T7)

Activity concentration of HTO in soil water at
distance, x, from discharge point

Csw,HTO ( x )Equation (T7)

Bq/L

[I5]

Empirical constantCR s-a

0.3

—

[I5]

Activity concentration of HTO in air at
distance, x, from discharge point

Cair,HTO ( x )Equation (T1)

Bq/m3

—

Absolute humidityHa

6 × 10−3

L/m3

[I5]

Calculated using:



C
( x)
C f ,HTO ( x=
) WC p ⋅  RH air,HTO + (1 − RH ) Csw,HTO ( x ) γ
H
a



(T8)

Activity concentration of HTO in food, f, atdistance, x, from discharge point (fresh
weight)C f ,HTO ( x )Equation (T8)

Bq/kg

[I5]

Fractional water content of the plant (fresh
weight)WC p

Cereal: 0.12

L/kg

[I5]

Vegetables: 0.92
Pasture: 0.76

Relative humidity

RH

0.7

—

[I4]

Activity concentration of HTO in air at
distance, x, from discharge point

Cair,HTO ( x )Equation (T1)

Bq/m3

—

Absolute humidityHa

6 × 10−3

L/m3

[I2]

Activity concentration of HTO in soil water at
distance, x, from discharge point

Csw,HTO ( x )Equation (T7)

Bq/L

[I5]

Ratio of HTO vapour pressure to H2O vapour
pressure

γ

0.909

—

[I5]

Weighted concentration of HTO in drinking water and feed calculated using:

CR f ,HTO ( x ) = 0.5 ⋅ Csw,HTO ( x ) + 0.5 ⋅Weighted concentration of HTO in drinking
water and of water in feed (assumes that
50% intake is drinking water and 50% is
pasture)CR f ,HTO ( x )Equation (T9)

Cpasture,HTO ( x )

(T9)

WC p
Bq/L

—

102 UNSCEAR 2016 REPORTParameter

Symbol

Value

Units

Reference

Activity concentration of HTO in soil water at
distance, x, from discharge point

Csw,HTO ( x )Equation (T7)

Bq/L

[I5]

Activity concentration of HTO in pasture at
distance, x, from discharge point (fresh
weight)Cpasture,HTO ( x )Equation (T8)

Bq/kg

[I5]

Fractional water content of the plant

WCp

0.76

L/kg

[I5]

HTO activity concentration in terrestrial food (animal products) calculated using:C f ,HTO=( x ) CR a ,HTO ⋅ CR f ,HTO ( x )(T10)

Activity concentration of HTO in food, f, atdistance, x, from discharge pointC f ,HTO ( x )Equation (T10)

Bq/kg

[I5]

Concentration ratio: concentration of HTO in
product (fresh weight) divided byconcentration of HTO in water intake

CR a ,HTO

Milk: 0.87

L/kg

[I5]

Weighted concentration of HTO in drinking
water and of water in feed

CR f ,HTO ( x )

Bq/L

—

Beef: 0.66
Equation (T9)

OBT activity concentration in terrestrial food (cereal, vegetables and pasture) calculated using:C f ,OBT ( x ) =(1 − WC p ) ⋅WEQ p ⋅ R p ⋅ C f ,HTO ( x )

(T11)

WC p

Activity concentration of OBT in food, f, atdistance, x, from discharge pointC f ,OBT ( x )Equation (T11)

Bq/kg

[I5]

Fractional water content of the plant (fresh
weight)WC p

Cereal: 0.12

L/kg

[I5]

L/kg

[I5]

Vegetables: 0.92
Pasture: 0.76

Water equivalent factor

WEQ p

Cereal: 0.56
Vegetables: 0.51
Pasture: 0.56

Partition factor

Rp

0.54

—

[I5]

Activity concentration of HTO in food, f, atdistance, x, from discharge pointC f ,HTO ( x )Equation (T8)

Bq/kg

[I5]

OBT activity concentration in terrestrial food (animal products)

C f ,OBT=( x ) CR a ,OBT ⋅ Cpasture,OBT ( x )(T12)Activity concentration of OBT in food, f, atdistance, x, from discharge pointC f ,OBT ( x )Equation (T12)

Bq/kg

[I5]

Concentration ratio: concentration of OBT inmilk (fresh weight) divided byconcentration of OBT in intake (dry weight)

CR a ,OBT

Milk: 0.24

kg/L

[I5]

Mean concentration of OBT in feed

Cpasture,OBT ( x )

Bq/kg

—

Beef: 0.40
Equation (T11)ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 103

Parameter

Symbol

Value

Units

Reference

INDIVIDUAL DOSE CALCULATIONS
Effective dose from inhalation at a specified distance calculated using:

H E ( inh ),i ( x=) Cair,i ( x ) ⋅ Dinh,i ⋅I inh⋅ Oann86 400(T13)

Effective dose from inhalation ofradionuclide, i, at distance, x, fromdischarge point

H E (inh ),i ( x )Equation (T13)

Sv—

Activity concentration of radionuclide, i, in air
at distance, x, from discharge point

Cair,i ( x )Equation (T1)

Bq/m3

—

Dose coefficient for inhalation ofradionuclide, i, (see equation (T14) for
calculation of dose coefficient for 222Rn plus
short-lived progeny)

Dinh,i

Dependent on radionuclide

Sv/Bq

[I14]

Adult inhalation rate

I inh

20

m3/d

[U6]

86 400s/d—3.15 × 107 (rounded)s—

Conversion factor
Annual occupancy at location

Oann

Dose coefficient for inhalation of radon and its short-lived progeny calculated using:

Dinh,222Rn= 24 ⋅Dose coefficient for inhalation of 222Rn plus
short-lived progeny(

DC 222 Rn EF222 Rn,in ⋅ O222 Rn,in + EF222 Rn,out ⋅ O222 Rn,out

Dinh,222Rn

Conversion factor

)

I inh

(T14)

Equation (T14)

Sv/Bq

[U6]

24

h/d

—

Dose conversion factor for radon

DC 222 Rn

9 × 10−9

Sv per (Bq h/m3)[U6]

Indoor equilibrium factor for radon

EF222 Rn,in

0.4

—

[U6]

Indoor occupancy factor

O222 Rn,in

0.8

—

[U6]

Outdoor equilibrium factor for radon

EF222 Rn,out

0.6

—

[U6]

Outdoor occupancy factor

O222 Rn,out

0.2

—

[U6]

Adult inhalation rate

I inh

20

m3/d

[U6]

Dose from external exposure from immersion in the plume at a specified distance calculated using:

H E (ex,cloud),i=( x ) Cair,i ( x ) ⋅ Dex,cloud,i ⋅ Oann ⋅ ( Oout + (1 − Oout ) Lcloud )Effective gamma dose from external
exposure to radionuclide, i, at distance, x,from discharge point

H E (ex,cloud),i ( x )Equation (T15)

Sv

(T15)
—

104 UNSCEAR 2016 REPORTParameter

Symbol

Value

Units

Reference

Activity concentration of radionuclide, i, in air
at distance, x, from discharge point

C air,i ( x )Equation (T1)

Bq/m3

External dose coefficient for immersion in theplume

Dex,cloud,iDependent on radionuclide

Sv per (Bq s/m3)

Annual occupancy at location

Oann

3.15 × 107 (rounded)s—

Fraction of occupancy that is outdoors

Oout

0.2

—

[U4]

Building shielding factor in plume (location
factor)Lcloud0.2——[E2, P2]

Dose from external exposure from deposits calculated using:

H E ( ex,deposit ),i ( x=) di ( x ) ⋅ tdischarge ⋅ Dex,deposit,i ⋅ ( Oout + (1 − Oout ) Ldeposit )(T16)

Annual effective dose from the depositedradionuclide, i, at distance, x, fromdischarge point

H E ( ex,deposit ),i ( x )Equation (T16)

Sv—

Total daily average deposition rate at
distance, x, of radionuclide, i, from both
wet and dry processesdi ( x )Equation (T3)

Bq s/m2—

Duration of deposition

tdischarge

3.15 × 107 (rounded)s

1 year

Dose coefficient for external exposure from
deposited material, integrated to 100 years(time dependence taken into account using
equation given by [G2])Dex,deposit,iDependent on radionuclide

Sv per (Bq/m2)[E2, P2]

Fraction of occupancy that is outdoors

Oout

0.2

—

[U4]

Building shielding factor from depositedmaterial (location factor)

Ldeposit

0.1—

Effective dose from ingestion calculated using:

H E (ing ), f ,r ,i ( x=
) C f ,i ( x ) ⋅ Ding,i ⋅ Flocal ⋅ I f ,r(T17)

Annual effective dose from the ingestion offood, f, containing radionuclide, i, atdistance, x, from discharge point in region r

H E (ing ), f ,r ,i ( x )Equation (T17)

Sv—

Activity concentration of radionuclide, i, in
terrestrial food, f, at distance, x, fromdischarge pointC f ,i ( x )Equations (T4), (T5), (T6), (T8),(T10), (T11), (T12)

Bq/kg or Bq/L—

Dose coefficient for ingestion

Ding,i

Dependent on radionuclide

Sv/Bq

[I14]

Fraction of food that is locally produced

FlocalIndividual dose: 0.25—

Collective dose: 1.0
Amount of food, f, ingested per caput in ayear in region, rI f ,rDependent on region andfood

kg or L[W2]

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 105

Parameter

Symbol

Value

Units

Reference

Total effective dose from ingestion calculated using:

H E ( ing ),r ,i ( x ) = ∑ H E (ing ), f ,r ,i ( x )

(T18)

f

Total annual effective dose in region, r, fromingestion of radionuclide, i, in food at
distance, x, from discharge point

H E (ing ),r ,i ( x )Equation (T18)

Sv—

Annual effective dose from ingestion of food,f, containing radionuclide, i, at distance, x,from discharge point in region, r

H E (ing ), f ,r ,i ( x )Equation (T17)

Sv—

Total individual effective dose calculated using:

H E ( atmos ),r ,i ( x ) =H E (inh ),i ( x ) + H E (ex,cloud),i ( x ) + H E ( ex,deposit ),i ( x ) + H E (ing ),r ,i ( x )(T19)

Total effective dose in region, r, fromradionuclide, i, at distance, x, fromdischarge point

H E ( atmos ),r ,i ( x )Equation (T19)

Sv—

Effective dose from inhalation ofradionuclide, i, at distance, x, fromdischarge point

H E (inh ),i ( x )Equation (T13)

Sv—

Effective dose from external exposure toradionuclide, i, at distance, x, fromdischarge point

H E (ex,cloud),i ( x )Equation (T15)

Sv—

Effective dose from the depositedradionuclide, i, at distance, x, fromdischarge point

H E ( ex,deposit ),i ( x )Equation (T16)

Sv—

Total effective dose in region, r, fromingestion of radionuclide, i, in food at
distance, x, from discharge point

H E (ing ),r ,i ( x )Equation (T18)

Sv—

COLLECTIVE DOSE CALCULATIONS
Collective effective dose calculated using:S E ( atmos ),r=x , xk +1 ) N r ( xk , xk +1 ) ⋅ H E ( atmos ),r ,i ( x ),i ( k

(T20)

Collective effective dose in region, r, in theannulus limited by xk and xk+1

S E ( atmos ),r ,i ( xk , xk +1 ) Equation (T20)

Total population in region, r, in the annulus
limited by xk and xk+1

N r ( xk , xk +1 )Dependent on region

Annular distances (the outer distance foreach annuli)( x1 , x2 , x3 , x4 , x5 ){0,100,500,1 000,1 500}

km—

Total effective dose from radionuclide, i, atdistance, x, from discharge point where x is
the midpoint between xk and xk+1

H E ( atmos ),r ,i ( x )Equation (T19)

Sv—

man Sv—see electronic
attachment 3
and [F1]

106 UNSCEAR 2016 REPORTParameter

Symbol

Value

Units

Reference

Local collective effective dose calculated using:2

S E ( atmos ),r ,i ( local ) = ∑ S E (atmos),r ,i ( xk , xk +1 )

(T21)

k =1

Local (0 km to 100 km) collective effective
dose in region, r, from radionuclide, i

S E ( atmos ),r ,i ( local )Equation (T21)

man Sv—

Total annual collective effective dose inregion, r, from radionuclide, i, in theannulus limited by xk and xk+1

S E (atmos),r ,i ( xk , xk +1 ) Equation (T20)

man Sv—

Regional component of collective effective dose
5

S E ( atmos ),r ,i ( regional ) = ∑ S E (atmos),r ,i ( xk , xk +1 )

(T22)

k =3

Regional component (100 km to 1500 km) of
collective effective dose in region, r, fromradionuclide, i

S E ( atmos ),r ,i ( regiona Equation (T21)

man Sv—

Total annual collective effective dose inregion, r, from radionuclide, i, in theannulus limited by xk and xk+1

S E (atmos),r ,i ( xk , xk +1 ) Equation (T20)

man Sv—

GLOBAL DISPERSION
Collective doses from tritium, 14C, 85Kr and 129I were calculated by PC-CREAM (incorporates global dispersion models referenced in [U6])ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 107

Table A3. Parameter and equation list: Freshwater environment
Parameter

Symbol

Value

Units

Reference

FRESHWATER DISPERSION
Activity concentration in unfiltered water calculated using:Cuw , s =Q
Friver, s(F1)

Activity concentration in unfiltered water

Cuw ,sEquation (F1)

Bq/m3

—

Annual discharge rate of radionuclide, i

Q

1

Bq/s

—

Volumetric flow rate of river of size, sFriver,s

Large: 1 000

m3/s

—

Small: 10
Activity concentration in filtered water calculated using:Cfw , s ,i =Cuw , s1 + K d ,i ⋅ α s(F2)Activity concentration of radionuclide, i, in filtered waterfrom river of size, s

Cfw , s ,iEquation (F2)

Bq/m3

—

Activity concentration in unfiltered water

Cuw ,sEquation (F1)

Bq/m3

—

Sediment–water distribution factor

K d ,i

Dependent on elementm3/t[I2, I5]

Suspended sediment load in river of size, sαsLarge: 5 × 10−4

t/m3

[O1]

Small: 2 × 10

−5ENVIRONMENTAL TRANSFERS
Activity concentrations in freshwater fish calculated using:

Cfish, s ,i =Cfw , s ,i ⋅ Bfish,i

(F3)

1 000

Activity concentration of radionuclide, i, in freshwater fish

Cfish, s ,iEquation (F3)

Bq/kg

—

Activity concentration of radionuclide, i, in filtered waterfrom river of size, s

Cfw , s ,iEquation (F2)

Bq/m3

—

Concentration factor for freshwater fish (fresh weight)

Bfish,iDependent on food andelement

Bq/kg per
Bq/L

[I5]

1 000

L/m3

—

Conversion factor
Activity concentration of HTO in fish calculated using:

Cfish, s ,HTO =Activity concentration of HTO in freshwater fish

WC f ⋅ Cfw , s ,HTO

Cfish, s ,HTO

(F4)

1 000
Equation (F4)

Bq/kg

—

108 UNSCEAR 2016 REPORTParameter

Symbol

Value

Units

Reference

Fractional water content of fish (fresh weight)

WC f

0.78

L/kg

[I5]

Activity concentration of HTO in filtered water from river ofsize, sCfw , s ,HTO

Equation (F2)

Bq/m3

—

1 000

L/m3

—

Conversion factor
Activity concentration of OBT in fish calculated using:

Cfish, s ,OBT =(1 − WC ) ⋅ WEQ ⋅ R ⋅Cff

f

fw , s ,HTO

(F5)

1 000

Activity concentration of OBT in freshwater fish

Cfish, s ,OBT

Equation (F5)

Bq/kg

—

Fractional water content of fish (fresh weight)

WC f

0.78

L/kg

[I5]

Water equivalent factor for fish

WEQ f

0.65

—

[I5]

Partition factor for fish (geometric mean, fresh weight)

Rf

6.6 × 10−4

kg/kg

[I5]

Activity concentration of HTO in filtered water from river ofsize, sCfw , s ,HTO

Equation (F2)

Bq/m3

—

1 000

L/m3

—

Conversion factor
Activity concentrations of 14C in fish calculated using:

Cfish, s ,14 C =Cfw , s ,14 C ⋅ Bfish,14 C

(F6)

1 000

Activity concentration of 14C in freshwater fish (adaptedfrom Equation (F3))

Cfish, s ,14 C

Equation (F6)

Bq/kg

—

Activity concentration of 14C in filtered water from river ofsize, sCfw , s ,14 C

Equation (F2)

Bq/m3

—

Concentration factor for 14C in freshwater fish

Bfish,14 C

5 × 104

Bq/kg per
Bq/L

[I5]

1 000

L/m3

—

Conversion factor
Deposition rate from irrigation calculated usingdirr, s= Cuw , s ⋅ I irr ⋅ Firr

(F7)

Total daily average deposition rate of radionuclide, i, from
irrigationdirr,sEquation (F7)

Bq/m2

—

Average concentration of radionuclide, i, in river water used
for irrigation

Cuw ,sEquation (F1)

Bq/m3

—

Daily irrigation rate

I irr

0.005

m3/m2

[N2]

Fraction of the year for which irrigation occurs

Firr

150
=0.41
365

—

[N2]

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 109

Parameter

Symbol

Value

Units

Reference

Activity concentration in terrestrial food (cereals, vegetables) calculated using:C f , s ,i =C f ,unit,i ⋅ dirr, s(F8)

86 400

Activity concentration of radionuclide, i, in irrigated,terrestrial food, f

C f , s ,iEquation (F8)

Bq/kg

—

Activity concentration of radionuclide, i, in terrestrial food, f,
per unit deposition rate

C f ,unit,i

PC-CREAM output

Bq/kg perBq s/m2—

Total daily average deposition rate of radionuclide, i, from
irrigationdirr,sEquation (F7)

Bq/m2

—

86 400s/d—Conversion factor
Activity concentration of HTO in soil water calculated using:Csw , s ,HTO =I irr ⋅ Cuw , s(F9)

P + 1 000 ⋅ I irr

Activity concentration of HTO in soil water irrigated fromriver of size, s

Csw , s ,HTO

Equation (F9)

Bq/L

[I4]

Daily irrigation rate (averaged over 150-day growing
season)

I irr

0.005

m3/m2

[N2]

Average concentration of radionuclide, i, in river water used
for irrigation

Cuw ,sEquation (F1)

Bq/m3

—

Daily precipitation rate (averaged over 150-day growing
season)

P

5

L/m2

[U6]

1 000

L/m3

—

Conversion factor

HTO activity concentration in terrestrial food (cereals, vegetables) calculated using:C f , s ,HTO
= WC p ⋅  RH ⋅ Cam + (1 − RH ) ⋅ Csw , s ,HTO  γ

(F10)

Activity concentration of HTO in food, f

C f , s ,HTO

Equation (F10)

Bq/kg

[I5]

Fractional water content of the plant

WC pCereal: 0.12

L/kg

[I5]

Vegetables: 0.92
Relative humidity

RH

0.7

—

[I4]

HTO concentration in air moisture

Cam

Assumed approximately
equal to Csw , s ,HTO

Bq/L

[I4]

Activity concentration of HTO in soil water

Csw , s ,HTO

Equation (F9)

Bq/L

—

Ratio of HTO vapour pressure to H2O vapour pressure

γ

0.909

—

[I5]

Activity concentration of OBT in terrestrial food (cereals, vegetables) calculated using:C f , s ,OBT =(1 − WC p ) ⋅WEQ p ⋅ R p ⋅ C f , s ,HTO
WC p

(F11)

110 UNSCEAR 2016 REPORTParameter

Symbol

Value

Units

Reference

Activity concentration of OBT in terrestrial foodC f , s ,OBT

Equation (F11)

Bq/kg

—

Fractional water content of plant (fresh weight)

WCp

Cereal: 0.12

L/kg

[I5]

L/kg

[I5]

Vegetables: 0.92
Water equivalent factor for plant

WEQp

Cereal: 0.56
Vegetables: 0.51

Partition factor

Rp

0.54

—

[I5]

Activity concentration of HTO in food, f

C f , s ,HTO

Equation (F10)

Bq/kg

[I5]

Specific activity of 14C in plant calculated using:

I ⋅C
SA
SA
CDc ⋅ irr uw , s
=
=
p
air
Fc

(F12)

Specific activity of 14C in plant

SAp

Equation (F12)Bq/(g C)

[I4]

Specific activity of 14C in air

SAair

Equation (F12)Bq/(g C)

[I4]

Canopy dilution factor

CDc

0.15

—

[I4]

Daily irrigation rate (averaged over growing season)

Iirr

0.005

m3/m2

[N2]

Activity concentration of dissolved radionuclide i in
unfiltered water

Cuw,sEquation (F1)

Bq/m3

—

Average production rate of carbon by decomposition ofcrop residues

Fc

0.66

(g C)/m2[I4]

14

C activity concentration in terrestrial food (cereals, vegetables) calculated using:C f , s ,14 C=

Cair,14C
Sair⋅ S p= SA air ⋅ S p(F13)

C activity concentration in food, f, irrigated from river ofsize, sC f , s ,14 C

Equation (F13)

Bq/kg

[I4]

Specific activity of 14C in air

SAair

Equation (F12)Bq/(g C)

[I4]

Concentration of stable carbon in the plant

Sp

Cereal: 3.9 × 102(g C)/kg[I5]

14

Vegetables: 30

INDIVIDUAL DOSE CALCULATIONS
Individual dose from ingestion of drinking water calculated using

H E ( water ), s ,i = Cuw , s ⋅ Ding,i ⋅ I water ⋅ FWT ⋅ Priver(F14)

Annual effective dose from the ingestion of radionuclide, i,in drinking water

H E ( water ), s ,iEquation (F14)

Sv—

Average activity concentration of radionuclide, i, in water
assuming complete mixing

Cuw ,sEquation (F1)

Bq/m3

—

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 111

Parameter

Symbol

Value

Units

Reference

Dose coefficient for ingestion of radionuclide, iDing,i

Dependent on radionuclide

Sv/Bq

[I14]

Amount of water ingested in one year

I water

0.5

m3

[U6]

Drinking water treatment removal factor

FWT

Dependent on element

—

[B7]

Fraction of individual water intake from river

Priver1——

Dose from external exposure from freshwater sediments calculated using:

H E ( ex,riverbank ), s=
Csed, s ,i ⋅ rsed ⋅ tsed ⋅ Fgeom ⋅ Dex,deposit ⋅ Oriverbank
,i

(F15)

Annual effective dose from radionuclide, i, from occupancy
on river banks

H E ( ex,riverbank ), s ,iEquation (F15)

Sv—

Activity concentration of radionuclide, i, in freshwater
riverbank sediment based on discharge rate of 1 Bq/s
(assumed to be the same as bed sediment) (dry weight)

Csed, s ,i

PC-CREAM output

Bq/kg

—

Density of riverbank sediment

ρsed

1 200

kg/m3

[I2]

Thickness of riverbank sediment

tsed

0.05

m

[I2]

Geometry factor for river banks for external dosecoefficientsFgeom

0.2 (rounded)

—

[A2]

Integrated dose coefficient for irradiation from surface
deposits

Dex,deposit

Dependent on radionuclide

Sv/s per
Bq/m2

[E2]

River bank occupancy

Oriverbank

1.8 × 105 (50 hours)sAssumed value

Ingestion of freshwater fish

H E (ing,fish ), s ,r ,i = Cfish, s ,i ⋅ Ding,i ⋅ Flocal ⋅ I fish,r(F16)

Annual effective dose from radionuclide, i, from ingestionof freshwater fish in region, r

H E (ing,fish ), s ,r ,iEquation (F16)

Sv—

Average activity concentration of radionuclide, i, infreshwater fish

Cfish, s ,iEquations (F3), (F4), (F5),
(F6)

Bq/kg

—

Dose coefficient for adults for ingestion

Ding,i

Dependent on radionuclide

Sv/Bq

[I14]

Fraction of food that is locally produced

Flocal0.25—

See para.11[U12]

Amount of freshwater fish ingested in a year in region, rI fish,r

Dependent on region

kg

[F3]

Dose from ingestion of irrigated, terrestrial foods calculated using:

H E (ing,irr ), f , s ,r ,i = C f , s ,i ⋅ Ding,i ⋅ Flocal ⋅ Fspray, s ,r ⋅ I f ,rAnnual effective dose from the ingestion of irrigated,terrestrial food, f, containing radionuclide, i, in region, r

H E (ing,irr ), f , s ,r ,iEquation (F17)

(F17)
Sv

—

112 UNSCEAR 2016 REPORTParameter

Symbol

Value

Units

Reference

Activity concentration of radionuclide, i, in irrigated,terrestrial food, f

C f , s ,iEquations (F8), (F10), (F11),
(F13)

Bq/kg

—

Adult dose coefficient for ingestion

Ding,i

Dependent on radionuclide

Sv/Bq

[I14]

Fraction of food that is locally produced

Flocal0.25—

See para.11[U12]

Fraction of food that is spray irrigatedFspray, s ,rSmall rivers: 0

—

[I7]

kg

[W2]

Large rivers: dependent on
regionAmount of food, f, ingested in a year in region, rDependent on region andfood

I f ,rTotal effective dose from ingestion of irrigated, terrestrial foods calculated using:

H E (ing,irr ), s ,r ,i = ∑ H E (ing,irr ), f , s ,r ,i

(F18)

f

Total effective dose in region, r, from ingestion of irrigated,terrestrial food containing radionuclide, i

H E (ing,irr ), s ,r ,iEquation (F18)

Sv—

Annual effective dose from the ingestion of irrigated,terrestrial food, f, containing radionuclide i in region, r

H E (ing,irr ), f , s ,r ,iEquation (F17)

Sv—

Total individual dose from freshwater pathways calculated using:

H E ( fw ), s ,r ,i = H E ( water ), s ,i + H E ( ex,riverbank ), s ,i + H E ( fish ), s ,r ,i + H E (ing,irr ), s ,r ,i(F19)

Total individual dose from freshwater pathways

H E ( fw ), s ,r ,iEquation (F19)

Sv—

Annual effective dose from the ingestion of radionuclide, i,in drinking water

H E ( water ), s ,iEquation (F14)

Sv—

Annual effective dose from radionuclide, i, from occupancy
on river banks

H E ( ex,riverbank ), s ,iEquation (F15)

Sv—

Annual effective dose from radionuclide, i, from ingestion of
freshwater fish in region, r

H E ( fish ), s ,r ,iEquation (F16)

Sv—

Total effective dose in region, r, from ingestion of irrigated,terrestrial food containing radionuclide, i

H E (ing,irr ), s ,r ,iEquation (F18)

Sv—

COLLECTIVE DOSE CALCULATIONS
Volumetric abstraction rate of water calculated using:

Virr,surface,=
Friver, s ⋅ Firr,surface,r
s ,r

(F20)

Volumetric abstraction rate of surface water from river ofsize, s, in region, rVirr,surface, s ,rEquation (F20)

m3/s

—

Volumetric flow rate of river of size, sFriver,s

Large: 1 000

m3/s

—

m2 a/m3[F4]

Small: 10
Fraction of surface water withdrawn for irrigation in region, rFirr,surface,rDependent on region

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 113

Parameter

Symbol

Value

Units

Reference

Area irrigated by abstracted water calculated using:Airr,surface, s ,r = 3.15 × 107 ⋅ Airr,unit,r ⋅ Virr,surface, s ,r(F21)

Area irrigated by water abstracted from river of size, s, inregion, r

Airr,surface, s ,rEquation (F21)m2—

Area irrigated per unit volume water withdrawn in a year inregion, r

Airr,unit,rDependent on region

m2 a/m3[F4]

Volumetric abstraction rate of surface water from river ofsize, s, in region, rVirr,surface, s ,rEquation (F20)

m3/s

—

3.15 × 107 (rounded)s/aConversion factor
Irrigation rate calculated using:C ⋅Vdirr,surface, s ,r = uw , s irr,surface, s ,rAirr,surface, s ,r(F22)

Irrigation rate, assumed to be the same volume of waterwith the same activity concentration abstracted from riverof size, s, in region, rdirr,surface, s ,rEquation (F22)

Bq s/m2—

Activity concentration in unfiltered water

Cuw ,sEquation (F1)

Bq/m3

—

Volumetric abstraction rate of surface water from river ofsize, s, in region, rVirr,surface, s ,rEquation (F20)

m3/s

—

Area irrigated by water abstracted from river of size, s, inregion, r

Airr,surface, s ,rEquation (F21)m2—

Activity concentration in irrigated, terrestrial foods calculated using:C=
Ccereal,unit,i ⋅ dirr,surface, s ,rcereal, s , r ,i(F23)

Activity concentration of radionuclide, i, in terrestrial food(represented by cereal) irrigated from river of size, s, inregion, r

Ccereal, s ,r ,iEquation (F23)

Bq/kg

Activity concentration of radionuclide, i, in terrestrial food,f, per unit deposition rate

Ccereal,unit,i

PC-CREAM output

Bq/kg perBq s/m2

Irrigation ratedirr,surface, s ,rEquation (F22)

Bq s/m2——

Activity concentration of 14C in cereals calculated using:Cair,14C

Ccereal, s ,r ,14=C

Sair

SA air
⋅ S=
p
(F24)

Cuw , s
CDc ⋅
=
⋅ Sp
365 ⋅ Airr,unit,r ⋅ Fc
Activity concentration of 14C in food, f, irrigated from river ofsize, sCcereal, s ,r ,14 C

Equation (F24)

Bq/kg

Canopy dilution factor

CDc

0.15

—

Equations(F13) & (F12)[I4]

114 UNSCEAR 2016 REPORTParameter

Symbol

Value

Units

Reference

Activity concentration in unfiltered water

Cuw ,sEquation (F1)

Bq/m3

—

Area irrigated per unit volume in region, rAirr,unit,rDependent on region

m2 a/m3[F4]

Average daily production rate of carbon by decompositionof crop residues

Fc

0.66

(g C)/m2[I4]

Concentration of stable carbon in the plantSp

3.9 × 102(g C)/kg

[I5]

365

d/a

—

Conversion factor
Collective dose from ingestion of irrigated, terrestrial food calculated using:S E (ing,irr ), s ,r ,=Ccereal, s ,r ,i ⋅ Ycereal,r ⋅ Fspray, s ,r ⋅ Airr,surface, s ,r ⋅ Ding,i
i

(F25)

Collective dose from consumption of food (represented bymajor food group, cereal) in region, r

S E (ing,irr ), s ,r ,iEquation (F25)man Sv—

Activity concentration of radionuclide, i, in terrestrial food(represented by cereal) irrigated from river of size, s, inregion, r

Ccereal, s ,r ,iEquations (F23), (F10), (F11),
(F24)

Bq/kg

—

Yield per unit area of cereal in region, r

Ycereal,rDependent on region

kg/m3

[F1]

Fraction of food that is spray irrigatedFspray, s ,rSmall rivers: 0

—

[I7]

Large rivers: dependent on
regionArea irrigated by water abstracted from river of size, s, inregion, r

Airr,surface, s ,rEquation (F21)m2—

Adult dose coefficient for ingestion

Ding,i

Dependent on radionuclide

Sv/Bq

[I14]

Collective dose from drinking water calculated using:S E ( water ), s ,i = Cuw ,i ⋅ Ding,i ⋅ Friver, s ⋅ Tdw ⋅ Amun ⋅ Fdw ⋅ FWT(F26)

Collective dose from drinking water for river of size, s

S E ( water ), s ,iEquation (F26)

man Sv—

Activity concentration of dissolved radionuclide, i, in
unfiltered water

Cuw ,iEquation (F1)

Bq/m3

—

Dose coefficient for ingestion of radionuclide, iDing,i

Dependent on radionuclide

Sv/ q

[I14]

Volumetric flow rate of river of size, sFriver,s

Large: 1 000

m3/s

—

Small: 10
Duration of abstraction

Tdw

3.15 × 107 (rounded)

S—

Abstraction fraction for municipal water

Amun

0.01

—

—

Fraction of municipal water that is drunk

Fdw

0.01

—

—

Drinking water treatment removal factor

FWT

Dependent on element

—

[B7]

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 115

Parameter

Symbol

Value

Units

Reference

Collective dose from freshwater fish ingestion calculated using:S E (ing,fish ), s ,i = Cfish, s ,i ⋅ Ding,i ⋅ Yfish, s ⋅ Lriver, s(F27)Collective dose from ingestion of freshwater fish from riverof size, s

S E (ing,fish ), s ,iEquation (F27)

man Sv—

Activity concentration of radionuclide, i, in filtered waterfrom river of size, s

Cfish, s ,iEquations (F3), (F4), (F5),
(F6)

Bq/m3

—

Dose coefficient for ingestion of radionuclide, iDing,i

Dependent on radionuclide

Sv/Bq

[I14]

Representative freshwater fish catch in river of size, s

Yfish,sLarge: 104

kg/km

[F1]

km

[I2]

Small: 10

3

Length of river of size, s

Lriver,s

Large: 5 × 102
Small: 1 × 102

Total collective dose from all freshwater pathways calculated using:S E ( fw ), s ,r ,i = S E (ing,irr ), s ,r ,i + S E ( water ), s ,i + S E (ing,fish ), s ,i(F28)

Total collective dose in region, r, from discharges into riverof size, s

S E ( fw ), s ,r ,iEquation (F28)

man Sv—

Collective dose from consumption of food (represented bymajor food group, cereal) in region, r

S E (ing,irr ), s ,r ,iEquation (F25)man Sv—

Collective dose from drinking water for river of size, s

S E ( water ), s ,iEquation (F26)

man Sv—

Collective dose from ingestion of freshwater fish of size, s

S E (ing,fish ), s ,iEquation (F27)

man Sv—SE(fish),s,i

GLOBAL DISPERSION
Collective doses from tritium, 14C, 85Kr and 129I were calculated by PC-CREAM (incorporates global dispersion models referenced in [U6])116 UNSCEAR 2016 REPORT

Table A4. Parameter and equation list: Marine environment
Parameter

Symbol

Value

Units

Reference

MARINE DISPERSION
Sedimentation decay constant

λs ,c ,i =K d , c ,i ⋅ S c(M38)

hc ⋅ (1 + K d ,c ,i ⋅ α c )Sedimentation decay constant for compartment, cls,c,iEquation (M38)

a−1

[C2]

Marine sediment distribution factor for compartment
c and radionuclide, i

Kd,c,i

Dependent on radionuclide

m3 t−1[I3]

Sedimentation rate in compartment, c

Sc

Local: 1.0 × 10−4

t m−2 a−1(see section III.C of main
text)

m

(see section III.C of main
text)

t m−3

(see section III.C of maintext)

Regional: 1.0 × 10−5
Depth of compartment, chcLocal: 10
Regional: 1000

Suspended sediment load in compartment, c

αcLocal: 2.0 × 10−4
Regional: 1.0 × 10−7

Flow rate constant, local to regionalll ,r =

∆Vl ,rVl(M39)

Flow rate constant, local compartment to regional
compartment

ll,rEquation (M39)a−1(see section III.C of maintext)

Volumetric exchange between regional compartment
and global oceans∆Vl ,r2 × 1010

m3 a−1(see section III.C of maintext)

Volume of local compartment

Vl

109

m3

(see section III.C of maintext)

Flow rate constant, regional compartment to global oceans

λr , g =∆Vr , g

(M40)

Vr

Flow rate constant, regional compartment to global
oceanslr,gEquation (M40)a−1(see section III.C of maintext)

Volumetric exchange between local and regional
compartments∆Vr , g1015

m3 a−1(see section III.C of maintext)

Volume of regional compartment

Vl

1015

m3

(see section III.C of maintext)

Combined decay constant, local to regionalΛ l ,i = li + ls ,l ,i + ll ,rCombined decay constant for local compartment andradionuclide, iΛ l ,iEquation (M41)

(M41)
a−1

[C2]

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 117

Parameter

Symbol

Value

Units

ReferenceRadioactive decay constant for radionuclide, iliDependent on radionuclide

a−1

[I9]

Sedimentation decay constant for local compartment

ls,l,iEquation (M38)

a−1

[C2]

Flow rate constant, local compartment to regional
compartment

ll,rEquation (M39)a−1(see section III.C of maintext)

Combined decay constant, regional compartment to global oceansΛ r ,i = λλλi + s , r ,i + r , g(M42)

Combined decay constant for regional compartment
and radionuclide, iΛ r ,iEquation (M42)

a−1

[C2]

Radioactive decay constant for radionuclide, iliDependent on radionuclide

a−1

[I9]

Sedimentation decay constant for regional
compartment

ls,r,iEquation (M38)

a−1

[C2]

Flow rate constant, regional compartment to global
oceanslr,gEquation (M40)a−1(see section III.C of main
text)

Activity in local compartment

A=l ,i ( t )

(

Q
−Λ ⋅t
1 − e l ,rΛ l ,r)(M43)

Activity of radionuclide, i, in local compartment attime, tAl ,i ( t )Equation (M43)

Bq

[C2]

Discharge rate

Q

3.15 × 107 (rounded)

Bq a−1= 1 Bq s−1

Combined decay constant for local compartment andradionuclide, iΛ l ,rEquation (M41)

a−1

[C2]

Duration of discharge

t

100

a

—

Activity in regional compartment=
Ar ,i ( t )ll ,r ⋅ Al ,i ( t )Λ l ,r(1 − e−Λ l ,r ⋅t

)

(M44)

Activity of radionuclide, i, in regional compartment attime, t

Ar ,i ( t )Equation (M44)

Bq

[C2]

Rate constant for movement of water from local
compartment to regional compartment

ll,r

20

a−1

(see section III.C of main
text)

Activity of radionuclide, i, in local compartment attime, tAl ,i ( t )Equation (M43)

Bq

[C2]

Combined decay constant for regional compartment
and radionuclide, iΛ l ,rEquation (M41)

a−1

[C2]

Duration of discharge

t

100

a

—

118 UNSCEAR 2016 REPORTParameter

Symbol

Value

Units

Reference

Activity concentration in water

Cc ,i ( t ) =Ac ,i ( t )
Vc

(M45)

Activity concentration of radionuclide, i, in water incompartment, c, at time, t

Cc ,i ( t )Equation (M45)

Bq m−3—

Activity of radionuclide, i, in compartment, c, at time,

Ac ,i ( t )Equations (M43), (M44)

Bq

[C2]

Vc

Local: 1.0 × 109

m3(see section III.C of maintext)tVolume of compartment, cRegional: 1.0 × 10

15ENVIRONMENTAL TRANSFERS
Activity concentrations in marine foods

C f , c ,i =Cc ,i ⋅ B f ,i

(M46)

1 000

Activity concentration of radionuclide, i, in marine
food, f, in compartment, c

Cf,c,iEquation (M46)

Bq kg−1—

Activity concentration of radionuclide, i, in water attime, tCc,iEquation (M45)

Bq m−3—

Bioaccumulation factor for marine food, f,
concentration in marine food (fresh weight) divided
by concentration in seawater

Bf,iDependent on food andelement

L kg−1

[I3]

1 000

L m−3

—

Conversion factor
Activity concentration in beach sediment

Cbeach,i = 0.1⋅ tsed ⋅ ρsed ⋅ K d ,local,i ⋅ Clocal,i(M47)

Activity concentration per unit area in beachsediment

Cbeach,iEquation (M47)

Bq m−2—

Kd adjustment factor for coarser grains (i.e. sand

—

0.1

—

[S7]

Average thickness of beach sediment

tsed

0.05

m

[I2]

Average density of beach sediment

rsed

1.2

t m−3

[I2]

Marine sediment distribution factor in local
compartment for radionuclide, iKd,local,iDependent on radionuclide

m3 t−1(L kg−1)[I3]

Activity concentration of radionuclide, i, in water

Clocal,i

Equation (M45)

Bq m−3—rather than fine bed sediment)

INDIVIDUAL DOSE CALCULATIONS
External dose from beach occupancy

H E ( ex,beach ),i = Cbeach,i ⋅ Fgeom ⋅ Dex,deposit,i ⋅ Obeach(M48)

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 119

Parameter

Symbol

Value

Units

Reference

Effective dose from radionuclide, i, from beachoccupancy

H E ( ex,beach ),iEquation (M48)

Sv—

Activity concentration of radionuclide, i, in beachsediment

Cbeach,iEquation (M47)

Bq m−2—

Geometry factor for beaches for external dosecoefficients

Fgeom

0.5

—

[E2]

Dose coefficient for irradiation from surface deposits

Dex,deposit,iDependent on radionuclide

Sv s−1 Bq−1
m2

[E2]

Beach occupancy (typical rather than critical)

Obeach

3.6 × 105s

[R1]

Ingestion of marine food

H E (ing,marine ), f ,r=,i(C

f ,local,i⋅ f f ,local + C f ,regional,i ⋅ f f ,regional ) ⋅ Ding,i ⋅ I f ,rAnnual effective dose from radionuclide, i, fromingestion of freshwater fish in region, r

H E (ing,marine ), f ,r Equation (M49)

Average activity concentration of radionuclide, i, infood, f, in the local marine compartment

C f ,local,iFraction of food, f, that is caught in the local marine
compartment

f f ,local

(M49)

Sv

—

Equation (M46)

Bq kg−1—

Fish: 0.25—

Crustaceans: 1.0
Molluscs: 1.0

Fish: consistent withfreshwater andterrestrial
Crustaceans andmolluscs: [S7]Average activity concentration of radionuclide, i, infood, f, in the regional marine compartment

C f ,regional,iEquation (M46)

Bq kg−1

Fraction of food, f, that is caught in the regional
marine compartment

f f ,regional

Fish: 0.75—

Crustaceans: 0.0
Molluscs: 0.0—

Fish: consistent withfreshwater andterrestrial
environments
Crustaceans & molluscs:[S7]

Adult dose coefficient for ingestion

Ding,i

Dependent on radionuclide

Sv Bq−1[I14]

Amount of marine food, f, ingested in a year inregion, r

If,rDependent on region

kg

[W2]

Total effective dose from ingestion of marine foods

H E (ing,marine ),r ,i = ∑ H E (ing,marine), f ,r ,i

(M50)

f

Total effective dose in region, r, from ingestion ofmarine food containing radionuclide, i

H E (ing,marine ),r ,i Equation (M50)

Sv—

Effective dose from the ingestion of marine food fcontaining radionuclide, i, in region, r

H E (ing,marine ), f ,r Equation (M49)

Sv—

120 UNSCEAR 2016 REPORTParameter

Symbol

Value

Units

Reference

Total individual dose from marine pathwaysH=
H E ( ex,beach ),i + H E (ing,marine ),r ,iE ( marine ), r ,iTotal individual dose from marine pathways in region,(M51)

H E ( marine),r ,iEquation (M51)

Sv—

Effective dose from beach occupancy

H E ( ex,beach ),iEquation (M48)

Sv—

Effective dose from the ingestion of radionuclide, i, in
marine foods in region, r

H E (ing,marine ),r ,i Equation (M50)

Sv

—

r

COLLECTIVE DOSE CALCULATIONS
Collective dose from ingestion of marine fish

S E (ing,marine ),fish,c ,=Cfish,c ,i ⋅ Yfish,r ⋅ Fed,fish ⋅ Ac ⋅ Ding,i
r ,i

(M52)

Collective dose from ingestion of marine fish forradionuclide, i

S E (ing,marine ),fish, Equation (M52)

man Sv—

Activity concentration of radionuclide, i, in fish incompartment, cCfish,c ,iEquation (M46)

Bq kg−1—

Fish catch in 1 year in region, r

Yfish,rDependent on region

kg km−2[F2]

Edible fraction of catch

Fed,fish

0.5

—

[S6]

Area of compartment, cAc

Local: 1 × 102

km2(see section III.C of maintext)

Sv Bq−1[I14]

Regional: 1 × 10

6Adults dose coefficient for ingestion

Ding,i

Dependent on radionuclide

Collective dose from ingestion of shellfish

S E (ing,marine ), f ,c ,r ,=i C f ,c ,i ⋅ Y f ,r ⋅ Fed, f ⋅ Lc ⋅ Ding,iCollective dose from ingestion of marine food, f,(where f ∈ {molluscs,crustaceans} ) in

(M53)

S E (ing,marine ), f ,c ,r Equation (M53)

man Sv—

compartment, c, for radionuclide, iActivity concentration of radionuclide, i, in marine
food, f, in compartment, c

C f , c ,iEquation (M46)

Bq kg−1—

Harvest of marine food, f, in 1 year per unit length ofcoastline in region, rY f ,rDependent on region

kg km−1[F2]

Edible fraction of catch for marine food, fFed, fCrustaceans: 0.35

—

[S6]

km

(see section III.C of maintext)

Molluscs: 0.15
Length of coastline in compartment, c

Lc

Local: 10
Regional: 1 000

Adults dose coefficient for ingestion

Ding,i

Dependent on radionuclide

Sv Bq−1[I14]

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 121

Parameter

Symbol

Value

Units

Reference

Total collective dose from all marine pathways

S E ( marine ),c ,r ,i = ∑ S E (ing,marine ), f ,c ,r ,i

(M54)

f

Total collective dose in compartment, c

S E ( marine ),c ,r ,iEquation (M54)

man Sv—

Collective dose from consumption of food, f, incompartment, c

S E (ing,marine ), f ,c ,r Equations (M52) and (M53)

man Sv—

GLOBAL DISPERSION
Collective doses from tritium, 14C, 85Kr and 129I calculated by PC-CREAM (incorporates global dispersion models referenced in UNSCEAR 2000
Report [U6])122 UNSCEAR 2016 REPORT

Table A5. Treatment of progeny for atmospheric and terrestrial pathways
Half-lives are given in brackets to two significant figures [I9]Discharged
parentProgeny

Exposure pathway
Plume (external irradiationand inhalation)

Deposited material(external irradiation)

Food (ingestion)

Y (64 h)Dose coefficient for
inhalation of parent
accounts for progeny.
Activity concentration ofprogeny assumed equal toparent for estimatingexternal exposure

Deposition and dose
rates calculated forparent and progeny
separately

Dose coefficient for ingestion ofparent accounts for progeny

Sr (29 a)

90

106

Ru (370 d)

106

Rh (30 s)

Secular equilibrium
assumed. Dose rates
calculated for both parent
and progeny

Deposition and dose
rates calculated forparent and progeny
separately

Dose coefficient for ingestion ofparent accounts for progeny

135

Xe (9.1h)135

Cs (2.3 × 106 a)

Activity concentrations ofprogeny in air calculated
explicitly based on transit
time

Deposition and dose rate
calculated for progeny

Activity concentration calculated from
deposition rate for progeny138

Xe (14 m)

138

Cs (32 min)

Activity concentration ofprogeny in air calculated
explicitly based on transit
time

Deposition and dose rate
calculated for progeny

Activity concentration calculated from
deposition rate for progeny137

Cs (30 a)137m

Ba (2.6 min)Dose coefficient for
inhalation of parent
accounts for progeny.
Activity concentration ofprogeny calculated at yieldof 94.6% for estimatingexternal exposure

Deposition and dose
rates calculated forparent and progeny
separately

Dose coefficient for ingestion ofparent accounts for progeny

210

Pb (22 a)210

Bi (5.0 d)

210

Po (140 d)

Ingrowth insignificant
during plume transit;
progeny not considered

Negligible dose fromprogeny following
ingrowth after
deposition of 210Pb; notconsidered

Ingrowth following deposition of 210Pb
to soil and plant, but activity
concentration of 210Po in terrestrial
foods significantly lower than fordeposition of 210Po; subsequent
small doses compared to those from
210
Pb; ingrowth not considered218

Po (3.1 min)214

Pb (27 min)214

Bi (20 min)214

Po (164 μs)

210

Pb (22 a)

Negligible contributions
to dose from
deposition of progeny
compared with dosefrom inhalation; notconsidered

Because of time taken for ingrowth
and reduced activities of long-lived
progeny compared to 222Rn,
negligible doses from ingestion ofprogeny in terrestrial foods; notconsidered210

Bi (5.0 d)

210

Po (140 d)

Dose coefficient for
inhalation of 222Rn includes
contribution from four
immediate short-lived
progeny [U9]. Longer lived
progeny (210Pb, 210Bi and
210Po) will not grow induring plume transit andso not considered

90

222

Rn (3.8 d)

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 123

Discharged
parent226

232

Ra (1 600 a)

Th (1.4 × 1010 a)Progeny222

Rn (3.8 d)

218

Po (3.1 min)214

Pb (27 min)214

Bi (20 min)214

Po (164 μs)

210

Pb (22a)210

Bi (5.0 d)

210

Po (140 d)

228

Ra (5.8 a)

228

Ac (6.1 h)

228

Th (1.9 a)

212

Pb (11 h)

234

U (2.4 × 105 a)

230

Th (7.7 × 104 a)

238

U (4.5 × 109 a)

234

Th (24 d)
Pa (1.2 min)

234m

239

Pu (2.4 × 104 a)

235

U (7.0 × 108 a)

240

Pu (6 500 a)

236

U (2.3 × 107 a)

241

Am (430 a)

237

Np (2.1 × 106 a)

Exposure pathway
Plume (external irradiationand inhalation)

Deposited material(external irradiation)

Food (ingestion)

Ingrowth insignificant
during plume transit;
progeny not considered

Ingrowth following
deposition of 226Ra, butradon gas largely
emitted to atmosphereand further diluted.
Small doses fromprogeny comparedwith parent; notincluded

Ingrowth following deposition of226
Ra, but radon gas largely emitted toatmosphere and further diluted. Smalldoses from progeny compared withparent; not included

Secular equilibrium
assumed; activity
concentration in air
assumed equal to parent

Deposition rate
calculated for eachprogeny in secularequilibrium

Activity concentration of parent in air
used as proxy for secular equilibrium
in soil after 100 years. Only 228Ra,
228
Th and 212Pb included

Parent half-life sufficiently long that ingrowth not relevant at 100 years
Secular equilibrium
assumed; activity
concentration in air
assumed equal to parent

Deposition rate
calculated for eachprogeny

Activity concentration of parent in air
used as proxy for secular equilibrium
in soil after 100 years. Only 234Th
includedParent half-life sufficiently long that ingrowth not relevant at 100 years

124 UNSCEAR 2016 REPORT

Table A6. Treatment of progeny for freshwater pathways
Half-lives are given in brackets to two significant figures [I9]Discharged parent

Progeny

Exposure pathway
Riverbank occupancy90

Sr (29 a)90

Ru (370 d)

106

106

Y (64 h)Secular equilibrium
assumeda

Rh (30 s)

Secular equilibrium
assumeda137

Cs (30 a)

137m

210

Pb (22 a)210

Bi (5.0 d)

210

Po (140 d)

222

Rn (3.8 d)

218

Po (3.1 min)214

Pb (27 min)214

Bi (20 min)214

Po (164 μs)

210

Pb (22 a)210

Bi (5.0 d)

210

Po (140 d)

228

Ra (5.8 a)

228

Ac (6.1 h)

228

Th (1.9 a)

212

Pb (11 h)

226

232

Ra (1 600 a)

Th (1.4 × 1010 a)Ba (2.6 min)234

U (2.4 × 105 a)

230

Th (7.7 × 104 a)

238

U (4.5 × 109 a)

234

Th (24 d)
Pa (1.2 min)

234m

a

239

Pu (2.4 × 104 a)

235

U (7.0 × 108 a)

240

Pu (6 500 a)

236

U (2.3 × 107 a)

241

Am (430 a)

237

Np (2.1 × 106 a)

Freshwater fish

Drinking water

Irrigated food

Dose coefficient for ingestion of parent accounts for progeny

Ba activity
concentration in filteredwater calculated at yieldof 94.6% a137m

Negligible doses from
ingrowth compared withtotal dose for 210Pb; notconsidered
Rn gas mainly emitted toatmosphere;
contributions from
subsequent progeny
omitted222

Secular equilibrium
assumeda

Ingrowth limited during transit time in freshwater body;doses from progeny not included

Secular equilibrium
assumed in water a

Secular equilibrium
assumed in
untreated water a

Deposition rate ofparent used asproxy for secularequilibrium in soil
after 100 years.
Only 228Ra, 228Th
and 212Pb included

Parent half-life sufficiently long that ingrowth not relevant at 100 years
Secular equilibrium
assumeda

Dose rates calculated from parent and each progeny separately.

Secular equilibrium
assumed in water a

Secular equilibrium
assumed in
untreated water a

Deposition rate ofparent used asproxy for secularequilibrium in soil
after 100 years.
Only 234Th included

Parent half-life sufficiently long that ingrowth not relevant at 100 years

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 125

Table A7. Treatment of progeny for marine pathways
Half-lives are given in brackets to two significant figures [I9]Discharged
parentProgeny

Exposure pathway
Beach occupancy

Sr (29 a)

90

106

Ru (370 d)

106

137

Cs (30 a)

137m

210

Pb (22 a)210

Bi (5.0 d)

210

Po (140 d)

222

Rn (3.8 d)

218

Po (3.1 min)214

Pb (27 min)214

Bi (20 m)

214

Po (164 μs)

210

Pb (22 a)210

Bi (5.0 d)

210

Po (140 d)

228

Ra (5.8 a)

228

Ac (6.1 h)

228

Th (1.9 a)

212

Pb (11 h)

90

Marine foodY (64 h)Secular equilibrium assumeda226

Ra (1 600 a)Th(1.4 × 1010 a)

232

Dose coefficient for ingestion of parent accounts forprogeny

Rh (30 s)
Ba (2.6 min)234

U (2.4 × 105 a)

230

Th (7.7 × 104 a)

238

U (4.5 × 109 a)

234

Th (24 d)

Ba activity in local/regional compartment
and on beach calculated at yield of 94.6%a137m

Secular equilibrium assumeda

Secular equilibrium assumed in water;
bioaccumulation factors for parent and progeny
explicitly applied

Most radon is assumed not emitted toatmosphere; secular equilibrium assumed a

Secular equilibrium assumed in water;
bioaccumulation factors for parent and progeny
explicitly applied

Secular equilibrium assumed a

Secular equilibrium assumed in water;
bioaccumulation factors for parent and progeny
explicitly applied

Parent half-life sufficiently long that ingrowth not relevant at 100 years
Secular equilibrium assumed aPa (1.2 min)

234m

a

239

Pu (2.4 × 104 a)

235

U (7.0 × 108 a)

240

Pu (6 500 a)

236

U (2.3 × 107 a)

241

Am (430 a)

237

Np (2.1 × 106 a)

Secular equilibrium assumed in water;
bioaccumulation factors for parent and progeny
explicitly applied

Parent half-life sufficiently long that ingrowth not relevant at 100 years

Dose rates from external exposure calculated from parent and each progeny separately.

ANNEX A: METHODOLOGY FOR ESTIMATING PUBLIC EXPOSURES DUE TO RADIOACTIVE DISCHARGES 127

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ANNEX B
RADIATION EXPOSURES FROMELECTRICITY GENERATION

135

Corrigendum to Sales No. E.17.IX.1

1 August 2017

Sources, Effects and Risks of Ionizing Radiation: United
Nations Scientific Committee on the Effects of Atomic
Radiation 2016 Report to the General Assembly, with
Scientific Annexes — Scientific Annexes A, B, C and D
Corrigendum
1. Annex B (Radiation exposures from electricity generation), page 142, table 1,
footnote a
The footnote should read
The value for the nuclear fuel cycle was 5.7 man Sv/(GW a); for globally-circulating
radionuclides after 100 years the value was 12 man Sv/(GW a); for uranium mine andmill tailings after 100 years of radon releases, the value was 0.25 man Sv/(GW a).
2. Annex B (Radiation exposures from electricity generation), page 142, table 1,
footnote e
The footnote should read
The value for the nuclear fuel cycle was 3.0 man Sv/(GW a); for globally-circulating
radionuclides after 10,000 years the value was 50 man Sv/(GW a); for uranium mineand mill tailings after 10,000 years of radon releases, the value was 150 man Sv/(GW a).

V.17-05540 (E)*1705540*

Corrigendum to Sales No. E.17.IX.1

15 December 2020

Sources, Effects and Risks of Ionizing Radiation: United
Nations Scientific Committee on the Effects of Atomic
Radiation 2016 Report to the General Assembly, with
Scientific Annexes
Corrigendum
Annex B (Radiation exposures from electricity generation), page 164, table 14
Column headed “Asia and Pacific”, subsection entitled “Aquatic discharges”
In the entry for Power plants, for 1.5 × 10−4 read 8.4 × 10−5
Column headed “Europe”, subsection entitled “Aquatic discharges”
In the entry for Power plants, for 1.2 × 10−3 read 6.6 × 10−4
In the entry for Reprocessing, for 6.9 × 10−4 read 6.0 × 10−4
Column headed “Latin America”, subsection entitled “Aquatic discharges”
In the entry for Power plants, for 1.5 × 10−4 read 1.0 × 10−4
Column headed “North America”, subsection entitled “Aquatic discharges”
In the entry for Power plants, for 5.0 × 10−5 read 4.1 × 10−5
Column headed “Asia and Pacific”, subsection entitled “Total for all discharges”
In the entry for Power plants, for 2.0 × 10−4 read 1.3 × 10−4
Column headed “Europe”, subsection entitled “Total for all discharges”
In the entry for Power plants, for 1.3 × 10−3 read 7.3 × 10−4
In the entry for Reprocessing, for 7.5 × 10−4 read 6.6 × 10−4
Column headed “Latin America”, subsection entitled “Total for all discharges”
In the entry for Power plants, for 2.6 × 10−4 read 2.1 × 10−4
Column headed “North America”, subsection entitled “Total for all discharges”
In the entry for Power plants, for 1.1 × 10−4 read 9.9 × 10−5

V.20-07527 (E)*2007527*

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 137

CONTENTSI.

INTRODUCTION ................................................................................................................................. 139
A. Scope............................................................................................................................................. 140

II. BACKGROUND ................................................................................................................................... 141
A. Global trends in electricity generation technology ..................................................... 142
B. Electricity generation worldwide in 2010 ........................................................................ 144
III. ASSESSMENT APPROACH AND END POINTS .......................................................................... 145
A. Recapitulation of key features of the methodology for estimating
public exposures due to radioactive discharges ........................................................... 147
IV. RADIATION EXPOSURES ARISING FROM ELECTRICITY GENERATION BY NUCLEAR
POWER ......................................................................................................................................................... 151
A. Introduction................................................................................................................................ 151
B. Public exposure ......................................................................................................................... 152
C. Occupational exposure for the nuclear fuel cycle ........................................................ 174
D. Decommissioning..................................................................................................................... 178
V. RADIATION EXPOSURES ARISING FROM ELECTRICITY GENERATION BY COMBUSTION
OF FOSSIL FUELS ...................................................................................................................................... 182
A. Coal ................................................................................................................................................ 182
B. Natural gas .................................................................................................................................. 194
C. Oil ................................................................................................................................................... 197
D. Comparisons of exposures from fossil-fuel electricity generation ......................... 199
VI. RADIATION EXPOSURES ARISING FROM ELECTRICITY GENERATION FROMGEOTHERMAL ENERGY........................................................................................................................... 199
VII. ASSESSING DOSES FROM THE CONSTRUCTION PHASE OF ELECTRICITY-GENERATINGTECHNOLOGIES ........................................................................................................................................ 201
VIII. COMPARISON OF RADIATION EXPOSURES FROM ELECTRICITY-GENERATINGTECHNOLOGIES ........................................................................................................................................ 211
A. Comparison of public exposures due to radioactive discharges fromthe electricity-generating technologies based on the nuclear fuelcycle, coal cycle, combustion of natural gas and oil, and use ofgeothermal energy .................................................................................................................. 211
B. Comparison of public and occupational exposures .................................................... 220
C. Commentary on significance of accidents ...................................................................... 224

138 UNSCEAR 2016 REPORTIX. RESEARCH NEEDS ............................................................................................................................. 225
X. CONCLUSIONS................................................................................................................................... 227
XI. ACKNOWLEDGEMENTS .................................................................................................................. 229
REFERENCES ............................................................................................................................................... 231

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 139

I. INTRODUCTION
1.
A reliable and affordable supply of electricity is important to improve human health and welfare
worldwide, an objective recognized by both the United Nations Millennium Development Goals [U2,U3] and Sustainable Development Goals. 1 However, policy makers and the general public also haveinterest in the impacts of electricity generation on humankind and the environment.
2.
Interest in exposure of the public and workers to radiation due to nuclear power dates back to theearliest use of the technology. The first UNSCEAR Report to the General Assembly in 1958 included
data on exposures of contract employees of the United States Atomic Energy Commission and arecognition of the need to keep track of the exposure of workers in view of the anticipated growth in theuse of nuclear technology and the associated worker population [U4]. Since then, the Committee has
periodically reviewed exposures of both the public and of workers, related to nuclear power [U5, U6,
U7, U8, U9, U11]. As a result, the Committee’s records of exposures of workers and the public from
nuclear power provide one of the most complete and accurate global pictures of radiation exposuresfrom any source of ionizing radiation.
3.
The Committee has conducted evaluations of radiation exposures of the public arising from formsof electricity generation other than nuclear power to a much more limited extent. Radiation exposuresfrom industrial activities outside of the nuclear sector are generally not monitored or reported in asystematic manner; consequently, the assessment of these exposures has often relied on incomplete datafrom isolated surveys or ad hoc collection of data from various reports and publications.
4.
The Committee’s thematic priorities identified in its strategic plan 2 for 2009–2013 included
radiation levels of energy production and in its plan 3 for 2014–2019 the global impact of energy
production. The Committee decided to update its assessments of the exposures from electricity
generation, considering the principal relevant commercial technologies, both nuclear and non-nuclear.
5.
The world’s mix of electricity-generating technologies changes over time in response to thelandscape of climatic, environmental, resource, political and economic challenges. Governments andresearchers may conduct various comparative studies that among other things take into account thevarious implications for the public and the environment of the different technologies. Exposure toionizing radiation is only one of the many factors that such assessments may take into account.
However, the Committee considers that an update and extension of its past assessments of radiation
exposures of the public and workers from electricity generation could be a useful source of information
for such studies.
6.
This scientific annex thus presents an analysis of the total population exposure, public andoccupational, to ionizing radiation from the different life cycle stages of electricity-generating
technologies, normalized to the electricity generated during one year for that technology. The annex isalso an update and extension of the Committee’s earlier evaluations of radiation exposures resulting
from discharges associated with different electricity-generating technologies. The Committee
1

Transforming our world: the 2030 Agenda for Sustainable Development. Resolution adopted by the General Assembly on
25 September 2015. General Assembly, Seventieth session, Agenda items 15 and 116 (A/RES/70/1).
2
Official Records of the General Assembly, Sixty-third session, Supplement No. 46 (A/63/46).
3
Official Records of the General Assembly, Sixty-eighth session, Supplement No. 46 (A/68/46).

140 UNSCEAR 2016 REPORT

emphasizes that the objective of this study was comparative in nature. Common approaches, available
data and balanced assumptions for assessing the main exposures and identifying the dominantcomponents of those exposures for different electricity-generating technologies have been used to theextent possible.
7.

Specifically, the following core questions have been addressed:− How do the individual and population exposures of humans to ionizing radiation that
result from different phases of the life cycle of each electricity-generating technology
compare with each other across the world and by region?
− What are the main factors, in terms of principal sources, radionuclides, time periods andexposed populations, contributing to the exposure for each technology?
− What would be needed to improve such assessments in the future in terms of data andresearch?

8.
To conduct the assessment, the Committee has in parallel reviewed and revised its methodology
for estimating exposures of the public due to radioactive discharges (annex A). Other exposures, suchas occupational exposures, are assessed using other methods described here.

A. Scope
9.
This annex describes an assessment of the exposures to ionizing radiation in the life cycle ofelectricity-generating technologies that are currently deployed commercially, and fuelled by nuclear
energy, combustion of coal, oil or gas, geothermal energy, solar energy, wind or biomass. The
assessment includes exposures from activities in the life cycle of each technology that are relevant to theradiological impact, from construction to decommissioning. For electricity-generating technologies
using solar energy, wind and biomass, the only activity that has been assessed in this context is themining of metals needed for construction. Exposures from electrical energy storage, transmission anddistribution are not considered. Moreover, the evaluation does not address energy generation other thanelectricity generation, such as process heat generation or other means of distributing energy, such as
district heating.
10. The assessment considers normal operations only. Exposures from incidents and accidents are notconsidered in detail here; the Committee has earlier reviewed radiation exposures in accidents (annex C[U12]), and considered in detail the radiation exposures from the 1986 accident at the Chernobyl nuclear
power plant (annex D [U12]) and the 2011 accident at the Fukushima-Daiichi nuclear power station(annex A [U13]) and continues to monitor developments after these accidents. The annex discusses theirsignificance only in drawing conclusions from this assessment for the General Assembly.
11. The assessment is limited to considering individual and population exposure of humans.
Exposures of non-human biota in the environment are not considered. Assessing impacts other thanradiation exposure from electricity generation, such as social benefits, economic issues, or nonproliferation or security matters, is outside of the Committee’s remit and competence.
12. This work aims to compare exposures to ionizing radiation for different electricity-generating
technologies where individual and collective doses may be used for performing a comparative exposure
assessment. However, calculated doses are recommended only for comparative purposes and not forANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 141

estimations related to health effects. It is important to state that collective doses, as used here, are solely
an instrument to compare radiation exposures for different technologies used for electricity generation.
Collective dose is not intended as a tool for epidemiological risk assessment. Moreover, the aggregationof very low individual doses over extended time periods is inappropriate for use in risk projections and,in particular, the calculation of numbers of cancer deaths from collective doses based on individual
doses that are well within the variation in background exposure should be avoided. Collective doses
estimated in this annex only provide information for decision makers and researchers on radiation
exposures from different electricity generation technologies. Dose estimations for the evaluation of
implications for health should be more specific to each exposure situation.
13. The annex begins with a chapter providing background information to support the study, followed
by a chapter discussing the assessment approach and the end points for the study. That chapter also
includes a description of terminology relevant to applying the methodology for estimating public
exposures due to radioactive discharges (see annex A). The following chapters cover the radiation
exposures arising from electricity generation from each of the electricity-generating technologies: nuclear
fuel cycle, fossil fuel energy (coal, oil and gas) and geothermal energy. Next comes a chapter on assessing
occupational doses from the mining of metals for the construction phase of the electricity-generating
technologies: nuclear, coal, natural gas and the renewable technologies (biomass, solar and wind). Finally,
there are chapters that include comparisons of exposures from the principal electricity-generating
technologies, discussions of uncertainties, suggestions for future work that could improve understanding,and concluding remarks. Note that all supporting values and calculations were manipulated with full
precision and any discrepancies in the numbers presented in tables and figures are due to rounding.

II. BACKGROUND
14. The Committee’s past assessments in this field have covered each stage of the nuclear fuel cycle:
resource extraction (uranium mining and milling); fuel manufacture (uranium enrichment and fuelfabrication); power generation (nuclear power reactors); reprocessing of spent nuclear fuel to recoveruranium and plutonium for subsequent use in nuclear fuels; and the management of solid wastes
generated at the various stages. They have also included exposures from transportation activities within
and between the fuel-cycle stages [U5, U6, U7, U8, U9, U11].
15. The Committee’s most recent and comprehensive review of public exposures due to non-nuclear
electricity generation (coal, oil, natural gas, peat and geothermal technologies) was published in 1993
[U8], based principally on the parameters and assessments presented in the UNSCEAR 1988 Report[U7] and the assessment methodology adopted in the UNSCEAR 1982 Report [U6]. Public exposures
were expressed in terms of the collective effective dose normalized to electricity generated, in units ofman–sieverts per gigawatt–year. The results from those earlier UNSCEAR assessments of normalized
collective effective doses to the public from discharges for various electricity-generating technologies
are summarized in table 1.
16. According to the UNSCEAR 1993 Report, public exposures in terms of collective effective dose per
year of practice and per unit of electricity generated due to the burning of coal and peat, and due togeothermal sources exceeded that due to operational discharges from nuclear power generation. However,with continued improvements in efficiency of electricity generation from fossil-fuel plants and in emission
control technology, those earlier estimates of the Committee were deemed very likely to be outdated andpotentially misleading. Reductions in exposures were expected from newer and retrofitted power plants.

142 UNSCEAR 2016 REPORT

Since its 1993 evaluation, the Committee had only updated its estimates of exposures from the generationof nuclear power, where the most recent estimates showed a decrease for the normalized collective
effective dose to the public from power plant operation from 0.45 man Sv/(GW a) for the period 1990–
1994 to 0.27 man Sv/(GW a) for the period 1998–2002 [U11].
Table 1. Collective effective dose to the public normalized to electricity generated due todischarges from different electricity-generating technologies from previous UNSCEAR assessments

UNSCEAR ReportaNormalized collective effective dose to public(man Sv/(GW a))
Nuclear power
plant operation

Coal

Peat

Gas

Oil

Geothermal

1982 [U6]

4.2a

2

—

—

—

6

1988 [U7]

2.5

c

4

2

0.03

0.5

2

1993 [U8]d

1.34e

20f

—

(0.03)(0.5)(2)

2000 (for 1990-1994) [U9]

0.45g

—

—

—

—

—

2008 (for 1998-2002) [U11]

0.27

—

—

—

—

—

b

h

The value for the nuclear fuel cycle was 5.7 man Sv/(GW a); for globally-circulating radionuclides after 100 years the value

was 12 man Sv/(GW a); for uranium mine and mill tailings after 100 years of radon releases, the value was 0.25 man Sv/(GW a).
b

The value for the nuclear fuel cycle was 4.0 man Sv/(GW a); for mine and mill tailings and globally-circulating radionuclides
the total was 200 man Sv/(GW a).
c

The value was 6 man Sv/(GW a) for older coal-fired power plants, which were considered to constitute two thirds of the world
total and 0.3 man Sv/(GW a) for modern coal-fired power plants, which were considered to constitute one third of the world total.

d

The values reported in the UNSCEAR 1993 Report [U8] for gas, oil and geothermal technologies were not from newassessments but taken from the UNSCEAR 1988 Report [U7].
eThe value for the nuclear fuel cycle was 3.0 man Sv/(GW a); for globally-circulating radionuclides after 10,000 years the value

was 50 man Sv/(GW a); for uranium mine and mill tailings after 10,000 years of radon releases, the value was 150 man Sv/(GW a).
f

From coal-fired power plants (assumed one third modern-style, one third old-style and one third Chinese-style [U8]).

g

The reported dose value is for power plant operation only.

h

The value for the nuclear fuel cycle was 0.72 man Sv/(GW a).

A. Global trends in electricity generation technology
17. Figure I shows the global trends in electricity generation and contributions made by the various
electricity-generating technologies between 1980 and 2014. Combustion of coal for electricity
generation has dominated during this period with nearly 40% of total electricity generated; at the sametime, total global electricity generation has grown by about a factor of three. The use of both natural gas
and nuclear fuels to generate electricity has grown as a percentage of the total and also in total amountof electricity generated over the same period. The use of oil has fallen as a percentage of the total, but
only slightly in total electricity generated. The use of hydropower has increased in the total amount ofelectricity generation, although it has fallen as a percentage of the total. The contribution of the “others”
category, which includes renewable energies (geothermal, wind, solar and biomass), has increased from
1% to 7% during the 34 years, with a faster increase in their use since about 2008.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 143

Figure I. Trends in worldwide electricity generation (GW a) from 1980 to 2014
The “others” category comprises geothermal, solar, wind, biomass and waste [W13]. Supporting values were expressed as full
precision and any discrepancies in the numbers presented are due to rounding

144 UNSCEAR 2016 REPORTB. Electricity generation worldwide in 2010
18. The reference year for assessments presented in this annex is 2010. Table 2 and figure II
summarize worldwide electricity generation in 2010 [I15]. Total electricity generation worldwide in2010 is reported as 2,452 GW a (table 2), which also shows the contribution by type for the six
geographical regions of the world as adopted by the United Nations Environment Programme (UNEP).
These regions are the same as those used in annex A. Marginally different values for electricity
generated by nuclear power in 2010 were obtained from IAEA [I6] and used in the application of therevised methodology for estimating public exposures due to radioactive discharges (annex A). The
IAEA database [I6] provided detailed data on electricity generated in 2010 both by region and as afunction of reactor type; these data were not used to generate table 2, but were used in the more detailed
assessments discussed in chapter IV.B.2.
19. Total annual average worldwide electricity generated from nuclear power plants 1998–2002 wasreported as 278 GW a in table 16 and table A-5 from [U11], which can be compared to 314 GW a for 2010.
Table 2. Summary of worldwide electricity generated in 2010
Supporting values were manipulated with full precision and any discrepancies in the numbers presented are due to rounding

Electricity generated in 2010 (GW a) [I15, I16]

Type
AfricaAsia and
Pacific

EuropeLatin America
and Caribbean

North
America

West
Asiaa

Total

Nuclearb

1.38

66.4

137.4

3.14

106.0

—

314.3

Coal

29.6

567.7

150.6

6.26

237.4

0.00

991.6

Gas

22.7

146.9

177.0

35.8

122.0

39.4

543.8

Oil

9.26

36.9

12.8

20.2

6.33

27.1

112.6

Geothermal

0.17

3.19

1.28

1.13

2.00

—

7.77

Solar

0.06

0.68

2.62

0.03

0.47

0.00

3.86

Wind

0.23

8.80

17.4

0.14

11.9

0.00

38.5

Biomass

0.09

5.95

17.3

4.85

9.59

0.00

37.8

Hydro

12.4

128.6

105.3

80.9

72.7

0.94

400.8

Tide

—

—

0.11

0.02

0.00

—

0.13

Other

—

0.00

0.55

—

0.10

—

0.65

Total

76

965

622

153

569

67

2 452

a

The numerical entry 0.00 corresponds to no electricity generated and the entry — corresponds to no data available. West Asia isomitted in all tables generated from the assessments reported in this annex, except for the assessments on gas and oil.

b

Marginally different values, obtained from IAEA [I6], were used in the application of the revised methodology for estimating
public exposures due to radioactive discharges (annex A) for the electricity generated by nuclear power in 2010.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 145

Figure II. Breakdown of the electricity generated worldwide in 2010 by generation technology
Derived from data in table 2 (excluding contributions from “Tide” and “Other”. Supporting calculations were made to full
precision and any discrepancies in the final numbers are due to rounding

III. ASSESSMENT APPROACH AND END POINTS
20. The main end point in this annex is an assessment and comparison of collective doses, andcollective doses normalized to the electricity produced, for one year of different electricity-generating
technologies, using consistent assessment approaches. To this end, the dominant exposure pathways foreach of the electricity-generating technologies considered in this study have been assessed.
21. The Committee conducted this study by investigating sources of exposure from electricitygenerating technologies based on (a) nuclear power, (b) the combustion of coal, natural gas, oil andbiofuels, and (c) geothermal, wind and solar power. Two electricity-generating technologies (nuclear
power and combustion of coal) were investigated in detail, because a more robust database existed forthese technologies. The Committee evaluated the main sources of radioactive discharges from the life
cycle of these electricity-generating technologies. For nuclear power, these sources included uranium
mining and milling, mill tailings, power plant operation and reprocessing activities. For combustion ofcoal, they were the mining for coal, power plant operation for a prototype of both a modern coal plant
and an older-style coal plant, and deposits of coal ash. These sets of sources will hereafter be called the“nuclear fuel cycle” and “coal cycle”, respectively, for simplicity.

146 UNSCEAR 2016 REPORT

22. The Committee pursued two paths for the assessments presented in this annex. First,
UNSCEAR’s revised methodology for estimating public exposures due to radioactive discharges(annex A) was used to assess components of exposure due to electricity generation from the nuclear
fuel cycle and from the coal cycle in a consistent manner. These two technologies were chosen for
substantive treatment, because reasonably reliable input data existed and, from previous assessments,
they were known to be important. The use of a common methodology for these assessments allowed
consistent comparison of estimated exposures from the two technologies. A significant end point tothese assessments was to identify and discuss dominant sources of exposure from these two electricitygenerating technologies, and the dominant radionuclides contributing to those exposures.
23. The second path was to derive reasonable estimates of public exposures arising from the otherelectricity-generating technologies, and occupational exposures from all the electricity-generating
technologies considered. This permits a rough comparison of the total human exposure from each of theelectricity-generating technologies normalized to the energy generated by each technology.
24. For those electricity-generating technologies relying on power plants that burn natural gas or oil
as fuel, and power plants driven by geothermal energy, an important exposure pathway is radon gas
discharged to atmosphere. The Committee’s revised methodology for estimating public exposures due
to radioactive discharges (annex A) has been used for assessing atmospheric discharges from thesetechnologies.
25. The revised methodology (annex A) was designed for assessing public exposures from routine
releases of radionuclides to the environment during normal operations, and not for assessing occupational
exposures. Thus, occupational exposures due to electricity generation from the nuclear fuel cycle and fromthe coal cycle have been assessed here primarily from data on occupational exposure from the UNSCEAR
2008 Report [U11], adjusted for electricity generated in 2010. In some cases, particularly occupational
exposures from mining for coal and uranium and other metals, relevant new data have been obtained andestimates of dose commitments have been updated for 2010. Occupational exposures from the mining of
metal ores needed for the construction of power plants or power-generating equipment have been assessedand compared for the various electricity-generating technologies.
26. Assessments have also been made of collective effective doses for the entire decommissioning
phase of nuclear power plants based on available data on occupational exposures from existing
decommissioned plants.
27. Exposures of the public and workers due to radioactive materials from conventionalhydroelectricity generation were not included in this annex. This is because no radioactive materials aredischarged during normal operations, and the magnitude of discharges of radioactive materials andresulting exposures when constructing dams were deemed negligible when compared to otherexposures (for example, the mining of metals or mining for other fuels, such as coal or uranium). The
main exposure situation associated with hydroelectricity generation was considered to be occupational
exposure during the mining of metals needed for construction of the power plants. However, thisexposure was not evaluated because hydroelectric plants have considerable variation in their size,
leading to large uncertainties when adopting a standard plant design for assessment.
28. Finally, the total exposures during one year, including both public and occupational, for eachelectricity-generating technology (nuclear fuel cycle, coal cycle, oil, natural gas, geothermal, solar andwind), have been compared. Three main phases that contribute to the total exposure of the public andworkers for each technology have been considered: preparation, operation and decommissioning. For
all discharges, the same methodology has been used to estimate public exposures—a significant
improvement over the Committee’s earlier assessments.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 147

A. Recapitulation of key features of the methodology for estimating
public exposures due to radioactive discharges
29. The following paragraphs recapitulate some of the key features of the methodology described inannex A that are relevant for the assessment here.
30. World-average. This term is used to qualify data that are intended to represent a value averaged
across the whole world. For example, for the default population distributions (shown in table 3 ofannex A), the world-average population for a distance band is derived from the average of the values
for the distance band within each UNEP region, weighted by the population in each region. Similarly,
world-average consumption rates for terrestrial and aquatic foods are derived from the values for each
UNEP region weighted by the populations in each region.
31. Dose. In all cases and unless otherwise stated, dose refers to the protection quantity, effective
dose. The annual effective dose is the sum of the dose from external exposure in that year plus thecommitted dose from intakes by inhalation and ingestion in that year. For doses integrated over a
period, it is the sum of integrated doses from external exposure over the period and committed doses
from the integrated intakes by inhalation and ingestion over the period. Doses to the public are only
estimated for adults and the period considered for the committed dose is 50 years, i.e. adult ages from
20 to 70 years.
32. Characteristic individual dose. The individuals considered are those living in the area local to thepoint of discharge with behaviour indicative of the majority of people living in that area. The dose tothese characteristic individuals is referred to here as the “characteristic individual dose”.
33. For discharges to atmosphere, the characteristic individuals were assumed to live 5 km from thedischarge point and to obtain 25% of their food from this distance. For discharges to rivers, thecharacteristic individual was assumed to be exposed to riverbank sediment 5 km downstream from thedischarge point, to drink all of their water from the river, and to eat 25% of the freshwater fish in theirdiet from the river plus 25% of their dietary grain and leafy vegetables irrigated with water from theriver. For discharges to marine environments, the characteristic individual was assumed to be exposed
to external irradiation from radionuclides in sediments from the local marine compartment using afactor to represent average occupancy of beaches. The characteristic individual is also assumed toconsume marine foods with 25% of their dietary fish from the local marine compartment and 75% fromthe regional compartment, while 100% of the dietary consumption of crustacea and molluscs wereassumed to be from the local compartment. In all cases the consumption rates used were the average fora population. These assumptions are discussed in detail in annex A.
34. Collective doses. The standard collective dose calculated here is the integrated dose to 100 years
from one year’s discharge. (For the modelling of the globally-circulating radionuclides, collective doses
integrated to 500 and 10,000 years are also calculated, see paragraph 39.) The end point for all
radionuclides calculated in annex A was the collective dose per unit discharge integrated to 100 years.
For releases to atmosphere, the local and regional components of the collective doses derive from doses
to populations within a distance of 1,500 km from the discharge point and are based on four different
types of population distributions:(a) A default population distribution based on population densities for 2010. Values of population
density are given for six different regions (as adopted by UNEP) and also a world-average value (apopulation-weighted average of the population density for the UNEP regions). These are used toassess collective doses due to discharges from the coal cycle, combustion of gas and oil, andgeothermal energy.

148 UNSCEAR 2016 REPORT(b) The population distribution around nuclear power stations situated on the coast. Values of theaverage population within distance bands around coastal nuclear power stations are given forfive regions as adopted by UNEP (but not West Asia because there was no operating nuclear power
station in this region at the time of the assessment) and also an average value for the world. The
population densities were based on data for coastal sites of actual nuclear power stations
throughout the world and the distributions are arithmetic means of the data for distance bands
around each coastal site located in the region of interest.
(c) The population distribution around nuclear power stations situated inland. Values of theaverage population within distance bands around inland nuclear power stations are given for
four regions as adopted by UNEP (but not Africa or West Asia because there were no operating
nuclear power stations located inland in these regions at the time of the assessment) and also anaverage value for the world. The population densities were based on data for inland sites of actual
nuclear power stations throughout the world and the distributions are arithmetic means of the datafor distance bands around each inland site located in the region of interest.
(d) A single value of population density for remote sites of 5 km-2. This is used to assess collective
doses due to discharges from uranium mines and mill tailings, which are located in areas of very low
population density. This value is also used in an alternate calculation to assess collective doses due todischarges from geothermal sites. Values for regions of the world are not provided for these cases.
35. For releases to atmosphere, a distinction could be made between the local component of collective
dose (to populations within a distance of 100 km) and a regional component of the collective doses (topopulations between 100 and 1,500 km). For aquatic releases to rivers and lakes, it was not possible todistinguish between local and regional components of the collective dose because of the method used to
calculate collective dose. However, for discharges to marine environments, local and regional
compartments were used to model the dispersion of the released radionuclide, and therefore local andregional components of the collective doses could be provided. Note that the use of the word “regional”
here is distinct from its potential use as a qualifier meaning “related to the geographical regions adopted
by UNEP”.
36. Globally-circulating radionuclides. Four radionuclides (3H, 14C, 85Kr and 129I) were assumed to
become globally circulated and to continue to expose the world population for decades and, withregards to 14C, centuries, and 129I, millions of years. These radionuclides contribute to the local andregional component of collective dose when they are initially discharged (due to the so-called “firstpass”) and this component was modelled as for all other radionuclides. However, a global component
needs to be added separately for these radionuclides. Thus, an additional collective dose due to global
circulation was also modelled and results calculated. Results of the global model are provided fordischarges to atmosphere and to marine environments; for aquatic discharges to freshwater systems, theresults were assumed identical to those for a marine environment for the purposes of calculating the
global component of collective dose.
37. Integration times. The methodology applies to discharges that can be assumed to be continuous.
Account is taken of the build-up of long-lived radionuclides in the environment and the associatedcontinued exposure after the discharges have stopped. This is done by considering a year’s discharge ofa radionuclide, its dispersion in the environment and the subsequent exposure of people over many
years; the resulting dose rates are then integrated. A value of 100 years is used for the integration period
for the standard case (see further details in annex A).
38. Using the integrals it is also possible to consider that the discharges continue for many years fromthe same site. This is because the integrated dose to 100 years from one year’s discharge is numerically

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 149

equal to the dose in the 100th year from a continuous discharge at a constant rate over a 100-year
period. The characteristic individual doses are calculated by integrating dose rates to 100 years, which
also represent the annual dose that would be received in the 100th year of discharge.
39. Collective doses can be integrated to various times, but for most radionuclides it was sufficient tointegrate to 100 years because most of the collective dose commitment (i.e. the collective dosetheoretically integrated to infinity) is delivered during this period. However, for globally-circulating
radionuclides (i.e. 3H, 14C, 85Kr and 129I), the integrated collective dose continues to increase for many
years beyond 100 years (see annex A, paragraph 14). Therefore, for globally-circulating radionuclides,
collective doses integrated to 500 and 10,000 years were also calculated.
40. The main results of the revised methodology (annex A) are for a unit discharge for one year
integrated to 100 years and can be used to estimate the impact of one-year’s practice by, for example,
scaling the results by the discharge of radionuclides per unit of electricity generated, or the total
discharges from a particular type of electricity generation.
41. Waste residues from uranium mining and from coal-fired power stations. These wastes, uranium
mill tailings and coal ash, are often disposed of on the surface of the ground and they contain enhanced
levels of naturally occurring radionuclides, notably 226Ra which decays into the gas 222Rn, which can beemitted into the air for many years after the disposal occurs. The radon and its decay products give riseto human exposure (see figure III).
42. In this case one-year’s practice gives rise to a continuing discharge to atmosphere for many years.
Current best practice would rehabilitate mill tailings such that emissions were reduced to backgroundlevels (i.e. the levels that would occur from that area if the mine had not been present). Informationfrom Australia and elsewhere indicates that this process might take some 50 years and that monitoringis expected to continue for some 20 to 30 years to ensure that the rehabilitation has been successful andto monitor for any deterioration [S2, W1]. The question remains whether the site could deteriorate once
institutional control has finished, and radon emissions increase again. A base case integration time of
100 years seems reasonable with 10,000 years a cautious upper estimate.
43. Therefore, for radon discharges from land disposal of mill tailings or coal ash, allowance has to bemade for this continued discharge and this was done by multiplying the results for one year’s discharge
by 100 to allow for the discharge continuing for 100 years (see annex A, paragraph 17). This 100-year
multiplication is chosen for both individual and collective doses because 100 years is not muchdifferent from a human lifetime. The Committee has also considered the effect on the integrated
collective dose of a continuing discharge for 500 and 10,000 years.

150 UNSCEAR 2016 REPORT

Figure III. The 238U and 232Th decay series
Half-life is expressed in a = year; d = day; h = hours; m = minutes; s = seconds

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 151

44. Representative radionuclides for nuclear power plants. Discharges of radionuclides from nuclear
power plants are commonly reported by operators and regulatory bodies aggregated in groups. For
example in UNSCEAR 2008 discharges were reported as “noble gases” and “particulates” foratmospheric discharges, or “other nuclides” for liquid discharges. Accurate assessments of doses canonly be carried out on a radionuclide-specific basis and so where groups of radionuclides had beenreported, it was necessary either to apportion the discharge for the groups of nuclides among theradionuclides, or to use a “representative” radionuclide. The European Commission [E1] gives thepercentage of radionuclides discharged for different reactor types sited in Europe for different
groupings. For AGRs, FBRs and GCRs, 4 the discharges were split between the radionuclides
considered in the workbooks for the appropriate groupings (atmospheric noble gases, atmosphericparticulates and other liquids). The breakdown for GCRs was based on United Kingdom sites because
these were the only ones operational in 2010. For atmospheric particulate discharges from BWRs and
PWRs, 54Mn was selected as the representative radionuclide for any radionuclides not considered
explicitly in the workbooks. The grouping of “other liquids” did not previously include 14C and, formany countries and most reactor types, liquid discharges of 14C are not reported. However, examinationof EC RADD [E2] shows that for European PWRs liquid discharges of 14C are significant and therefore
30% of discharges from “other liquids” were assumed to be 14C for PWRs. For BWRs and PWRs any
liquid discharges for radionuclides not included in the workbooks were allocated equally between 54Mn
and 58Co. For HWRs, 41Ar, 60Co and 137Cs were assumed to represent discharges for noble gases,
atmospheric particulates and other liquids respectively, based on information in [B3, B5, C2]. For
LWGRs information was taken from the EC RADD database [E2].

IV. RADIATION EXPOSURES ARISING FROM ELECTRICITY
GENERATION BY NUCLEAR POWER
A. Introduction
45. The Committee has studied radiation exposures arising from electricity generation by nuclear
power repeatedly over the years since its first publication on this subject in 1958 (see especially [U6,U7, U8, U9, U11]). These studies have consistently shown that the major contribution to public
exposures has been through the discharges of natural radionuclides, primarily from: radon and its
progeny released during uranium mining and milling, and from mill tailings, associated with the nuclear
fuel cycle; and from carbon-14, primarily associated with reactor operation and fuel reprocessing.
46. The revised methodology for estimating public exposures due to radioactive discharges (annex A)
has been used here to assess public exposures from the following processes: uranium mining andmilling; electricity generation from nuclear power reactors; and fuel reprocessing. The assessments andresults for these processes are presented in the following section. The processes—uranium enrichment,
fuel fabrication and solid waste disposal—and their exposure characteristics are discussed in sections
4

The following abbreviations are used for the different nuclear power reactor types, categorized according to their coolant
systems and moderators: light-water-moderated and cooled pressurized or boiling-water reactors (PWRs and BWRs); heavywater-cooled and moderated reactors (HWRs); gas-cooled, graphite-moderated reactors (GCRs and AGRs); light-water cooled,graphite-moderated reactors (LWGRs); and the liquid metal cooled fast-breeder reactors (FBRs).

152 UNSCEAR 2016 REPORT

IV.B.4 and IV.B.5, and dose estimates for these processes that were published in the UNSCEAR 2008
Report [U11] have been used here. This chapter begins with the assessment of public exposures,
followed by that of occupational exposures, and ends with considering the occupational exposuresassociated specifically with the decommissioning process.

B. Public exposure
1.

Mining and milling

47. Uranium ore can be extracted from the earth by physically removing it through conventionalsurface or underground mining methods or by chemically dissolving the uranium out of the rock ore
through either heap leaching or in-situ leaching (ISL) (sometimes referred to as in situ recovery or ISR)
[U18]. Surface mining (also referred to as opencast, open pit or strip mining) techniques are applied toore bodies that are close to the surface, and are also a generally cost-effective method for extracting
large volumes of lower-grade ore, that may then be combined with other bulk extraction techniques(such as conventional milling, leaching and extraction, or alternative techniques such as heap leaching)which would be uneconomical for underground operations. Underground mining involves extractingrock through a tunnel or opening in the side of a hill or mountain and is generally applied for the
extraction of higher-grade ores. In-situ leaching is generally applied to shallow deposits that exist in
non-porous shale or mudstone, or in situations where uranium can be recovered from otherwise
inaccessible or uneconomical formations [U18].
48. Uranium ores typically contain from about 0.05 to 0.3% uranium oxide (U3O8) [U18]. After
extraction from the ground, the mined uranium ore is sent to a mill, which is usually located close to themine. This next step in the nuclear fuel cycle, called “milling”, involves the extraction and purificationof uranium from the uranium ore. After a first purification process, the uranium is precipitated in apartially refined form, known as “yellowcake”. The uranium concentrate, typically containing 75 to
95% uranium, is shipped to a chemical plant for further purification and chemical conversion [B4].
49. In-situ leaching is an alternative method for extracting uranium from low-grade ores or shallow
deposits that exist in non-porous shale or mudstone. An ISL plant chemically alters the uranium ore
underground before it is pumped out for processing. In the ISL process, wells are drilled into rockcontaining uranium ore. An alkaline or acidic solution (known as lixiviant) is injected down the wells todissolve the uranium in the rock. In the case of alkaline solutions, the lixiviant is usually (a) water
mixed with oxygen and/or (b) hydrogen peroxide mixed with sodium carbonate or carbon dioxide. In
the acid leach processes (used in Kazakhstan and Australia), sulphuric acid usage can result in lixiviant
solutions with pH as low as 1 being circulated into the ore body (the process in Kazakhstan is typicallystrongly acidic, while in Australia only mildly acidic). In some cases, an oxidant is added to increase
the efficiency of leaching. The lixiviant is then collected in a series of recovery wells, through which itis pumped to a processing plant. At the processing plant, the uranium is extracted from the solution
through an ion-exchange process, and the uranium oxide concentrate (yellowcake) is then precipitated,
dried and packed.
50. After recovery of the uranium, the barren solution is re-fortified with oxidant (if required) before
being returned to the well field via the injection wells. However, a small flow (about 0.5%) is bled off tomaintain a pressure gradient in the well field and this, with some solutions from surface processing, istreated as waste. This waste water contains various dissolved ions such as chloride, sulphate, sodium,radium, arsenic and iron from the ore body and is re-injected into approved disposal wells in a depleted

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 153

portion of the ore body. Wells must be monitored to ensure that extraction fluids do not leave the facility
or contaminate groundwater. Waste from this process, usually filters and piping, can be disposed in a
tailings pile at a mill site or a licensed disposal facility. In-situ leaching facilities have no radon dischargeduring the “mining” phase, and no surface tailings and little radon emission after closure [U11]. There canhowever be discharge of radon during the “leaching” phase of the ISL mining process, which here has
been assumed to be the same as the radon discharge occurring during the milling phase [B6, B7].
Chapter IX discusses the need for further study regarding the ISL mining process.
51. In 2014, 51% of the world’s uranium was mined using ISL operations, a share that has risen
steadily mainly because of mining operations in Kazakhstan [W8]. Table 3 shows the percentage
contribution from ISL-mining to total uranium production during the period from 2008 to 2012 in eachof the UNEP regions. For the reference year of 2010 used in this annex, the total uranium production
for 15 countries is given in table 4.
Table 3. Contribution of ISL-mining to total uranium production between 2008 and 2012
Information on percentage ISL compared with other types of mining is taken from [W10]RegionContribution to world production (%)
Total

ISL

OthersAsia and Pacific

51.6%

38%

13.6%

North America

21.1%

2%

19.1%

Africa

18.4%

0%

18.4%

Europe

8.6%

1%

7.6%

Latin America

0.3%

0%

0.3%

World

100%

41%

59%

Table 4. Uranium production in 2010 [O8]The production numbers come from OECD/NEA (table 1.21, page 60, Historical uranium production [O8]). Countries that hada label “Secretariat estimate” are not included here (Pakistan, Romania and India). Also excluded are a few countries that hadvery small production which came from mine rehabilitation efforts only

Country (UNEP region)

Uranium production (t U)

Kazakhstan (Asia and Pacific)

17 803

Canada (North America)

9 775

Australia (Asia and Pacific)

5 900

Namibia (Africa)

4 503

Niger (Africa)

4 197

Russian Federation (Europe)

3 563

Uzbekistan (Asia and Pacific)

2 874

United States (North America)

1 630

China (Asia and Pacific)

1 350

154 UNSCEAR 2016 REPORTCountry (UNEP region)

Uranium production (t U)

Ukraine (Europe)

837

Malawi (Africa)

681

South Africa (Africa)

582

Czech Republic (Europe)

254

Brazil (Latin America)

148

Iran, Islamic Rep. of (Asia and Pacific)

7

Total

54 104(a) Input to assessments for mining and milling
52. The Committee’s revised methodology for estimating public exposures due to radioactive
discharges (annex A) has been used to assess the collective doses and characteristic individual doses
from four categories of sources for radon discharges:(a) Mining (underground uranium mining)(b) Milling (natural radionuclides other than radon are also discharged)(c) Operational mill tailings(d) Mill tailings
53. The third category, operational mill tailings, refers to the mill tailings produced during the milling
process, and the fourth category, mill tailings, represents the tailings that are positioned in some otherplace for more permanent holding. They can be eventually treated or mitigated to reduce the dischargeof radionuclides. Assessments were conducted using the methodology described in annex A for the four
categories of sources above. They were performed using the discharges normalized to energy generated
shown in table 5 and table 6, and the values for electricity generation for each region given in table 7 toestimate characteristic individual doses to the public. As noted in the tables, many of the values for thenormalized discharges were taken from previous UNSCEAR reports. However, the estimate for thedischarge of radon from mill tailings has been updated. A value for the discharge of radon per energy
produced of 0.3 TBq/(GW a), based on a tailings area of 1 hectare and a radon emanation rate of
1 Bq s−1 m−2, was used in [U9]. This was lower than the previous value of 1 TBq/(GW a). Information
on emissions from Australian mine tailings [L2] is that the value of 1 Bq s−1 m−2 is high compared to
the range of levels that are measured. Measurements taken over a number of years at the El Sharena
mines were in the range 9–36 mBq s−1 m−2 with an average baseline value of 17 mBq which isequivalent to 0.005 TBq/(GW a) based on a tailings area per electricity generated of 1 ha/(GW a). In
other cases where the tailings have been rehabilitated, the emissions rates are at background levels. A
study at the Ranger uranium mine found that the emission rates ranged from 0.2 to about 0.9 Bq s−1 m−2.
The studies related to trial landforms intended to develop their rehabilitation strategies with the aim of
reducing the long-term radon emission rates. For Olympic Dam, an emission rate of 0.5 Bq s−1 m−2 waspublished, a value equivalent to 0.15 TBq/(GW a) while other information gives emission rates of 0.3,
0.14 and 0.04 Bq s−1 m−2 for the inner mine, outer mine and region. These are equivalent to 0.1, 0.04
and 0.01 TBq/(GW a) based on 1 ha/(GW a). These Australian values indicate that the UNSCEAR
2000 value (0.3 TBq/(GW a)) may be too high based on mines operating with best practice for which avalue of 0.1 or lower may be more appropriate. In the light of this, the Committee agreed that a roundedemission rate of 0.1 TBq/(GW a) should be used in this study.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 155

Table 5. Discharges of radon from the uranium mining and milling process normalized toelectricity generated
SourceNormalized discharges(TBq/(GW a))

Mining (world average, all mines except ISL, only radon discharges)

66aMining (world average, all ISL mines, only radon discharges from leaching phase of
ISL mining process)3

Milling (world average, all mines, radon discharges plus nuclides in table 6)

3bOperational mill tailings (world average, all mines except ISL, only radon discharges)

3cMill tailings (world average, all mines except ISL, only radon discharges)

0.1d

a

A value of 75 TBq/(GW a) was used in earlier UNSCEAR assessments [U8], and was based on the assumption that there wereon average 300 GBq of radon released per tonne of uranium oxide [U8].

b

From UNSCEAR 1993 Report [U8].

c

From UNSCEAR 2000 Report [U9].

dFrom annex A.

Table 6. Airborne discharges during the milling process (all mines including ISL) ([U8] tables 19 and 22)
Radionuclide

Activity discharged per unit of electricity generated(TBq/(GW a))

210

Po

2 × 10−5

210

Pb

2 × 10−5Ra

2 × 10−5

234

U

4 × 10−4

238

U

4 × 10−4

Th

2 × 10−5

226

230

54. The estimate for the discharge of radon from uranium mining (non-ISL mines) has also been
updated. Total uranium production in 2010, shown in table 4 [O8], was assumed to be the amount thatwas mined in that year. Values were summed for all countries to give a rounded total of 54,100 tonnesof uranium. Similarly, total uranium production in 2010 from table 4 is assumed to equal the total
amount of uranium milled. The value for the mass of uranium required per unit of electricity generated
given in the UNSCEAR 2008 Report was 220 t/(GW a) ([U11] table 18A), which is lower than thevalue of 250 t/(GW a) used earlier by the Committee [U7, U8, U9]. The trend in uranium required perunit of electricity generated (the uranium requirement) is generally downwards, because of increasing
efficiencies in power plant operation and lower enrichment tails assays. For example, the 2014 Red
Book [O8] reported that the generic reactor uranium consumption had reduced from 175 t/GW(e) per
year at 0.30% tails assay [O6] to 163 t/GW(e) per year at 0.25% tails assay. The corresponding figures
for U3O8 are 206 and 192 tonnes, respectively.
55. The value of 220 t/(GW a) used in the UNSCEAR 2008 Report is the value used in this assessment.
Given the information in the previous paragraph, this is possibly a slight overestimate; however, it was
retained for continuity with the previous UNSCEAR assessments. This gives a value for the discharge ofradon from the mining process normalized to electricity generated of 66 TBq/(GW a), which is lower thanthe value of 75 TBq/(GW a) used in earlier UNSCEAR studies [U7, U8, U9].

156 UNSCEAR 2016 REPORT

56. In modern ISL facilities, the circulating lixiviant goes directly from the well field to the header
houses and the dissolved uranium is extracted from the lixiviant through the ion exchange process. The
lixiviant is then reconstituted and returned directly to the well field in an essentially closed(pressurized) system. In older non-pressurized systems (such as at Crow Butte, United States), the resin
column is not pressurized, so the lixiviant is exposed to normal atmospheric conditions, which results insuper-saturated radon being discharged. The discharged radon gas is vented through an exhaust stack to
the atmosphere. This discharge is most similar to that from the milling phase, not from the mining
phase. The assessments in this annex have assumed radon discharges during ISL uranium mining are
the same as those from the milling phase.
57. In order to perform assessments of the characteristic individual doses to the public for eachregion, electricity generation in 2010—broken down into ISL and non-ISL mining sources—wasneeded for the different regions. Total uranium production for 15 countries in 2010 is given in table 4.
Table 3 shows the contribution from ISL-mining to the total uranium production in each UNEP region
between 2008 and 2012. Using the value for uranium requirement of 220 t/(GW a), the electricity
generated from the total uranium production for each region could be estimated as shown in table 7.
Assessments for mining and milling using the revised methodology for estimating public exposures due
to radioactive discharges (annex A) used the values for total electricity generated with the normalizeddischarges for milling, the total electricity generation from ISL mining with the normalized discharges
for ISL mining processes, and the total electricity generation excluding ISL mining sources with thenormalized discharges for the processes mining (non-ISL), operational mill tailings and mill tailings, tocalculate characteristic individual doses. Collective doses resulting from mining and milling wereassessed assuming a low population density (defined in annex A), and no aquatic discharges were
included in the assessment.

Table 7. Inferred electricity generation from the uranium production in each region and from ISL and non-ISL contributions
The assessment assumes 220 tonnes of uranium per gigawatt-year. Note that the total inferred electricity generation is 246 GW a, which is somewhat less than the world total nuclear electricity generation in 2010at 314 GW a (table 2). However, this difference is not unreasonable because current reactor fuel requirements were met from primary supply (direct mine output: 78% in 2009) and secondary sources: commercial
stockpiles, nuclear weapons stockpiles, recycled plutonium and uranium from reprocessing used fuel, and some from re-enrichment of depleted uranium tailings (left over from original enrichment) [O6, W11]
Region% from ISL% from otherproduction methods

Mass of uranium
from ISL (tonnes)

Mass of uranium fromother production
methods (tonnes)

Inferred electricity
generated from ISL(GW a)

Inferred electricity
generated from otherproduction methods(GW a)

Total inferred
electricity generated(GW a)

Africa

18.4

0

18.4

0

9 954

0

45

45

Asia and Pacific

51.6

38

13.6

20 558

7 358

93

33

126

Europe

8.6

1

7.6

541

4 112

3

19

22

Latin America

0.3

0

0.3

0

162

0

1

1

North America

21.1

2

19.1

1 082

10 333

5

47

52

Total

100

41

59.0

22 18131 919

101

145

246

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 157% of world
production

158 UNSCEAR 2016 REPORT

2.

Electricity generation from nuclear power reactors

58. The nuclear power reactors used for electricity generation that are treated here can be categorizedaccording to their coolant systems and moderators: light-water-moderated and cooled pressurized orboiling-water reactors (PWRs and BWRs); heavy-water-cooled and moderated reactors (HWRs); gascooled, graphite-moderated reactors (GCRs and AGRs); and light-water cooled, graphite-moderated
reactors (LWGRs). These reactor types are all thermal reactors that use the moderator material to slow
down the fast fission neutrons to thermal energies. In fast-breeder reactors (FBRs), the coolant is aliquid metal, there is no moderator and fast neutrons induce fission.

(a) Assessments for nuclear reactor operation
59. The distribution of electricity generated by geographic region and type of location of the nuclear
power plants in the world are presented in table 8. The electricity generated in 2010 by nuclear power
was about 300 GW a according to the IAEA PRIS database [I6] (table 9). PWRs contributed most to
the total electricity generated worldwide from nuclear reactors (about 68%), followed by BWRs (about21%) [I6]. The distribution of electricity generated by geographic region from nuclear power plants in2010 by reactor type is shown in table 9. (The ordering of the individual reactor types in all relevant
tables and figures in this annex is according to their world share of electricity generated in 2010, as
shown in table 9.)60. Radionuclide discharges and electricity generation data used with the revised methodology for
estimating public exposures due to radioactive discharges (annex A), for assessing discharges from
nuclear power plants in 2010 normalized to the electricity generated, are described in table 10. The
main basis for deriving the normalized discharge values was the IAEA PRIS database [I6], the EC
RADD database [E2] and the discharge data for the year 2002 in the UNSCEAR 2008 Report [U11].
Using these data the normalized discharges for 2002 were recalculated and adjusted to electricity
generation in 2010, for use with the revised methodology for estimating public exposures due toradioactive discharges (annex A) to assess the exposures for the reference year 2010.
61. A summary of the activities of radionuclides, or radionuclide groups, discharged in airborne andliquid effluents from nuclear reactors during routine operation in 2010, normalized to the electricity
generated in 2010, are reported in table 11. The method for determining how the aggregated nuclides
were characterized, i.e. the noble gases and particulates in the atmospheric discharges, and the “otherliquids” listed in the liquid discharges, has been addressed in earlier studies by UNSCEAR (e.g. [U7,U8, U9]). The aggregated nuclides were called “representative radionuclides for nuclear power plantdischarges” in those studies, and the term “representative radionuclides” is used in this annex also.
Chapter III.A, paragraph 44 in this annex explains how the representative radionuclides, were chosen
for use in the current study with the revised methodology for estimating public exposures due toradioactive discharges (annex A) applied to nuclear power plants.
62. The revised methodology for estimating public exposures due to radioactive discharges (annex A)
was used to assess doses due to discharges from nuclear power plants. The assessments considered
atmospheric and aquatic discharges. Collective dose estimates for discharges to atmosphere were basedon the population distributions around coastal or inland nuclear power plants, using the population
distributions generated for the different regions adopted by UNEP (see annex A). Where appropriate,
contributions from globally-circulating radionuclides were included in the assessment of collective

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 159

doses. The characteristic individual doses and collective doses were calculated separately for each of
the reactor types (PWR, BWR, HWR, LWGR, AGR, GCR and FBR).
Table 8. Number of reactors by type, location and UNEP region (2010 data)
Data from the IAEA PRIS database [I6]Reactor type

AfricaAsia and
Pacific

EuropeLatin America
and Caribbean

North
America

Grand
total

PWR total

2

55

141

2

69

269

Coastal

2

54

34

2

16

108

Inland

0

1

107

0

53

161

BWR total

0

36

19

2

35

92

Coastal

0

36

9

2

2

49

Inland

0

0

10

0

33

43

HWR total

0

24

2

2

18

46

Coastal

0

11

0

0

1

12

Inland

0

13

2

2

17

34

LWGR total

0

0

15

0

0

15

Coastal

0

0

4

0

0

4

Inland

0

0

11

0

0

11

AGR (all coastal)

0

0

14

0

0

14

GCR total

0

0

4

0

0

4

Coastal

0

0

2

0

0

2

Inland

0

0

2

0

0

2

FBR (all inland)

0

0

2

0

0

2

Grand total

2

115

197

6

122

442

160 UNSCEAR 2016 REPORT

Table 9. Electricity generation from nuclear power plants in 2010 by reactor type and UNEP region(from [I6])
The values for the total electricity generation from nuclear power plants for each UNEP region [I9] are marginally different
from those shown in table 2 [I15]. The IAEA database [I6] provided detailed data on electricity generation in 2010 both byregion and as a function of reactor type, which was not available from the database used to generate table 2, and the moredetailed data were therefore chosen for use in the assessments discussed in this chapter

Reactortype

Electricity generation (GW a)
AfricaAsia and
Pacific

EuropeLatin America
and Caribbean

North
America

West
Asia

Total

% totalperreactor
type

PWR

1.47

38.47

102.47

1.57

60.89

0.00

204.88

68

BWR

0.00

19.66

12.18

0.64

31.25

0.00

63.72

21

HWR

0.00

5.34

1.22

0.76

9.76

0.00

17.09

6

LWGR

0.00

0.00

8.15

0.00

0.00

0.00

8.15

3

AGR

0.00

0.00

5.01

0.00

0.00

0.00

5.01

2

GCR

0.00

0.00

0.93

0.00

0.00

0.00

0.93

0

FBR

0.00

0.00

0.42

0.00

0.00

0.00

0.42

0

Total

1.47

63.47

130.392.97

101.890.00

300.21100

Table 10. Key data sources and assumptions used to assess public exposures due to radioactive
discharges from nuclear power plants
Data sources

Commentary

Discharges, 2002

Taken from tables A6-A12 in UNSCEAR 2008 Report [U11]. Note that WWER was
regrouped under PWR. Data for discharges from FBR are from 2001. Discharges ofparticulates to air and other liquids for United Kingdom AGRs/GCRs were takenfrom the EC RADD database [E2]Electricity generation,
2002

Taken from [I6]. Note that HWLWR was regrouped under HWR and PHWR under
HWR. Generation data for FBR are from 2001

Electricity generation,
2010

Taken from [I6]. Note that HWLWR was regrouped under HWR and PHWR under
HWR. Includes a calculation of generation from reprocessing countries as a fractionof all generation

Countries

Countries and regions as adopted by UNEP (see annex A)

Location dataInland/coastal location of reactors, taken from [I6]

Normalized discharge data

For each reactor type:a,b the normalized discharges were scaled to the electricity
generation ratio using the 2002 to 2010 electricity generation data; information
summarized in paragraph 44 was used to apportion discharge between specificradionuclides; and location data were used to apportion between inland/coastal
locationsa

For FBRs, values of normalized discharges of 3H, 131I, 14C, liquid tritium and other liquids were taken from table 17, annex B,
UNSCEAR 2008 Report [U11] because no discharge data were available. Data for the radionuclide mix were taken from [E1].

bFor LWGRs, normalized discharge data were taken from table 17, annex B, UNSCEAR 2008 Report [U11]. EC RADD [E2]data for European LWGR reactors were used to infer mix of noble gases, particulates and other liquids.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 161

Table 11. Estimated normalized discharges from nuclear power plants 2010
Estimated normalized discharges per unit of electricity generated (TBq per GW a)
Reactortype

Discharges to atmosphere
Noble
gases

Tritium131

PWR

5.8 × 100

1.5 × 100BWR

1.8 × 101

HWR

CParticulates

S

LiquidtritiumOtherliquid

8.0 × 10−5

8.3 × 10−2

3.6 × 10−50

1.8 × 101

3.8 × 10−3

1.3 × 100

4.2 × 10−4

1.3 × 10−1

1.8 × 10−30

8.2 × 10−1

2.1 × 10−3

3.5 × 101

2.0 × 102

2.3 × 10−5

6.0 × 10−1

1.7 × 10−50

1.7 × 102

3.1 × 10−2

LWGR

4.6 × 102

2.6 × 101

9.9 × 10−3

1.3 × 100

2.7 × 10−30

7.8 × 10−1

2.0 × 10−3AGR

1.9 × 101

4.0 × 100

3.2 × 10−5

1.4 × 100

2.2 × 10−5

6.6 × 10−2

4.1 × 102

8.1 × 10−1

GCR

1.7 × 103

5.0 × 1000

5.5 × 100

3.0 × 10−4

3.7 × 10−1

4.7 × 100

1.2 × 100

FBR

4.4 × 101

4.9 × 101

2.0 × 10−4

1.2 × 10−1

1.4 × 10−40

1.7 × 100

2.3 × 10−2

3.

I

Aquatic discharges
35

14

Fuel reprocessing

63. Reprocessing of used nuclear fuel has been practised for several decades in a number of countries,
mainly for extracting and recycling fissile materials. Current recycling practices are primarily focusedon the conversion of fertile 5 238U to fissile plutonium, for which a significant amount of the plutonium
recovered from used fuel has been recycled into mixed oxide (MOX) fuel. Table 12 shows the current
location of reprocessing facilities in the world, with operations in France, the Russian Federation andthe United Kingdom dominating. Not shown in the table is the Rokkasho facility in Japan, which is
expected to start operations in 2018 and to have a commercial reprocessing capacity of 800 tonnes per
year. As of 2015, about 31% (90,000 tonnes of 290,000 tonnes) of used fuel from commercial powerreactors has been reprocessed [W12].
64. Characteristic individual doses and collective doses were calculated for Sellafield in the United
Kingdom, La Hague in France and Ozersk (Mayak) in the Russian Federation, which are also the threemain reprocessing facilities (table 12), using available data on discharges. The doses due to discharges
from fuel reprocessing facilities were assessed using the revised methodology for estimating public
exposures due to radioactive discharges (annex A).
65. Data for both atmospheric and aquatic discharges were obtained for the reprocessing facilities at
Sellafield, United Kingdom and La Hague in France for the year 2010 [E2] and for atmosphericdischarges only at Ozersk (Mayak), Russian Federation [F2]. These data include discharges from otheractivities on site but it could be assumed that most discharges are related to reprocessing activities.
Discharges normalized to electricity generation were calculated for La Hague based on the assumptionthat discharges from the reprocessing facility were related to the reprocessing of fuel equivalent to that
required to power French PWRs in 2010, 47 GW a [I6, W9]. However, for Sellafield and Mayak it was
not possible to relate the discharges to electricity generated.
5

Fertile material is a material that is not fissionable by thermal neutrons, but can be converted into a fissile material by neutron
absorption and subsequent nuclei conversions.

162 UNSCEAR 2016 REPORT

66. The discharges for reprocessing facilities are shown in table 13. Consequently, characteristic
individual doses and collective doses (as defined in section III) could be assessed for all three of thesefacilities, all located in the region named Europe (as adopted by UNEP). However, because discharges
normalized to electricity generation were only available for La Hague in France, the characteristic dosenormalized to electricity generated could be assessed only for this facility. Collective doses normalized
to electricity generation for Europe and the whole world were also based on La Hague, but took into
account the fraction of electricity generated by all French PWRs in 2010 compared to Europe and thewhole world, 0.55 and 0.24 respectively, for the calculation of the doses due to the first pass.
Table 12. World commercial reprocessing capacity [N1, O2, W12]
Fuel type
LWR fuel

Facility

Commercial reprocessing capacity(tonnes per year)

France, La Hague

1 700

United Kingdom, Sellafield (THORP)

600

Russian Federation, Ozersk (Mayak)

400

Total LWR
Other nuclear fuelsapproximately 2 700

United Kingdom, Sellafield (Magnox)

1 500

India (PHWR, 4 plants)

330

Japan, Tokai (MOX)

40

Total otherapproximately 1 870

Total civil capacity

approximately 4 570

Table 13. Discharges from reprocessing facilities [E2]Discharged
radionuclide

Discharges to atmosphere
Sellafield

La HagueAquatic discharges
Mayak

Sellafield

La Hague

Mayak

DISCHARGES PER UNIT OF ELECTRICITY GENERATED (TBq/(GW a))H—

3.8 × 104——

6.7 × 106

—

14

C

—

1.1 × 104——

5.0 × 103

—

41

Ar

—

—

—

—

—

—

54

Mn

—

—

—

—

58

Co

—

—

—

60

Co

—

4.9 × 10−3

85

Kr

—

90

3

0

1.4 × 10——−2

7.0 × 10———

4.4 × 101—

1.5 × 108

—

—

—

—

Sr

—

—

106

—

—

9.1 × 10——−2

4.8 × 10

—

—

3

Ru

1.4 × 10

—

129

I

—

3.1 × 100——

9.3 × 102

—

131

I

—

1.6 × 10−1——

7.9 × 100

—

135

Xe

—

—

—

—

—

—

1

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 163

Discharged
radionuclide

Discharges to atmosphere

Aquatic discharges

Sellafield

La Hague

Mayak

Sellafield

La Hague

Mayak

134

Cs

—

3.6 × 10−3——

5.1 × 101

—

137

Cs

—

3.0 × 10——2

7.3 × 10

—

239

Pu

—

1.3 × 10——0

6.3 × 10

—

241

Am

—

—

—

—

1.3 × 101

—

−3
−3

AVERAGE DISCHARGE RATE OVER A YEAR (Bq/s)
H

6

3.1 × 10

1.8 × 106—

4.4 × 107

3.2 × 108

—

14

C

8.7 × 103

5.1 × 105—

1.4 × 105

2.3 × 105

—

41

Ar

—

—

1.2 × 106

—

—

—

54

Mn

—

—

—

58

Co

—

—

60

Co

—

85

Kr

90

3

—

1

6.4 × 10

—

—

—

0

3.3 × 10—

2.3 × 10−1

3.4 × 10−1

3.1 × 103

2.0 × 103—

1.4 × 109

7.1 × 109

—

—

—

—

Sr

1.3 × 10

—

106

Ru

1

2.4 × 10

2.2 × 10

5.7 × 10

3.7 × 10

6.5 × 10

—

129

I

3.1 × 102

1.4 × 102—

8.7 × 103

4.3 × 104

—

131

I

1.2 × 101

7.7 × 100

2.5 × 101—

3.7 × 102

—

135

Xe

—

—

1.8 × 10

—

—

—

134

Cs

—

1.7 × 10137

Cs

3.0 × 100

239

Pu

241

Am

0

0

1

4.2 × 104

3.2 × 103

4.2 × 10

—

4

4

0

5

—

3

3.5 × 103

2.4 × 10—

1.4 × 10−1

2.1 × 101

1.5 × 105

3.4 × 104—

6.2 × 10−1

5.9 × 10−2

1.3 × 101

4.2 × 103

2.9 × 102—

1.2 × 10——

1.1 × 10

6.2 × 10

—

−1

0

4

2

(a) Results—public exposures from mining and milling, electricity generation from
nuclear power reactors and fuel reprocessing
67. Characteristic individual doses to the public—for the nuclear fuel cycle processes: mining andmilling; electricity generation from nuclear power reactors; and fuel reprocessing—normalized to theelectricity each process generated in 2010 are summarized in table 14. Doses for the characteristic
individual from mining and milling are shown for both non-ISL and ISL mines. The characteristic
individual was assumed to be located 5 km from the source. The doses represent the radiation exposure
from all discharges in a year and, in the case of mill tailings, the doses are associated with thoseemissions for a period of 100 years (see chapter III).
68. The largest estimated characteristic individual doses normalized to electricity generation in 2010
for all discharges come from mining and milling activities in non-ISL mines in all regions, followed byoperational discharges from nuclear power plants. Estimated characteristic individual doses from ISL
mines are about an order of magnitude smaller than those from non-ISL mines, owing to the differencesin radon discharges from the two processes. Estimated characteristic individual doses from mining andmilling are primarily associated with radon exposures, while those from nuclear power plant operations
reflect differences in the type and relative proportion of the different reactor types and thus different

164

UNSCEAR 2016 REPORT

radionuclide discharge mixes within a region, and to a lesser extent, variations in food consumption
patterns across regions. Estimated characteristic individual doses for a region normalized to electricity
generated that are associated with reprocessing are only reported for the region Europe, where theFrench, British and Russian reprocessing facilities are located. All of the estimated characteristic
individual doses are very low.
Table 14. Summary of characteristic individual doses to the public normalized to electricity
generated in 2010 for mining and milling, electricity generation from nuclear power reactors, andfuel reprocessing (mSv/(GW a))
Characteristic individual doses a,b to the public (mSv/(GW a))

Discharge type and source

Asia and
Pacific

AfricaEuropeLatin
America

North
America

ATMOSPHERIC DISCHARGESc

6.9 × 10−3

6.9 × 10−3

6.9 × 10−3

6.9 × 10−3

6.9 × 10−3—

5.5 × 10−4

5.5 × 10−4—

5.5 × 10−4

Power plants

1.3 × 10−5

5.1 × 10−5

7.2 × 10−5

1.1 × 10−4

5.8 × 10−5

Reprocessing——

5.7 × 10−5

—

—

—

—

—

Mining and milling - non-ISL mines
Mining and milling - ISL mines

AQUATIC DISCHARGES
Mining and milling - non-ISL mines—

Mining and milling - ISL mines—

Power plants

7.0 × 10

Reprocessing

—

—
—
−5

—
−5

—
−4

—
−4

8.4 × 10

6.6 × 10

1.0 × 10

4.1 × 10−5—

6.0 × 10−4——

TOTAL FOR ALL DISCHARGESc

6.9 × 10−3

6.9 × 10−3

6.9 × 10−3

6.9 × 10−3

6.9 × 10−3—

5.5 × 10−4

5.5 × 10−4—

5.5 × 10−4

Power plants

8.2 × 10−5

1.3 × 10−4

7.3 × 10−4

2.1 × 10−4

9.9 × 10−5

Reprocessing——

6.6 × 10−4——

Mining and milling - non-ISL mines
Mining and milling - ISL minesa

The characteristic individuals are those living 5 km from the points of discharge with behaviour indicative of the majority ofpeople living the area.
b

It is only appropriate to present the characteristic individual doses for discharges normalized to electricity generated and notsummations of the various individual dose values, because the same individuals cannot be exposed to all discharges from thevarious sources in each region. Because collective doses are the sum of all individual doses, they can be presented both as totalcollective dose and as collective dose normalized to electricity generated.
c

The radon emissions from mill tailings produced per unit of electricity generated are assumed to continue for 100 years.

69. Collective doses to the public for discharges from mining and milling, power plant operation andreprocessing in 2010 are summarized in table 15, and in table 16 normalized to the electricity generated
by each process for that year. The local component (for the region-averaged or world-averaged
population within 100 km) and the regional component (for the region-averaged or world-averaged
population between 100 km and 1,500 km) of the collective doses due to the atmospheric releases arepresented separately, integrated to 100 years (first pass). Radon emissions from mill tailings produced
per unit of electricity generated are assumed to continue for 100 years. The global components ofcollective doses resulting from the globally-circulating radionuclides (3H, 14C, 85Kr and 129I) are also
shown, integrated to 100, 500 and 10,000 years. For ease of comparison, table 17 shows the final

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 165

aggregated values for the total collective dose to the world public and for the collective dosenormalized to electricity generated for each of the processes mining and milling, power plant operation
and reprocessing.
70. The results in tables 15 and 16 indicate that for discharges from both power plant operation andreprocessing facilities, the global component of the collective doses due to globally-circulating
radionuclides integrated to 100 years and longer exceed the local and regional components of thecollective doses due to the initial discharge (first pass) integrated out to 1,500 km. It is important torecognize that global components are summed over the population of the entire world, which is taken as1010 people to account for growth over the next 100 years (called “whole world population” in thetables 15 and 16). This component thus is the sum of a large number of very small doses with per caput
values of the order of 10−8 Sv, and per caput values normalized to electricity generated of the order of
10−10 Sv/(GW a). Values for the collective doses in both tables 15 and 16 for the first-pass calculation
are the doses integrated to 100 years and to a distance of 1,500 km for the local and regional
components due to atmospheric discharges. Based on the results of the assessment, doses beyond thisdistance were assumed to be negligible.
71. For the nuclear industry, a significant fraction of the collective dose is due to radon discharges
from mining and disposal of mill tailings. The importance of carbon-14 discharges from nuclear power
plants and reprocessing is also evident. Table 18 shows the contribution from individual radionuclides
to the collective doses to the public, and collective doses per unit of electricity generated, from
reprocessing integrated to 100 years, not including contributions from globally-circulating
radionuclides. Figure IV is derived from the data in table 18, showing collective doses per unit ofelectricity generated for individual radionuclides discharged from reprocessing. The importance of the
carbon-14 contribution to dose is evident.
72. Figure V shows collective doses per unit of electricity generated due to the globally-circulating
radionuclides (tritium, 14C and 85Kr) integrated to 100, 500 and 10,000 years, for each reactor typeconsidered. Besides indicating the variation in the dose contribution from globally-circulating
radionuclides for the different reactor types, the graphs also illustrate the importance of 14C, whose
contribution to collective dose increases with integration time. The global circulation for 129I is notincluded in figure V because it is only discharged during the reprocessing phase.
73. Figure VI compares the collective doses to the public from all discharges, atmospheric andaquatic, from all nuclear power plants by region and by reactor type, not including contributions fromglobally-circulating radionuclides. The data indicate that HWRs and PWRs are the reactor types that
contribute most to the collective doses. Table 19 shows the total collective dose for all nuclear reactor
types summed and by region, with North America and Europe showing the highest values.

Discharge type and source

Collective dosea, b (man Sv)
AfricaAsia and Pacific

EuropeLatin America

North America

World

LOCAL COMPONENT – ATMOSPHERIC RELEASESMining and milling

No data by region

1.0 × 101

Power plants
Inland0

1.2 × 100

2.0 × 100

1.7 × 10−1

2.2 × 100

5.3 × 100

Coastal

8.2 × 10−3

2.7 × 100

1.2 × 100

1.5 × 10−2

1.2 × 10−1

3.8 × 100——

1.3 × 100——

2.1 × 100

Reprocessing

REGIONAL COMPONENT – ATMOSPHERIC DISCHARGES
Mining and milling

No data by region

4.3 × 101

Power plants
Inland0

3.0 × 100

6.7 × 100

1.3 × 10−1

5.8 × 100

1.6 × 101

Coastal

1.6 × 10−3

6.8 × 100

3.5 × 100

2.9 × 10−2

1.5 × 10−1

8.4 × 100——

4.3 × 100——

4.4 × 100

Reprocessing

LOCAL AND REGIONAL COMPONENT – AQUATIC DISCHARGES
Mining and milling

No data by region—

Power plants
Inland0

1.4 × 100

1.4 × 101

8.1 × 10−1

1.8 × 101

3.4 × 101

Coastal

9.3 × 10−5

2.5 × 10−2

6.8 × 10−2

3.5 × 10−4

1.3 × 10−3

3.9 × 10−1——

2.6 × 10-1——

1.1 × 100

Reprocessing

LOCAL AND REGIONAL COMPONENTS (ATMOSPHERIC AND AQUATIC DISCHARGES)Mining and milling

No data by region

5.3 × 101

Power plants

9.9 × 10−3

1.5 × 101

2.8 × 101

1.2 × 100

2.6 × 101

6.8 × 101

Reprocessing——

5.8 × 100——

7.6 × 100

9.9 × 10−3

1.5 × 101

3.3 × 101

1.2 × 100

2.6 × 101

1.3 × 102TOTAL LOCAL AND REGIONAL COMPONENTSc

166 UNSCEAR 2016 REPORT

Table 15. Summary of collective doses to the public due to mining and milling, electricity generation from nuclear power reactors, and fuel reprocessing (in 2010)

GLOBAL COMPONENT – GLOBALLY-DISPERSED RADIONUCLIDES

Source and integration time

Collective dose (man Sv)
Whole world population

Mining and milling integrated to
100 years—

500 years—

10 000 years—

Power plants integrated to
100 years

5.3 × 102500 years

1.0 × 103

10 000 years

5.0 × 103

Reprocessing integrated to
100 years

2.6 × 102500 years

4.9 × 102

10 000 years

2.4 × 103

Local and regional components of the collective doses (due to the first pass) are integrated to 100 years. The local and regional components are explained in chapter III. Per caput values are not given in the table, but are
discussed in the text.

bc

For all tables in the annex, the calculations resulting from assessments are made to full precision and any discrepancies in the final sum of numbers in the tables are due to rounding.

Total of the local and regional components due to mining and milling, power plants and reprocessing from the first-pass. The totals for each region include only discharges from power plants, except for Europe, which
also has a value for reprocessing.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 167aDischarge type and source

Collective dosea per unit of electricity generated (man Sv/(GW a))
AfricaAsia and Pacific

EuropeLatin America

North America

World-averageb

LOCAL COMPONENT – ATMOSPHERIC DISCHARGES
Mining and milling

No data by region

4.1 × 10−2

Power plants
Inland0

8.0 × 10−1

2.2 × 10−2

2.2 × 10−1

2.5 × 10−2

2.9 × 10−2

Coastal

5.6 × 10−3

4.3 × 10−2

3.2 × 10−2

6.8 × 10−3

8.5 × 10−3

3.3 × 10−2——

1.2 × 10−2——

8.6 × 10−3

Reprocessing

REGIONAL COMPONENT – ATMOSPHERIC RELEASESMining and milling

No data by region

1.7 × 10−1

Power plants
Inland0

2.0 × 100

7.2 × 10−2

1.7 × 10−1

6.5 × 10−2

8.9 × 10−2

Coastal

1.1 × 10−3

1.1 × 10−1

9.4 × 10−2

1.3 × 10−2

1.1 × 10−2

7.2 × 10−2——

4.0 × 10−2——

1.8 × 10−2

Reprocessing

LOCAL AND REGIONAL COMPONENTS (AQUATIC DISCHARGES)Mining and milling

No data by region—

Power plants
Inland0

9.3 × 10−1

1.5 × 10−1

1.1 × 100

2.0 × 10−1

1.9 × 10−1

Coastal

6.3 × 10−5

4.1 × 10−4

1.8 × 10−3

1.6 × 10−4

9.6 × 10−5

3.3 × 10−3——

3.0 × 10−3——

2.1 × 10−3

Reprocessing

LOCAL AND REGIONAL COMPONENTS (ATMOSPHERIC AND AQUATIC DISCHARGES)Mining and milling

No data by region

2.2 × 10−1

Power plants

6.7 × 10−3

2.4 × 10−1

2.1 × 10−1

3.9 × 10−1

2.6 × 10−1

2.3 × 10−1

Reprocessing——

5.5 × 10−2——

2.9 × 10−2

6.7 × 10−3

2.4 × 10−1

2.6 × 10−1

3.9 × 10−1

2.6 × 10−1

4.3 × 10−1

TOTAL LOCAL AND REGIONAL COMPONENTSc

168 UNSCEAR 2016 REPORT

Table 16. Summary of collective doses to the public due to mining and milling, electricity generation from nuclear power reactors, and fuel reprocessing (in 2010)
normalized to electricity generated

GLOBAL COMPONENT – GLOBALLY-DISPERSED RADIONUCLIDES

Source and integration time

Collective dose per unit of electricity generated (man Sv/(GW a))
Whole world population

Mining and milling integrated to
100 years—

500 years—

10 000 years—

Power plants integrated to
100 years

1.8 × 100

500 years

3.4 × 100

10 000 years

1.7 × 101

Reprocessing integrated tod
100 years

1.2 × 100

500 years

2.1 × 100

10 000 years

1.0 × 101

Local and regional components of the collective doses (due to the first pass) are integrated to 100 years. The local and regional components are explained in chapter III. Per caput values are not given in the table, but are
discussed in the text.
bWorld-average is used to qualify the calculations and data that are intended to represent a value averaged across the whole world. In this case the world average is the average of the population in each annulus for all
nuclear sites for which data are available (annex A). Similarly, consumption rates for terrestrial and aquatic foods are average values for the world population. The world-average values can be used for comparative
purposes including for comparison with the previous versions of the methodology for estimating public exposures due to radioactive discharges (annex A).
c

Total of the local and regional components due to mining and milling, power plants and reprocessing from the first-pass. The totals for each region include only discharges from power plants, except for Europe, which
also has a value for reprocessing.

d

The normalized values were calculated as described in paragraphs 65 and 66.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 169a170 UNSCEAR 2016 REPORT

Table 17. Comparison of the worldwide collective dose, and associated collective dose normalized
to electricity generated, from mining and milling, power plant operation and reprocessing
Nuclear fuel cycle

Collective dose(man Sv)

Normalized collective dosea(man Sv/(GW a))

Local and regional component from mining andmilling, power plants and reprocessing (first pass)b

130

0.43

100 years

910

3.0

500 years

1 700

5.5

10 000 years

7 600

25

Local and regional component (integrated to
100 years) plus global component integrated toaThese values are averages for the whole world. The world average is the average of the population in each annulus for all
nuclear sites for which data are available (annex A). Similarly, consumption rates for terrestrial and aquatic foods were averagevalues for the world population. The results can be used for comparative purposes including for comparison with the previous
versions of the methodology for estimating public exposures due to radioactive discharges (annex A).
b

Local and regional components of the collective doses (due to the first pass) are integrated to 100 years. The local and regional
components are explained in chapter III.

Table 18. Radionuclide contributions to the local and regional components of collective doses tothe public from reprocessing integrated to 100 years

Radionuclide

Collective dose per unit of electricity
generated (man Sv/(GW a))

Collective dose(man Sv)

Atmosphere

Aquatic

Total

Atmosphere

Aquatic

Total

H

1.3 × 10−3

3.6 × 10−6

1.3 × 10−3

7.1 × 10−1

8.0 × 10−4

7.1 × 10−1

14C

2.0 × 10−2

1.7 × 10−3

2.2 × 10−2

4.0 × 100

5.4 × 10−1

4.5 × 100

41

Ar

0

0

0

1.7 × 10−30

1.7 × 10−3

54

Mn

0

8.4 × 10−7

8.4 × 10−70

1.7 × 10−4

1.7 × 10−4

58

Co

0

1.8 × 10−8

1.8 × 10−80

3.6 × 10−6

3.6 × 10−660

Co

3.4 × 10−7

6.1 × 10−5

6.1 × 10−5

1.6 × 10−4

3.0 × 10−2

3.0 × 10−285

Kr

4.5 × 10−30

4.5 × 10−3

1.1 × 1000

1.1 × 100

90

Sr

0

5.8 × 10−7

5.8 × 10−7

3.0 × 10−2

9.8 × 10−4

3.1 × 10−2

106

Ru

4.2 × 10−7

9.0 × 10−5

9.0 × 10−5

1.2 × 10−3

2.8 × 10−2

2.9 × 10−2

129I

8.3 × 10−4

2.3 × 10−5

8.5 × 10−4

5.1 × 10−1

5.3 × 10−3

5.2 × 10−1

131I

9.6 × 10−7

1.3 × 10−9

9.6 × 10−7

1.1 × 10−3

2.6 × 10−7

1.1 × 10−3

135

Xe

0

0

0

3.6 × 10−40

3.6 × 10−4

134

Cs

2.0 × 10−7

1.7 × 10−6

1.9 × 10−6

3.9 × 10−5

8.1 × 10−4

8.5 × 10−4

137

Cs

1.9 × 10−7

1.7 × 10−5

1.7 × 10−5

6.5 × 10−3

1.8 × 10−2

2.5 × 10−2

239

Pu

3.7 × 10−6

8.3 × 10−5

8.7 × 10−5

1.6 × 10−1

2.5 × 10−1

4.2 × 10−1

241

Am

0

5.7 × 10−5

5.7 × 10−5

1.2 × 10−2

2.0 × 10−1

2.2 × 10−1

Total

2.7 × 10−2

2.1 × 10−3

2.9 × 10−2

6.5 × 100

1.1 × 100

7.6 × 100

3

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 171

Table 19. Local and regional components of collective dose integrated to 100 years summed over
all reactor types

Collective dose (man Sv)

AfricaAsia and
Pacific

EuropeLatin
America

North
AmericaWorld

9.9 × 10−3

1.5 × 101

2.8 × 101

1.2 × 100

2.6 × 101

6.8 × 101

Figure IV. Contribution of radionuclides to the local and regional components of collective dose to
the public from reprocessing integrated to 100 years and normalized to electricity generation
See data in table 18

172 UNSCEAR 2016 REPORT

Figure V.

Collective doses from globally-circulating radionuclides per unit of electricity generated

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 173

Figure VI. Local and regional components of the collective doses to the public due to total estimateddischarges from nuclear power plants (2010) integrated to 100 years, by UNEP region and by reactor type4.

Uranium enrichment and fuel fabrication

74. The discharges from the processes of conversion, uranium enrichment and subsequent fuelfabrication needed for different reactor types have generally been determined in earlier studies to berelatively small, consisting mainly of radionuclides within the decay chains of uranium isotopes. The
collective dose to the public estimated on the basis of the electricity generated, derived in the
UNSCEAR 2000 Report [U9] and re-used in the UNSCEAR 2008 Report [U11], were used here toobtain an updated value based on the 2010 electricity generation from nuclear power. This value for
uranium enrichment and fuel fabrication facilities is 0.003 man Sv/(GW a). The Committee earlier
estimated the average annual collective dose to the public from all plants for the period 1998–2002 to
be 0.8 man Sv [U11]. Updating this value to the electricity generated from nuclear power in 2010 fromtable 2, the average annual collective dose to the public due to uranium enrichment and fuel fabricationin the year 2010 is about 0.9 man Sv.

174 UNSCEAR 2016 REPORT

5.

Solid waste disposal

75. Solid wastes arise at various stages in the nuclear fuel cycle. These include low- and intermediatelevel wastes, mainly from reactor operations, high-level wastes from fuel reprocessing, and spent fuelfor direct disposal. The activity concentrations of these wastes can range from just above natural
background levels, as in mill tailings, to much higher levels, such as in spent reactor fuel.
76. Low- and intermediate-level wastes are generally disposed of by shallow burial in trenches orconcrete-lined structures, but more advanced disposal sites also exist. Before disposal, all such material
is manipulated and transported, which gives rise to both occupational and public exposures. For the
nuclear fuel cycle, doses to members of the public from the transport of radioactive material have been
earlier estimated using the factor of 0.1 man Sv/(GW a) for the collective dose per unit of electricity
generated [U11]. Wastes from the nuclear fuel cycle also include large quantities of depleted uranium
from uranium enrichment operations [U18]. High-level wastes and spent fuel are presently retained ininterim storage facilities until adequate methods for disposal have been devised and disposal sites
selected [U11].
77. Estimates of the doses due to solid waste disposal have been based on the projected eventual
migration of radionuclides from the burial site into groundwater. The collective dose due to low- andintermediate-level waste disposal normalized to electricity generated is estimated to be approximately
0.5 man Sv/(GW a), due almost entirely to 14C. The average worldwide per caput annual effective dosewould be about 1 nSv per year of practice [U11].

C. Occupational exposure for the nuclear fuel cycle
78. The assessment of occupational exposures for the nuclear fuel cycle considers the practices ofmining, milling, uranium enrichment and conversion, fuel fabrication, reactor operation, fuelreprocessing and research.
79. Collective doses due to occupational exposure for each practice in the nuclear fuel cycle that werereported in the UNSCEAR 2008 Report [U11] for the years 2000–2002 were used as the starting point
to estimate the 2010 occupational exposures. This approach was taken since data were not readily
available on the total number of, or the average annual effective dose to, monitored workers worldwide
in 2010. As reported in the UNSCEAR 2008 Report, the Committee used the amount of electricity thatcan be generated from each practice (i.e. how much electricity can be generated from the uranium thatis mined, milled, converted, and so on) to normalize collective doses to the electricity generated.
80. Table 20 summarizes the data on occupational exposure for the nuclear fuel cycle in the period
2000−2002 that are relevant for estimating occupational exposure for the nuclear fuel cycle in 2010.
81. The ratio from the 2000–2002 data for the average annual collective dose per unit of electricity
generated was combined with the electricity generation for 2010 to estimate the annual collective dosefor 2010. Similarly, the ratio obtained from 2002 for the average annual collective dose per unit mass ofuranium mined, which can be further transformed to the collective dose per unit of electricity generated
is assumed appropriate for the mining process. Actual data from the Canada, the United Kingdom andthe United States for 2010 were obtained on the number of monitored workers and their average annual
effective dose, and were compared with the respective country values for 2000−2002 in the UNSCEAR
2008 Report [U11]. These comparisons showed reasonable agreement with the scaled average annual
effective dose, giving some confidence in the assumptions used.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 175

Table 20. Worldwide average annual individual and collective doses to workers due to the commercial
nuclear fuel cycle
From table 72 in the UNSCEAR 2008 Report [U11]; for the years 2000–2002a

Practice

Monitored
workers(thousands)

Average annualb
collective dose(man Sv)

Average annualb collective
dose per unit of electricity
generated(man Sv/(GW a))

Average annualb
effective dose tomonitored workers(mSv)

Mining

12

22

0.1

1.8

Milling

3

3

0.02

1.0

Enrichment

18

2

0.02

0.1

Fuel fabrication

20

31

0.1

1.6

Reactor operation

437

617

2.5

1.4

Reprocessing

76

68

Research

90

36

0.1

0.4

Total

656

779

2.8

1.2

c,d

0.9

a

Some values in the table have been corrected since table 72 was published in the UNSCEAR 2008 Report [U11]; for the years
2000–2002.
b

The words “average annual” used in the table denote the annual dose averaged over the 2000-2002 year span.

c

Also includes the reprocessing of some fuel associated with military application.

d

The average annual collective dose per unit of electricity generated was not possible to obtain, because there were no datareadily available on the amount of electricity generated from the reprocessed fuel.

82. Table 21 shows the current results for 2010 for the worldwide average annual collective dosecompared to the 2000–2002 results. The procedure for estimating the values for each practice for 2010is explained below.
−

Uranium mining. The amount of uranium ore mined was assumed equal to the amountproduced [U11]. Total uranium production in 2010 for 16 countries is shown in table 4 [O8].
Values were summed for all countries to give 54,100 t of uranium. Of this amount, 31,900 t
was produced from open pit and underground uranium mines and 22,200 t was produced fromin-situ leaching (ISL) processes. Assuming the same collective dose per unit mass of uranium
extracted of 0.623 man Sv/kt as for 2000–2002 [U11], then the collective dose (rounded) is
20 man Sv. This value excludes uranium produced from ISL.

−

Uranium milling. As for uranium mining, the total uranium production in 2010 was taken fromtable 4 [O8] and assumed to equal the total amount of uranium milled, 54,104 t. Applying thesame simplifying assumptions used in the UNSCEAR 2008 Report [U11] (i.e. that all milled
uranium is used in LWRs and that the uranium requirement is 220 t/(GW a) ([U11]; table 18-A)),the equivalent amount of electricity for 2010 is 246 GW a. The collective dose (rounded) for2010 is then 5 man Sv. While ISL extract is not milled, it does go through some treatment toproduce yellowcake. Thus the full amount of uranium mined, 54,100 t, is used here.

−

Uranium conversion and enrichment, and fuel fabrication. UNSCEAR 2008 Report (table 15)
shows the worldwide installed capacity for fuel cycle installations taken from the IAEA
Nuclear Fuel Cycle Information System, NFCIS. Compiling the data from NFCIS for 2010
and comparing with table 15 from the UNSCEAR 2008 Report shows little has changed in the176 UNSCEAR 2016 REPORT

capacity of these processes. All three processes show decreased capacities of between 8% and
9%. Using the assumption that these practices give the same collective dose per unit ofelectricity generated in 2010 as in 2002, the average annual collective dose for 2010 intable 21 has been adjusted to a lower value by 8%.
−

Reactor operation. The normalized average annual collective dose for 2000–2002 was
2.5 man Sv/(GW a). Instead of using this value directly, as explained above, data from theOECD/NEA [O4] indicated a general decrease in the average annual collective doses. The
reported average annual collective dose for operating nuclear power plants fell by 20%
between 2002 and 2010, attributed to additional operating experience and the global exchangeof best radiation protection practices. Using this, the normalized average annual collective
dose for 2010 can be taken as 2.0 man Sv/(GW a). Assuming the total worldwide electricity
generation in 2010 is 314 GW a, an average annual collective dose of 628 man Sv wascalculated, about the same as that reported for 2002, as shown in table 21.

−

Reprocessing. In the absence of updated data on occupational exposures during fuelreprocessing in 2010, it was assumed that reprocessing activities have remained more or lessconstant and the same collective dose was assigned as previously, i.e. no change. Data fromUNSCEAR [U11] were not presented in terms of dose per unit of electricity generated.

−

Research. In the absence of updated data on occupational exposures obtained during researchfor 2010, it was assumed that research activities have remained more or less constant and thesame collective dose was assigned as previously. However, it should be noted that a slightdownward trend in collective dose was observed over the previous two time periods in
UNSCEAR [U11]. Data from UNSCEAR [U11] were not presented in terms of dose per unitof electricity generated.

Table 21. Worldwide levels of exposure of workers due to the commercial nuclear fuel cycle for 2010 and 2000–2002
Practice

Average annual collective dose2010 (man Sv)(Estimates from this study)

Average annual collective dose per unit ofelectricity generated (man Sv/(GW a))(From [U11] for years 2000–2002)

Remarks

Mining

22

20

0.1

U production 2010=54.1 kt; of this 31.9 kt was the amount of
U produced excluding ISL production (35 ktU average in
1998–2003)

Milling

3

5

0.02

246 GW a calculated using same simplifying assumption
from UNSCEAR 2008 Report, table 18-A, that all milled
uranium was used in LWRs, and that uranium requirement
was 220 t/(GW a). Doses resulting from subsequent
remediation activities were not accounted for

Enrichment

2

2

0.02

8% lower capacity than in 2000–2002 (not noticeable
because numbers are rounded)

Fuel fabrication

31

29

0.1

8% lower throughput than in 2000–2002

Reactor operation

617

628

2.5

314 GW a total electricity generated by nuclear power plants
in 2010 [I15, I16]

Reprocessing

68

68

Research

36

36

0.1

Total

779

788

2.8

Assumed constant collective dose–i.e. no change. Data were
not presented [U11] in terms of dose per unit of electricity
generated
Assume constant collective dose–i.e. no change

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 177

Average annual collective effective
dose 2000–2002 (man Sv)(From [U11] for years 2000–2002)

178 UNSCEAR 2016 REPORT

D. Decommissioning
83. When a power company decides to close a nuclear power plant permanently, the facility must be
decommissioned by safely removing it from service and reducing residual radioactivity to a level that
permits release of the property and termination of the operating licence. As more commercial nuclear
power reactors reach the end of their operating licence, there is a commensurate increase indecontamination and decommissioning activities that involve radiation exposure. These activities
include decontamination of structures and components, dismantling of components and demolition ofbuildings, remediation of any contaminated ground, and removal of the resulting waste. As of 2015,
156 commercial and prototype reactors located in 19 countries with a total of 60.9 GW(e) capacity werepermanently shut down [I8]. Most of these reactors were PWRs, GCRs and BWRs. The current
assessment concentrated on occupational exposures during the decommissioning process.
84. The decommissioning process begins when a power company decides to permanently cease
operations. The operator of a nuclear facility may choose between three decommissioning strategies:(a) immediate dismantling, (b) deferred dismantling after a safe storage period, and (c) entombment of
the facility. Under immediate dismantling, decommissioning activities begin shortly after thepermanent cessation of operations. This strategy implies prompt completion of the decommissioning
project (approximately 10 to 20 years) and involves the removal of all radioactive material from thefacility to another new or existing licenced facility.
85. Under deferred dismantling (sometimes called safe storage, safe store or safe enclosure), a nuclear
facility is maintained and monitored in a condition that allows the radioactivity to decay and thus
reduce occupational exposure. Parts of a facility containing radioactive contaminants are safely stored
and maintained for upwards of 40 to 60 years until they can subsequently be decontaminated to levels
that permit parts of the facility to be dismantled and released for unrestricted use.
86. Under entombment, radioactive contaminants are encased in a structurally long-lived material
such as concrete until radioactivity decays to a level permitting the unrestricted release of the facility,
or release with restrictions imposed by the regulatory body.
87. As of December 2015, the International Atomic Energy Agency/Nuclear Energy Agency
Information System on Occupational Exposure (ISOE) database (http://www.isoe-network.net/)contained data on doses to workers from 84 reactors that were shut down, in some stage ofdecommissioning or fully decommissioned. These reactor units were generally of different types andsizes, and at different phases of their decommissioning programmes. At least 15 reactors had been fully
dismantled, over 50 reactors were being dismantled, over 50 reactors were in deferred dismantlement,
three had been entombed, and for others the decommissioning strategy was not specified yet.
88. Table 22 provides the average annual collective dose for occupational exposure per unit for up to
71 permanently shut down reactors by country and reactor type for 2008−2013, based on data recordedin the ISOE database and supplemented by individual country reports. Different decontamination anddecommissioning strategies were being employed and each facility shown in the database was in adifferent stage of decommissioning, making definitive trends difficult to deduce from this information.
Depending on the phase of decommissioning, there may be little or no worker exposure one year andthe next year collective dose to workers may rise as much as 100-fold (adapted from [O5]). Considering
the average collective doses for each of the reactor types, however, and the total average for all theunits, the average values demonstrate some stability over these six years, with the total average for allunits being about 0.06 man Sv per year and reactor.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 179

89. Table 23 shows data from the whole decommissioning period for nuclear power plants thathave been immediately decontaminated and decommissioned. The dose values given are theintegrated doses for the whole decommissioning process. There may be some additional dose due tospent fuel storage until a permanent repository is opened, but these occupational exposures can beconsidered negligible compared to active decommissioning. For example, for five of thedecommissioned units that reported to the REIRS database (Radiation Exposure and Information
Reporting Systems) ([U24]; table 3.1), collective doses from occupational exposure due to spent fuelstorage ranged from 0 to 1.86 × 10−3 man Sv annually. The values for electricity supplied were thetotal integrated electricity generation from the time the unit went on the grid until it ceased operation.
90. The relative quantity needed for the comparative study conducted here was an estimate of thecollective doses from occupational exposure for decommissioning, integrated over the period of
decommissioning. The fact that the annual collective dose varies from year to year is not relevant if areasonable average can be obtained. Although limited, the Committee considered that the data on
integrated doses for nuclear reactors shown in table 23 were sufficient for its use here.
91. The average total collective dose associated with the strategy of immediate decontamination anddecommissioning of the commercial nuclear power plants shown in table 23 was less than 10 man Sv
per reactor. While the alternative decommissioning strategies, deferred dismantling and entombment,
may result in some reduction in the total collective dose per reactor from decommissioning, it may be
another 10 to 20 years before any definitive conclusions can be documented.
92. Table 23 also shows the collective dose from occupational exposure normalized to the total
integrated electricity supplied for each of the reactors, with an average for these 8 reactors of
1.80 man Sv/(GW a).

The columns headed dose give the average annual collective dose from occupational exposure in man-millisieverts per reactor (man mSv)

2008

PWR

VVER

BWR

2009

2010

2011

2012

2013

No.

Dose

No.

Dose

No.

Dose

No.

Dose

No.

Dose

No.

Dose

France

1

23.2

1

62.1

1

117.2

1

264.1

1

275.6

1

189.3

Germany

5

160.0

5

128.0

3

278.6

3

126.3

3

114.4

4

77.7

Italy

1

1.1

1

1.7

1

3.2

1

1.8

1

3

1

5.2

Spain

1

134.7

1

244.0

1

53.0

1

190.0

1

307.9

1

468.9

United States

10

7.1

8

1.5

8

2.4

6

49.4

6

127.1

12

47.3

Average

18

57.2

16

60.0

14

73.5

12

94.4

12

141

19

81.1

Bulgaria

4

31.0

4

29.4

4

11.3

4

9.2

4

10.1

4

3.3

Germany

5

27.0

5

20.0

—

—

—

—

—

—

—

—

Russian Federation

2

78.0

2

84.0

2

77.6

2

66.3

2

79.2

2

49.6

Slovak Republic

1

48.2

2

106.0

2

12.4

2

10.0

2

4.3

—

—

Average

12

38.6

13

46.0

8

28.2

8

23.7

8

25.9

6

18.7

Germany

3

179.0

3

138.0

1

427.1

1

289.5

1

88.2

1

72.0

Italy

2

29.1

2

61.8

2

60.3

2

15.1

2

18.4

2

34.2

Japan

—

—

—

—

2

123.8

2

48.4

2

41.2

2

64.2

The Netherlands

1

0.3

1

0.6

n/a

n/a

1

10

—

—

—

—

Sweden

2

39.1

2

27.0

2

6.2

2

27.2

2

20

2

3.45

United States

3

13.4

4

5.1

5

21.8

4

30.7

4

59.4

5

55.7

Average

11

64.9

12

51.1

12

76.3

12

50.3

11

44.1

12

46.2

180 UNSCEAR 2016 REPORT

Table 22. Number of units and average annual collective dose from occupational exposure per reactor by country and reactor type for definitely shutdown reactors, 2008–2013

2008

2009

2010

2011

2012

2013

Dose

No.

Dose

No.

Dose

No.

Dose

No.

Dose

No.

Dose

France

6

2.8

6

8.8

6

1.3

6

2.4

6

7.4

6

8.2

Germany

2

13.0

2

17.0

—

—

—

—

—

—

—

—

Italy

1

2.9

1

0

1

1.7

1

10.4

1

0.2

1

2.2

Japan

1

20.0

1

20.0

1

50

1

50

1

70

1

10

Spain

—

—

—

—

—

—

1

0

1

0

1

0

United Kingdom

16

48.0

16

42.0

16

55

16

49

19

56

19

57.3

United States

—

—

—

—

—

—

—

—

1

0

1

0

Total number,average dose

26

32.1

26

30.0

24

39.1

25

34.4

28

42.1

28

41.1

HTGR

Germany

1

0

1

0

1

0

1

0

—

—

1

0

FBR

United States

—

—

1

80.14

1

77.9

1

294.9

1

2

1

0.1

LWGR

Lithuania

1

188.4

1

144.7

2

236.2

2

304.8

2

264.9

2

304.8

LWCHWR

Japan

1

431.3

1

114.6

1

111.6

1

126.6

1

148.8

1

134.1

All Units

Average

70

52.3

71

47.7

63

59.9

62

61.5

64

66.3

71

57.8

GCR

This table is adapted from table 4 in the 2012 ISOE annual report [O5]and table 5 from the 2010 ISOE annual report [O4]. Additional information was obtained from the individual country reports and the ISOE database,and the United States Nuclear Regulatory Commission Radiation Exposure and Information Reporting System (REIRS) [U24] database for United States definitely shut reactors. Data for 2013 were obtained from the2013 ISOE country reports [O7] and the ISOE database [I20].
Dashes (—) represent missing or partial data in the data source. The acronyms for the reactor type used in tables 22 and 23 differ somewhat from those in the other tables in this annex (see paragraph 58). The acronyms inthis table and table 23 are the designations used by the Nuclear Energy Agency.
The following explains the meaning of the acronyms that differ from previous tables:
LWCHWR (Light Water Cooled Heavy Water Reactor) also known as a CANDU reactor is the Canadian designed system. IAEA would refer to this as a PHWR.
VVER (water, water, energetic reactor) is a Soviet designed PWR.
HTGR, (also called HTGC, high temperature gas cooled) reactor. More of a prototype reactor but nuclear power plants include Peach Bottom and Fort Saint Vrain in the United States; Dragon reactor in the United
Kingdom, and THTR-300 and AVR in Germany.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 181

No.

182 UNSCEAR 2016 REPORT

Table 23. Collective dose, and collective dose normalized to electricity supplied, due to the complete
decommissioning process of immediately decontaminated and decommissioned commercial nuclear
power plants
Reactor gross electrical capacity and total electricity supplied were obtained from the IAEA PRIS database [I7]. Collective
doses were obtained from the ISOE database [I20] and REIRS reports [U24]Reactortype
PWR

Country

Plant

Gross
Capacity(MW(e))

Collective
dose(man Sv)

Electricity
supplied(GW a)

Normalized
collective dose(man Sv/ GW a)

United States

Haddem Neck

603

8.457

12.1

0.70

Maine Yankee

900

6.195

13.6

0.46

Rancho Seco

917

2.345

5.1

0.46

San Onofre 1

456

3.002

5.8

0.52

Trojan

1 155

2.973

9.6

0.31

Yankee Rowe

180

6.467

3.9

1.66

4.907

8.34

0.69

AVERAGE
BWR

United States

Big Rock Point

71

5.703

1.45

3.93

HTGR

United States

Fort St. Vrain

342

3.961

0.62

6.39

All units

Average

4.900

1.80

V. RADIATION EXPOSURES ARISING FROM ELECTRICITY
GENERATION BY COMBUSTION OF FOSSIL FUELS

A. Coal1.

Introduction

93. Coal is a family name for a variety of solid organic fuels and refers to a whole range of
combustible sedimentary rock materials spanning a continuous quality scale. For convenience, thiscontinuous series is often divided into two main categories: hard coal (which includes anthracite andbituminous coal) and brown coal (which includes sub-bituminous coal and lignite). The International
Energy Agency [I15] makes use of two broad categories of coal; hard coal as having a gross calorificvalue not less than 23.9 GJ/t (5,700 kcal/kg) and brown coal with a gross calorific value less than
23.9 GJ/t (5,700 kcal/kg). Often coal data are presented in tonnes of coal equivalent (TCE) where one
tonne of coal is equivalent to 7 million kilocalories. This description standardizes the carbon content
and heat value of a particular type of coal. The International Energy Agency estimates that there are
equal recoverable global reserves of both hard coal and brown coal [I14].

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 183

94. Combustion of coal for electricity generation is the largest contributor to worldwide electricity
generation (figure I). Coal has maintained a share of about 40% of the total electricity generated forsome decades, while the total electricity generation from coal and other sources has increased steadily
as shown in figure I [I15]. The UNEP geographic region—Asia and the Pacific—accounted for 57% of
the total electricity generation from coal in 2010, with China accounting for about 70% of this [U17]. In
contrast, the geographic region—West Asia—used no coal for electricity generation in 2010 (table 2).
95. Electricity generation is not the only end-product of coal combustion, others being for example
steel production and cement manufacturing. The largest 10 coal-producing countries are shown for theyears 2008–2012 in table 24. China was the largest producer with a 30% increase in production overthis five-year period. The total world primary coal production for this period had a 16% increased
production in 2012 compared with 2008, and about a 60% increased production during the 10-year
period since 2002 [U17].
96. Coal contains naturally occurring radionuclides from the uranium and thorium series (figure III)and potassium-40. The concentration of naturally occurring radioactive material in coal depends on thecharacteristics of the geological formation of the coal seams where it originated. Table 25 provides arepresentative overview of the range in concentrations, where large variations in the activity
concentrations can be noted.
Table 24. Total primary coal production 2008–2012 (million metric tonnes, Mt)
Source: United States Energy Information Administration, International Energy Statistics 2008–2012, open source data [U17]Country

2008

2009

2010

2011

2012

China

2 811

2 995

3 230

3 518

3 645

United States

1 063

975

984

994

922

India

517

558

562

575

589

Australia

392

408

424

402

421

Russian Federation

305

276

322

322

354

South Africa

252

249

255

253

259

Indonesia

249

291

325

360

443

Poland

143

135

133

139

144

Kazakhstan

111

101

111

116

126

Colombia

74

73

74

86

89

World

6 778

6 896

7 257

7 660

7 881

184 UNSCEAR 2016 REPORT

Table 25. Example ranges and/or averages of radionuclide activity concentrations in coal (Bq/kg)(table VII from IAEA [I2])Country

U

238

Th

230

226

21–68

Ra

Pb

Po

Th

Ra

210

210

232

228

19–24

20–33

16–28

11–69

11–64

72

62

62
8

K

40

Australia

8.5–47

Brazila

72

72

Egypt

59

26

8

10–145

10–63

10–700

32a

21a

225a

Lignite

<1–58

<1–58

<4–220(Former
Democratic
Republic of
Germany)

10a

8a

22a

Germany

Greeceb

117–390

Hungary

20–480

12–97

30–384Italyc23±3

18±4

218±15

Poland

<159
18d

<123
11d

<785

Romania<415
80a

UnitedKingdom

7–19

United States

6.3–73aAverage.

b

Lignite.

cLignite, average.

d

Average for all coal seams.

44–206<557
126a
8.5–25.5

59–205

23–140

<510
210a

9–41

<580
262a

7.8–21.8
8.9–59.2

<170
62a
7–19

12.2–77.7

3.3–51.8

55–314

3.7–21.1

97. When coal is combusted the majority of the non-combustible matter remains in the bottom and flyash. Because radionuclides are present in the mineral constituents, they also tend to remain in the ash,where the concentration of the radionuclides becomes enhanced [I2, S4, Z1]. However, radon is a noble
gas and all the radon present in the coal is emitted through the flue-gas stack. As coal-fired power
plants have modernized, the cleaning and filtering systems have become increasingly efficient andlarger fractions of the particulate matter in the fly ash have been captured and removed from thedischarge to atmosphere. The coal ash that is collected is either recycled for beneficial use or it isdisposed of in landfills or impoundments. 6
98. The components of the coal cycle that can lead to radiation exposures of the public are(a) discharges to atmosphere of radon from coal-mining activities and other wastes produced during
mining, (b) radioactive discharges from the operation of coal-fired power plants, and (c) radiation
exposures from the recycling and use of, or disposal of, coal ash. Occupational radiation exposures
6

Coal ash can be disposed as a coal slurry in impoundments, a dammed reservoir that contains the coal slurry.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 185

occur during (a) the mining for coal, and (b) work performed associated with power plant operation anddisposal operations. In earlier studies, occupational exposures from work performed at the power plant
were determined to be small [C5, G4, G5] in relation to the occupational radiation exposures incurred
during coal mining. For this reason, combined with the relatively small number of workers involved
compared to coal mining, this exposure source was not considered further in the current study. The
following sections consider the other exposure sources and pathways.

2.

Public exposure(a) Coal mining
99. Radon emissions from coal mining contribute to exposure of the public from the coal cycle.
Radon gas is emitted from the mines as a result of the coal mining activities, primarily through ventingof the mines. Although data that could enable the assessment of this impact have earlier been scarce,
new measurements on radon emissions from coal mining in China have become available. Radonemissions from 23 large-, medium- and small-sized coal mines in eight different provinces in China
have been assessed, based on the monitoring results of radon activity in the mines. The large- andmedium-sized coal mines are equipped with ventilation systems that are described as “good” or“relatively good.” The smaller, privately owned coal mines generally are poorly ventilated, with radon
concentrations 15 to 30 times higher than in the larger mines [L1]. The radon emissions have beennormalized to the unit of coal output and per unit of electricity generated for each coal mine and foreach size-class of coal mine. Typical values were obtained as shown in table 26 [W2].
Table 26. Radon emissions during one year from large-, medium- and small-sized coal mines in
China [W2]Type of coal mine

Coal output(t/a)

Number ofcoal mines

Radon emissions
per unit of coal
output (Bq/t)

Radon emissions per unit ofelectricity generated(TBq/(GW a))a

Large-sized

1.6 × 1097

1.9 × 105

0.57

Medium-sized

8

5.7 × 104

6

1.3 × 10

3.6

Small-sized

3.9 × 10812

3.5 × 10611

9.3 × 105

2.8

Typical valuea

From the data in table 26, obtained from the Chinese delegation to the Committee, the coal production per unit of electricity
generation could be calculated as between 2.8 and 3.1 × 106 t/(GW a).

100. The radioactive content of coal-mine tailings do not differ significantly from backgroundconcentrations of natural radionuclides in soil; this situation is different from uranium mining and milltailings, which have higher concentrations of the natural radionuclides [U9]. This exposure pathway isthus not considered further in this analysis.
101. The typical value for the radon emission (based on discharges of radon during one year)
normalized to electricity generated obtained from the Chinese data, 2.8 TBq/(GW a), has been usedwith the electricity generation data from table 2 and the Committee’s revised methodology for
estimating public exposures due to radioactive discharges (annex A) to assess the radiological impact
from this source. The default population distribution (see section III.A) was used and the radon was
assumed to be released on its own with account taken of its short-lived progeny (218Po, 214Pb, 214Bi, and
214Po) in estimating the doses from inhalation.

186 UNSCEAR 2016 REPORT(b) Electricity generation from combustion of coal
102. Coal combustion in power plants results in the direct release of gaseous radionuclides, and the
production of ash with enhanced concentrations of natural radionuclides relative to those of coal [S4].
A fraction of the ash produced is released to atmosphere, dependent on the particulate collection
devices used in the power plant. Modern coal-fired power plants have more efficient particulate
collection devices than older coal-fired power plants.
103. To assess the doses due to atmospheric releases from coal-fired power plants, a representative
case was considered. The coal (e.g. hard coal) was assumed to have an energy equivalent of 24 GJ/t
with a power generation efficiency of 38% [I13]. Although the values of 238U concentrations shown intable 25 show wide variations, an average 238U concentration of 20 Bq/kg was chosen, based on typicalvalues for Chinese coal reported as 10–25 Bq/kg [I10]. China is the largest producer of coal worldwide,
so the value of 20 Bq/kg for the 238U concentration in coal was considered reasonable as a typical
average global value. Other studies have also found a normal range of 20–25 Bq/kg observed globallyfor good quality coal [S4, T1, U6]. A value of 20 Bq/kg for the 238U concentration in coal had also beenused in the UNSCEAR 1988 Report [U7]. It should also be noted that the resulting individual andcollective doses calculated by the revised methodology for estimating public exposures due toradioactive discharges (annex A) are directly proportional to the 238U concentration in coal.
104. Two different release rates were evaluated to represent the discharge characteristics of older coalfired power plants and modern coal-fired power plants (table 27). The relative distribution among thedifferent radionuclides in table 27 for the older plants were taken from Hedvall and Erlandsson [H2]while the same for the modern plants were based on Zeevaert et al. [Z1]. The key difference between
the two data sets is that 222Rn concentrations in the discharges are 100 times larger than 226Ra in older
plants and 1,500 larger than 226Ra in modern plants that have more efficient filtering systems. However,in both cases the value for the discharge of lead and polonium particles has been chosen as twice that ofuranium and radium [C4, T1]. Because radon is a noble gas, it is unaffected by filters. As seen fromtable 27, these releases are not in secular equilibrium. The distribution of older versus modern coalfired power plants in the various regions of the world is not well-characterized so the results from bothcases are included to illustrate the difference in the resulting doses.
105. The assessments presented in this annex using the Committee’s revised methodology for
estimating public exposures due to radioactive discharges (annex A) do not include dose contributions
from 40K or the 232Th decay chain. The absorption of 40K in humans is homeostatically controlled andtherefore any doses from 40K discharges have been neglected. The contribution of the 232Th decay chain
to exposures due to electricity generation from coal combustion has been studied by Zeevaert et al.
[Z1]. They found that the concentrations of the radionuclides in the 238U and 232Th decay chains in fluegases discharged were similar. The largest contribution to individual doses came from deposition, with
consumption of food crops dominating. In this category, the contribution from the 238U decay series was
one order of magnitude larger than the contribution from the 232Th decay series. Although inclusion of
the 232Th decay series may lead to somewhat higher doses from electricity generation due to coal
combustion, it was excluded from the analysis in this annex.
106. Aquatic discharges from coal-fired power plants may lead to additional exposure of local
population groups, but this contribution has been found to be small and site-specific [L2, Z1], and were
therefore not considered here.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 187

Table 27. Releases to atmosphere (during one year) normalized to electricity generated from coalfired power plants, based on a representative coal containing 20 Bq/kg of 238U
Radionuclide

Normalized release (TBq/(GW a))
Older coal-fired power plantsa

Modern coal-fired power plantsb

222

Rn

0.07

0.07

210

Po

1.4 × 10−3

9.3 × 10−5

210

Pb

1.4 × 10−3

9.3 × 10−5Ra

0. 7 × 10−3

4.7 × 10−5

234

U

0. 7 × 10−3

4.7 × 10−5

238

U

0. 7 × 10−3

4.7 × 10−5Th

0. 7 × 10−3

4.7 × 10−5

226

230

a

Relative radionuclide distribution source: [H2].

b

Relative radionuclide distribution source: [Z1].

(c) Radon discharges from coal ash disposed in landfills
107. Coal ash is also referred to as coal combustion residuals (CCR) and coal combustion products(CCP) depending on the industry and the country, and is one of the larger streams of industrial waste
generated in the United States [U21]. Coal ash includes a number of by-products from combustion ofcoal including: fly ash; (furnace) bottom ash; fluidized bed combustion ash; boiler slag; semi-dry
absorption product; and flue gas desulphurization gypsum.
108. The coal ash collected from a power plant is either recycled for beneficial use or it is disposed ofin landfills or impoundments. Figure VII shows a time history of fly ash production and the fractionused for construction in the period 1966–2010 in the United States [A1]. Approximately 40% was usedcommercially, leaving the remaining 60% for disposal. This does not include bottom ash, but the
proportion sent for disposal was similar [A2].
109. Conventional back-filling or earthmoving operations are used to dispose of the coal combustion
residues (CCR) in dry or slightly moist conditions. Impoundments represent a wet disposal method
where the CCR is mixed with water at the power station and conveyed hydraulically through pipelines
to artificial lagoons where the slurry is discharged. According to the Electric Power Research Institute[E3], about 60% of the disposed fly ash in the United States is managed dry in landfills, and 40% ismanaged wet in impoundments, and it noted that there was a long-term trend toward increased use of
dry management practices [E3].
110. Earlier studies concluded that the individual doses to the public due to radioactive discharges
from deposited coal ash were below values relevant from an individual radiation protection viewpoint(e.g. [U23]). The main reason was that the fraction of radon emanating from coal ash is small. This
pathway was however considered relevant in the current assessment for collective doses given that theamount of coal ash produced is large and could be significant to the comparisons of collective doses.
Also, a similar pathway was considered from the uranium mill tailings in the nuclear fuel cycle
motivating a comparison between these two pathways.

188 UNSCEAR 2016 REPORT

Figure VII.

Fly ash production and use in the United States (1966–2010) [A1](Blue = production, Red = used as a construction material in a variety of applications)111. Method for treating the radon discharge from coal ash deposits. The amount of radon released toatmosphere from coal ash depends on the activity concentration of radon in the ash, which is related tothe activity concentration of 226Ra, and the extent to which the radon produced in the ash is emitted toatmosphere. A recent comprehensive review of radon emanation measurements for mineral, rock, soil,
mill tailings and fly ash found the radon emanation fraction (the ratio of the radon emitted to thatgenerated in the mill tailings (or ash deposits) per unit volume) for mill tailings (mostly uranium milltailings) was 0.17 and for fly ash was 0.03 [S1]. In another study [S6] the radon emanation fraction forcoal ash was found to be between 0.08 and 0.13 [S6]. In summary, the first cited study estimated aradon emanation from coal fly-ash of about 20% of the radon emanation from uranium mill tailings,
and the second study estimated a value of about 50%. A representative average concentration of 226Ra
in coal ashes (fly- and bottom-ash) was taken as 100 Bq/kg [M2].
112. The 222Rn flux (Bq/(m2 s)) was studied using data from a repository for uranium residues in India[I5] and a linear relationship was found between the radon flux and the activity concentration of 226Ra
in the source:
222

Rn flux (Bq/(m2 s)) = 8.3 × 10-4 kg/(m2 s) × 226Ra concentration (Bq/kg)(1)113. Because the relationship between the radon flux and the radon emanation factor is linear, theconstant in the above equation can be assumed to be proportional to the ratio of the radon emanation
factors for different sources. Therefore, for fly ash the constant above could be reduced to between

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 189

20 and 50% of its value (comparable with the values from the studies cited above for the radon
emanation from coal ash as compared to uranium mine tailings) to obtain an estimate of the 222Rn flux
from coal ash based on the 226Ra concentration in coal ash. A value of 20% was used in this analysis,
giving a linear constant in the above relationship of 1.7 × 10−4 kg/(m2 s).
114. According to a survey in the United States regarding the disposal of solid waste from coal-fired
power plants, a value for the area needed for disposing of solid wastes normalized to electricity
generated was reported as 2–11 m2/(GW h) [A9, F3]. A central value of 6 m2/(GW h) was used for thisanalysis. Multiplying the radon flux during one year, obtained by using equation (1), with the linear
constant adjusted for coal ash and a 226Ra concentration of 100 Bq/kg, with the area needed for
disposing of solid wastes normalized to electricity generated, gave the source term for the annual
discharge of 222Rn from deposited coal ashes of about 0.03 TBq/(GW a).
115. A similar value was reached by an independent method. This assessment used an annual dischargeof 0.1 TBq/(GW a) as the source term for radon discharges from uranium mill tailings, one part of thenuclear fuel cycle. Using the source term 0.1 TBq/(GW a) for radon discharge from uranium milltailings and reducing it to 20 to 50% gives an estimate for the annual discharge of radon from coal ash
of 0.02–0.05 TBq/(GW a), agreeing well with the value 0.03 TBq/(GW a) obtained from the alternate
assessment above.
116. The amount of ash placed in disposal was assumed to be 60% of the total ash produced in coal
power plants, based on data cited earlier from the United States and the United Kingdom. The larger six
coal-producing countries in table 24 are also countries with large land mass and therefore the value of
60% from the United States was assumed appropriate for use in this assessment. The remaining 40% of
the ashes were assumed to be used in a variety of building-related activities and the resulting radiation
exposures are discussed in subsection (e).
117. It was explained in paragraph 43 that radon discharges from one gigawatt-year of coal ash
deposits were assumed to continue for 100 years and that allowance had to be made in the methodology
for this continued discharge. This was done by multiplying the results for one year’s discharge by 100
to allow for the discharge continuing for 100 years (see annex A, paragraph 17). Taking into account
this multiplication by 100 years and that 60% of the total ash produced in coal-fired power plants was
assumed placed in disposals, gave a value for the radon discharge from coal ash deposit normalized toelectricity generation of 1.8 TBq/(GW a).

(d) Results—public exposures due to discharges from coal mining, coal-fired
power plants and ash deposits118. The annual discharges to atmosphere obtained as described above were used to derive individualand collective doses using the revised methodology for estimating public exposures due to radioactive
discharges (annex A). The results of the calculation of the characteristic individual doses per unit ofelectricity generated integrated to 100 years due to atmospheric discharges from the three sources ofpublic exposures from the coal cycle are shown in table 28. These three sources are (a) discharges fromcoal-fired power plants, (b) discharges from coal ash disposed in landfills and (c) discharges from coalmining. The radon discharges from one gigawatt-year of coal ash deposits were assumed to continue for
100 years. The results are presented for both older coal-fired and for modern coal-fired power plants. If
the coal-fired power plants are older, then the individual doses from all three routes are similar, withthose from the discharges from power plants being the highest. For newer power plants, the doses due
to discharges from the plants are about 10% of those from the other two sources.

190 UNSCEAR 2016 REPORT119. The values for each region that are shown in table 28 for the characteristic individual doses perunit of electricity generation due to radon discharges from coal mining and from coal ash deposits
remain constant across the regions. This is because the discharges from mining and from ash deposits
are from radon gas and therefore only result in inhalation doses. However, the discharges from coalfired power plants include the other radionuclides from the 238U series, as shown in table 27, and resultin doses due to deposition and food consumption, which do have variations with geographical region.
As described above, the 232Th series was not considered in these calculations, and earlier assessments
by [Z1], imply that this contribution is less than 10% of the contribution from 238U.
120. The results also show a difference in the characteristic individual doses per unit of electricity
generated due to coal-fired power plant discharges between the older and the modern coal-fired plants
with the values for the modern plants being about 10% of those for the older plants. It could be
expected that the values resulting from discharges from ash would differ between the older and themodern coal-fired power plants, but this is not reflected in table 28. This is because the actual
difference in the amount that remains in the ash for an older versus a modern coal-fired power plant isvery small, of the order of about 1%, and for both cases most of the radionuclides stay in the ash. Thisdifference therefore does not show up in the results.
121. The collective doses to the public integrated to 100 years due to electricity generation from thecoal cycle for each region and for the world as a whole are shown for the total electricity generated
from coal in the year 2010 in table 29. Both cases, assuming (a) that all coal-fired plants are older and(b) that all coal-fired plants are modern, are compared. Considering the three discharge sources (frommining, power plants and ash deposits), the contribution to the collective dose from the discharges fromcoal-fired plants dominates for the older coal-fired plants, but is the smallest term for the modern plants
where the contribution from mining is largest. To facilitate comparison, table 30 compares only theterms from the older versus modern coal-fired power plants.
Table 28. Summary of characteristic individual doses per unit of electricity generated integrated to
100 years due to atmospheric discharges from coal-fired power plant sources (mSv/(GW a))
SourceaAfricaAsia and
Pacific

EuropeLatin
America

North
America

From mining

2.4 × 10−4

2.4 × 10−4

2.4 × 10−4

2.4 × 10−4

2.4 × 10−4

From older coal plants

3.8 × 10−4

4.0 × 10−4

4.2 × 10−4

3.9 × 10−4

4.2 × 10−4

From modern coal plants

3.1 × 10−5

3.2 × 10−5

3.4 × 10−5

3.2 × 10−5

3.4 × 10−5

From asha

1.5 × 10−4

1.5 × 10−4

1.5 × 10−4

1.5 × 10−4

1.5 × 10−4

The annual discharges of radon from coal ash produced per unit of electricity generated are assumed to continue for 100 years.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 191

Table 29. Summary of collective doses to the public integrated to 100 years for the total electricity
generated from combustion of coal (in 2010) (man Sv)
Discharge type and source

AfricaAsia and
Pacific

EuropeLatin
America

North
AmericaWorld

LOCAL COMPONENTa – ATMOSPHERIC DISCHARGES
Mining1.0 × 100

7.1 × 101

8.6 × 100

3.8 × 10−1

3.4 × 100

7.0 × 101

Older coal plants

1.8 × 100

1.3 × 102

1.9 × 101

7.1 × 10−1

7.4 × 100

1.4 × 102

Modern coal plants

1.5 × 10−1

1.1 × 101

1.5 × 100

5.6 × 10−2

5.8 × 10−1

1.1 × 101

Ashb

6.7 × 10−1

4.6 × 101

5.5 × 100

2.4 × 10−1

2.2 × 100

4.5 × 101

Total (assuming all older coal plants)

3.5 × 100

2.5 × 102

3.3 × 101

1.3 × 100

1.3 × 101

2.5 × 102

Total (assuming all modern plants)

1.9 × 100

1.3 × 102

1.6 × 101

6.8 × 10−1

6.2 × 100

1.3 × 102REGIONAL COMPONENTc – ATMOSPHERIC DISCHARGES
Mining

4.5 × 100

3.0 × 102

3.7 × 101

1.6 × 100

1.5 × 101

3.0 × 102

Older coal plants

8.7 × 100

6.4 × 102

9.0 × 101

3.4 × 100

3.5 × 101

6.4 × 102

Modern coal plants

6.9 × 10−1

5.0 × 101

6.9 × 100

2.6 × 10−1

2.7 × 100

5.0 × 101Ashb

2.9 × 100

2.0 × 102

2.4 × 101

1.0 × 100

9.4 × 100

1.9 × 102

Total (assuming all older coal plants)

1.6 × 101

1.1 × 103

1.5 × 102

6.0 × 100

5.9 × 101

1.1 × 103

Total (assuming all modern plants)

8.0 × 100

5.5 × 102

6.7 × 101

2.9 × 100

2.7 × 101

5.5 × 102TOTAL LOCAL AND REGIONAL COMPONENTS – ATMOSPHERIC DISCHARGES
Mining

5.5 × 100

3.7 × 102

4.5 × 101

2.0 × 100

1.8 × 101

3.7 × 102

Older coal plants

1.1 × 101

7.7 × 102

1.1 × 102

4.1 × 100

4.3 × 101

7.8 × 102

Modern coal plants

8.3 × 10−1

6.0 × 101

8.3 × 100

3.2 × 10-1

3.3 × 100

6.0 × 101Ashb

3.5 × 100

2.4 × 102

2.9 × 101

1.3 × 100

1.2 × 101

2.4 × 102

Total (assuming all older coal plants)

2.0 × 101

1.4 × 103

1.8 × 102

7.4 × 100

7.2 × 101

1.4 × 103

Total (assuming all modern plants)

9.9 × 100

6.8 × 102

8.3 × 101

3.6 × 100

3.3 × 101

6.7 × 102a

The local component of collective dose is calculated for people living between 0 and 100 km from the discharge point.

b

The discharges of radon from coal ash produced in 2010 from electricity generation are assumed to continue for 100 years.

c

The regional component of collective dose is calculated for people living between 100 and 1,500 km from the discharge point.

192 UNSCEAR 2016 REPORT

Table 30. Comparison of collective doses to the public integrated to 100 years from coal-fired
power plants given alternative plant design (man Sv)
Collective doses to the public integrated to 100 years (man Sv)
Doses from coal-fired power
plantsAfricaAsia and
Pacific

EuropeLatin
America

North
AmericaWorld

LOCAL COMPONENTa – ATMOSPHERIC DISCHARGES
Older coal plants

1.8 × 100

1.3 × 102

1.9 × 101

7.1 × 10−1

7.4 × 100

1.4 × 102

Modern coal plants

1.5 × 10−1

1.1 × 101

1.5 × 100

5.6 × 10−2

5.8 × 10−1

1.1 × 101

REGIONAL COMPONENTb – ATMOSPHERIC DISCHARGES
Older coal plants

8.7 × 100

6.4 × 102

9.0 × 101

3.4 × 100

3.5 × 101

6.4 × 102

Modern coal plants

6.9 × 10−1

5.0 × 101

6.9 × 100

2.6 × 10−1

2.7 × 100

5.0 × 101

TOTAL LOCAL AND REGIONAL COMPONENTS – ATMOSPHERIC DISCHARGES
Older coal plants

1.1 × 101

7.7 × 102

1.1 × 102

4.1 × 100

4.3 × 101

7.8 × 102

Modern coal plants

8.3 × 10−1

6.0 × 101

8.3 × 100

3.2 × 10−1

3.3 × 100

6.0 × 101a

The local component of collective dose is calculated for people living between 0 and 100 km from the discharge point.

b

The regional component of collective dose is calculated for people living between 100 and 1,500 km from the discharge point.

(e) Coal ash in building materials
122. Coal combustion products (CCPs) are used as ingredients in a variety of construction materials
including concrete, grouting, fill material, lightweight aggregate, road construction or maintenancematerials, and in soil stabilization [U1].
123. Fly ash and furnace bottom ash are the main CCPs and the extent to which they are utilized varies
by country. Approximately 40% of CCPs produced in the United States are utilized in construction
materials [A2], and the percentage utilization was around 45% in the European Union [E2] for fly ash
and bottom ash CCPs produced in 2010. China is the largest producer of electricity from combustion ofcoal and consequently the largest producer of coal ash. According to Tang et al. [T2], China produces
proportionately more coal ash because it has a very low coal washing rate (51%). In 2010, the fly ash
utilization rate in China was over 65% [T2]. In contrast, in 2009 less than 20% of CCPs produced inAustralia were utilized [H3].
124. As noted previously, the activity concentration of natural radionuclides in coal ash is increased
relative to coal by a factor of 5–10. However, the concentration is again diluted when the fly ash ismixed with other materials to form construction material. The presence of fly ash in building materials
may increase indoor exposures to gamma radiation from 226Ra, 232Th and 40K contained in the coal ash,and exposures due to inhalation of radon emanating from the building construction material such asconcrete [D3, F1, T4].
125. The relationship between coal ash in construction material and exposure is not straightforward.
The use of coal ash has been shown to either increase, decrease or have essentially no effect on
exposures relative to traditional construction materials [T4]. This reflects differences in the source ofthe coal, the natural radionuclide content of the traditional construction materials, and to the changes instructural properties of the construction material. For example, the use of fly ash and furnace slag inconcrete has been shown to reduce radon exhalation rates because it alters the porosity and chemistry of

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 193

the construction material [D3, D4]. As of the year 2016 there are no consistent or internationallyaccepted limits for radionuclides in building materials.
126. A population-weighted average annual effective dose of 0.41 mSv from indoor exposure to
gamma rays mainly determined by construction materials was reported by UNSCEAR [U9]. The resultsfor individual countries generally ranged between 0.3 and 0.6 mSv. A more recent analysis by de Jonget al. [D3] calculated an annual effective dose of 0.32 mSv for the Netherlands, which is at the lower
end of the range given by UNSCEAR [U9]. In both evaluations, the results included dwellings where
fly ash was used in the construction material.
127. Collective doses either for geographical region or for the world as a whole due to the use of coal
ash in building materials have not been analysed further in this study. Although the individual dosesthat have been assessed are very low, some contribution to the total collective dose can be expectedfrom the use of coal ash in construction materials. However, the various factors causing variability inthe exposure characteristics cause major uncertainty in the assessments.

3.

Occupational exposures from coal mining

128. Previous UNSCEAR studies have estimated the collective doses due to occupational exposure
from coal mining. In the UNSCEAR 1988 Report [U7], based on exposure data from British coalmining and the worldwide coal production rate, the worldwide collective dose was estimated as
2,000 man Sv. The UNSCEAR 2000 Report [U9] updated this value to an estimated 2,600 man Sv. The
UNSCEAR 2008 Report estimated the worldwide collective dose to workers from coal mining in
2002-2004 at 16,560 man Sv in a year [U11]. Most of the data on occupational exposure from coalmining that were used in the UNSCEAR 2008 Report came from China. A study on radiation levels in
China reported the annual average collective dose to underground coal miners in 2003–2004 as
16,500 man Sv [L1]. In addition, occupational exposures from coal mining in the United Kingdom in
2002 were reported in the UNSCEAR 2008 Report [U11] as 3 man Sv for 5,000 coal miners and an
average individual dose of 0.6 mSv.
129. The worldwide collective dose to coal miners in 2010 could be estimated from the 2002 collective
dose value assuming that the increase in electricity generation from coal in the period 2002–2010 was
proportional to the increase in occupational exposure due to coal mining for the same time span. This
assumed that the same collective dose normalized to electricity generation that was assessed for
2002-2004 [U11] was applicable for 2010 and that the efficiency of coal mining remained the same, i.e.
the same number of coal miners was needed to obtain the same amount of coal. The electricity
generated from coal-fired power plants in the reference year of 2010 was 992 GW a (table 2). The
corresponding number for the period of 2002–2004 considered in the UNSCEAR 2008 Report was
720 GW a [I12]. This gave a rounded worldwide estimate of the collective dose to coal miners of
23,000 man Sv, as a first approximation.
130. Table 31 shows results from a Chinese study [L1], together with data on the number of coalminers in three categories of mines in China for 2010 [W2]. No new values of the average annual
effective dose to coal miners have been published since the 2010 values in table 31.
131. Chinese coal production between 2004 and 2010 increased by approximately 50%. During thisperiod, there was an increase in the number of underground miners working at medium- and large-sizedcoal mines and a significant decrease in the number working in small-sized township and privatelyowned coal mines [W2]. The last column in table 31 uses the average annual effective dose from 2004
and the number of coal miners in 2010 to estimate the collective dose in 2010 in China for the three194 UNSCEAR 2016 REPORT

coal-mine categories: small, medium and large. This can be compared to the collective dose in
2002-2004, also shown in the table. (The number of bone-coal miners in China in 2010 was notavailable, although this group represented a small fraction of the total in the 2004 data.) This approachgave an estimated collective dose in 2010 for coal miners in China of about 10,000 man Sv.
132. During the period 2002–2004, China accounted for almost 90% of the underground coal miners in
the world. Assuming that the proportion of underground coal miners from countries worldwide was about
the same in 2010 as in 2002–2004, the total world collective dose to coal miners was estimated to be about
11,000 man Sv (compared to 23,000 man Sv obtained using the simpler approach above). This can also becompared to the value 16,560 man Sv that was reported in the UNSCEAR 2008 Report [U11]. The
implied decrease would be due to more efficient coal-mining technology for underground coal extraction
and to the closing of many of the smaller, less efficient coal mines with poor ventilation [W2].
133. The dose contribution for thorium and its decay products and aerosols containing long-lived
alpha-emitting radionuclides were not assessed in this study [L1]. In the past, the effective dose fromaerosols containing long-lived alpha-emitting radionuclides was roughly estimated by using thecontents of radioactive material in the coal and the concentrations of dust in underground coal mines.
The estimation indicated that the annual individual dose ranges from several to ten microsieverts, which
has been assumed to be of little significance in this comparative study.
Table 31. Annual doses to underground coal miners in China [L1]
The information in the table for number of miners for the year 2010 was obtained from an official communication from the
Chinese delegation to the Committee

Type of coalmine

Average annual
effective dose
(2004)(mSv)

Number ofminers in2004(millions)

Number ofminers in2010(millions)Collective
dose
2002–2004(man Sv)

Collective dose in 2010,
using average annual
effective dose from2004 and number ofminers in 2010 (man Sv)

Large-sized

0.32

1

1.26

315

403

Medium-sized

0.63

1

1.31

630

825

Small-sized

3.78

4

2.27

15 1008 581

Bone-coal

11.3

0.05

Not available

567

Not available

2.75 (weighted
average)

6.05

4.84

16 612

9 809

Total

B. Natural gas
134. According to table 2 the global combustion of natural gas produced 544 GW a of electrical energy
in 2010. This amount of electrical energy was second only to that due to the combustion of coal at
992 GW a and was 1.7 times the amount of electricity generated by nuclear power plants. The
generation of 1 kW h of electrical energy has been estimated to require the combustion of 0.286 m3 ofnatural gas [U20].
135. Natural gas is essentially methane (CH4) with trace amounts of other materials. The mostimportant radionuclide released during the combustion of natural gas is 222Rn. Several authors have

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 195

reported on the concentration of 222Rn in natural gas as measured at the wellhead or at various otherlocations in the gas-delivery system. In developed countries, natural gas is typically not used directlyfrom the wellhead, but is processed to remove moisture and refrigerated to condense and remove higher
chain gases, including ethane (C2H6), propane (C3H8) and butane (C4H10). The condensation process isimportant, because radon tends to condense with ethane and propane [D6, G2]. Gesell [G2] studied nine
gas-processing plants and found that, on average, the 222Rn content of gas ready to be sold for
combustion (the sales gas) was 34% of that at the wellhead (the input gas) with a range of 4 to 90%.
Further, natural gas may be transported by pipeline over long distances and is also typically stored at
locations near its end use. Because the half-life of 222Rn is 3.8 days, appreciable decay can occur
between the production and combustion of natural gas.
136. Measured values of 222Rn in natural gas are summarized in table 32. The range of individual
values was large; the weighted average of 879 samples was 625 Bq/m3. Most of the values in table 32
were for samples taken at the wellhead. Considering the processes that reduce the concentration ofradon from the gas at the wellhead to the gas sold for combustion [V2], a value of 300 Bq/m3 wasconsidered to be a reasonable estimate.
Table 32. Summary of reported concentrations of 222Rn in samples of natural gas
Unless noted otherwise, the samples were taken at or near the wellhead. Concentration values are in units of Bq/m3

SourceArea

Number ofsamples

Minimumvalue

Maximum
value

Mean

[B9]

NW New Mexico
SW Colorado

307

7.4

5 880

910

[M4]

NW New Mexico
SW Colorado

42

11.8

2 130

610

[G2]

Texas, Oklahoma, Louisiana

15

37

4 400

1 330

[V1]

Netherlands, Germany, North Sea,
Borneo, Nigeria

~200

33

1 650

74a

[K2]

N. Germany

196

—

4 000

580

[V2]

British Columbia

~32

7

921

272

[O9]

Ireland

8b

116

918

638

[A4] cSyrian Arab Republic

36

15

1 142

400

[R4]

Pennsylvania

21

37

2 920

1 370

[P2]

Pennsylvania

22

110

5 476

1 770

Weighted average879

aAverage results for the Netherlands only.

bSame field sampled at various times.

c

Taken at many locations, including those in gas-processing plants.

625

196 UNSCEAR 2016 REPORT137. The calculation of the activity of 222Rn released normalized to electricity generated is thus given as

0.2860

m3
Bq
h
kW
TBq
TBq
× 300 3 × 8760 × 10 6
× 10 −12= 0.75kW ⋅ h
a
GW
Bq
GW
⋅a
m

(2)

This value for the release of 222Rn normalized to electricity generated was used with UNSCEAR’s
revised methodology for estimating public exposures due to radioactive discharges (annex A). The
2010 data on electricity generation from natural gas for each region, shown in table 2, were used, andthe default population distribution based on population densities for 2010 was used for the collective
doses (see paragraph 34). The resulting characteristic individual and collective doses are compared withthose for other electricity-generating technologies in table 33 (for characteristic individual doses) at theend of this chapter and in table 46 (for collective doses) in chapter VIII.
138. Other radionuclides can appear in natural gas. One of them is 220Rn, but its half-life (55.6 s) is too
short for it to appear at the point of end use of the gas for combustion. If water is co-produced withnatural gas, the water may contain soluble amounts of radium. Radium-226 is the parent of 222Rn and isa decay product of 238U; 224Ra is the parent of 220Rn and is a decay product of 228Ra and ultimately of232
Th (see figure III). The progeny of the three radium isotopes and 222Rn can also be present in naturalgas, but typically in small amounts. Lead-210 is a longer-lived (half-life of 22.2 years) decay product of226
Ra and 222Rn, and van der Heijde et al. noted that the activities of 210Pb and its decay product 210Po innatural gas are more than can be accounted for by the decay of 226Ra and 222Rn [V1]. These authors
concluded that the additional 210Pb and 210Po had been produced from the natural gas reservoir. In
general, these radionuclides other than radon do not reach the end point of combustion of natural gas,
but they typically plate out within the gas-distribution system fairly close to the point of withdrawal
and/or treatment. Because the radionuclides are contained within equipment, workers can incur external
exposure to radiation. Organo and Fenton [O9] concluded after investigations in Ireland that workers at
offshore locations would be exposed “at most” to around 100 μSv in a year. The Pennsylvania
Department of Environmental Protection [P2] measured the external ambient dose rate at several
locations within gas production and processing sites and estimated a maximum annual average dose of
270 μSv. The Committee has assumed a reasonably realistic estimate of 100 μSv in a year.
139. A problem in estimating the collective effective dose to workers is determining the number ofworkers in the industry apportioned to the amount of natural gas used to generate electricity. An
indirect method was used to make this determination. According to the United States Bureau of Labor
Statistics [U16], there were 89,000 production and non-supervisory workers in the oil- and gasextraction subsector. Most of these workers were not working directly in support of natural gas
production for electricity generation, but there was no further breakdown in the data. In order toestimate the fraction of workers supporting the generation of electrical energy by combustion of naturalgas, data from [I17] for the OECD countries for 2010 were used. In terms of energy produced, thesedata indicated that 12% of the energy produced from oil and natural gas together was used forelectricity generation due to combustion of natural gas. Thus, it was estimated that 10,700 workers were
employed in the United States to produce natural gas for the generation of electricity.
140. According to [U19] 209.2 billion cubic metres of natural gas were used in 2010 to produce
electricity in the United States. With the use of previously mentioned conversion factors this was
equivalent to 83.5 GW a of electrical energy. Thus, the normalized collective dose for workers based on
the data originating in the United States was 0.013 man Sv/(GW a).

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 197

C. Oil141. The combustion of oil for electricity generation accounted for less than 5% of the world’s
electricity generation in 2010, but accounted for more than 30% of electrical energy generated in West
Asia. The generation of 1 kW h of electrical energy is estimated to require the combustion of 0.278 L of
petroleum [U20]. Most of the “petroleum” combusted to generate electricity in 2010 in the OECD
countries consisted of fuel oil [I17], which has an approximate density of 0.95 kg/L.
142. Petroleum is produced from underground reservoirs in the presence of geological formations that
contain the primordial radionuclides 40K, 238U and 232Th; the latter two are parents of chains that
produce a series of radionuclides usually present in secular equilibrium (see figure III). A process thatcan perturb the secular equilibrium is the dissolution of some members of the chain in water, which at
depth can be at high temperature. Water containing dissolved radium is typically co-produced with oil.
Radium-226 is a member of the chain headed by 238U; 226Ra is the parent of 222Rn and an additional
series of decay products including 210Pb and 210Po. Radium-228 is a member of the chain headed by232
Th; 228Ra is the parent of 224Ra and 220Rn. Both 226Ra and 224Ra are parents of additional short-lived
beta or gamma-emitting radionuclides.
143. Apparently some power plants can burn crude (or heavy) oil with its associated natural gas [A5], but
it is more typical that associated natural gas and water produced in the process must be removed frompetroleum before delivery to a pipeline or refinery [S5]. Radon is more typically associated with natural
gas and the soluble isotopes of radium are typically associated with the produced water. As produced
water is withdrawn from depth, it cools and dissolved minerals, including radium, can form precipitatesand deposit on the production tubing and on various other production equipment. These deposits are
known as scale or sludge. With time the radium contained within scale and sludge will come into
equilibrium with gamma-emitting decay products and give rise to external exposure of workers.
144. There are many studies of the presence of radionuclides in scale, sludge, (e.g. [I1, R3]) andproduced water [S5], and of radon in natural gas (see section V.B above), but there are few
measurements of radium or radon in crude oil or its subsequent products. Crude oil generally contains
some remaining water within a water-in-oil emulsion and this water contains dissolved radium [S5].
Radon is also quite soluble in oil [B2].
145. The report by Bell et al. [B2] appears to be the only one that reports measurements of both 222Rn
and 226Ra in the same samples of crude oil. Because the half-life of 222Rn is 3.82 days, the samples had
to be delivered quickly and correction made for decay in transit. Seven samples of crude oil from wells
in Texas and Oklahoma were measured. The results for 222Rn varied from a minimum of 3.2 Bq/kg to a
maximum of 17 Bq/kg and had an average of 7 Bq/kg. The results for 226Ra had a minimum of
0.2 Bq/kg, a maximum of 13 Bq/kg and an average of 0.7 Bq/kg. The ratio of 222Rn to 226Ra had aminimum of 4, a maximum of 38, and an average of 15. Thus, most of the 222Rn in fresh samples ofcrude oil was unsupported, that is, it was far in excess of the activity that was in equilibrium with 226Ra.
Except for those power plants that burn crude oil directly, this excess 222Rn would decay or be removed
in the refining and/or transportation processes.
146. Hamlat et al. [H1] reported concentrations of 226Ra in an unspecified number of samples of crude
oil collected in Algeria. The range of values was from 6 Bq/kg to 20 Bq/kg; these values are reasonably
consistent with those reported by Bell et al. [B2]. Al-Saleh and Al-Harshan [A7] measured
radionuclides in 14 petroleum-product samples, including crude oil drawn from the Riyadh City
Refinery. The detection limit for 226Ra was 0.014 Bq/kg; the only products with detectable amounts of226
Ra were “sweet naphtha” (0.65±0.40 Bq/kg) and “flushing oil” (0.45±0.20 Bq/kg).

198 UNSCEAR 2016 REPORT147. The limited amount of data indicated above suggests a cautious approach in estimating the releaseof radionuclides from the combustion of oil products. For this procedure the Committee estimated that
there was a concentration of 226Ra in all oil products of 1 Bq/kg. Further, it assumed that unsupported
amounts of 222Rn decayed or otherwise were eliminated in the process, so that the concentration of
222
Rn in all oil products was also 1 Bq/kg.
148. Then, the normalized release of 222Rn was calculated as

0.278

L
kg Bq
h
kW
TBq
TBq
× 0.95 × 1
× 8760 × 10 6
× 10 −12= 0.002kW ⋅ hL
kga
GW
Bq
GW ⋅ a(3)This value for the release of 222Rn normalized to electricity generated was used with the revised
methodology for estimating public exposures due to radioactive discharges (annex A). The 2010 dataon electricity generation from oil for each region, shown in table 2, were used, and the default
population distribution based on population densities for 2010 was used for the collective doses (seeparagraph 34). The resulting characteristic individual and collective doses are compared with those forother electricity-generating technologies in table 33 (for characteristic individual doses) at the end of
this chapter and in table 46 (for collective doses) in chapter VIII.
149. There may be other radionuclides emitted by the combustion of oil products; Al-Masri and
Haddad [A5] reported values of about 60 Bq/kg of 210Po in heavy oil fuel used in three power plants in
the Syrian Arab Republic. The same authors [A6] also reported that soil in the vicinity of the three
power plants contained enhanced levels of 210Pb and 210Po. Other authors have not reported
measurements of 210Pb and 210Po associated with oil-fired power plants.
150. As mentioned above, 226Ra and 228Ra and their decay products accumulate in scale and sludge
around wellheads and processing equipment. This gives rise to occupational exposure, when workers
are in the vicinity of these elements. Kvasnicka [K4] estimated that the maximum effective dose on
offshore platforms could be managed to be below 1 mSv in a year. It is expected that the doses toworkers on offshore platforms would be higher than on onshore platforms, because of more substantial
problems with scale formation and the limited space in the working environment on the platforms.
Hamlat et al. [H1] measured ambient dose rates in the vicinity of onshore oil and natural gas production
equipment. They estimated that annual effective doses around oil extraction equipment might range
from 40 to 600 μSv for normal activities. A value of 300 μSv was adopted for this analysis.
151. There were no direct data on the number of workers in the oil industry apportioned to the
combustion of oil for the generation of electricity, so an indirect method was used. According to theUnited States Bureau of Labor Statistics [U16], there were 89,000 production and non-supervisory
workers in the oil- and gas-extraction subsector in the United States in 2010. In order to estimate thefraction of workers who were supporting the generation of electrical energy by the combustion of oil,data from [I17] were used. In terms of energy produced these data indicated that only 1.7% of energy
produced from oil and natural gas was used for the generation of electricity by the combustion of oil in
2010. Thus, it was estimated that 1,520 workers in the United States could be apportioned to generationof electrical energy by the combustion of oil.
152. According to [U19] 7.02 × 109 kg of oil were used in 2010 to generate electricity in the UnitedStates. Using the conversion factors given above, this is equivalent to 3.03 GW a of electricity. Thus,
the collective dose to workers normalized to the electricity generated, based on data from the UnitedStates, was estimated to be 0.15 man Sv/(GW a).

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 199

D. Comparisons of exposures from fossil-fuel electricity generation
153. Table 33 shows characteristic individual doses integrated to 100 years per unit of electricity
generated for the UNEP regions. These doses are due to atmospheric discharges from the coal cycle andthe combustion of gas or oil in power plants generating electricity. Although the doses resulting fromthe modern coal plants are about 10% of the doses from the older plants, in total the characteristic
individual doses per unit of electricity generated from both the older and the modern coal plants arelarger than from the gas or the oil combustion technologies. See chapter VIII for further comparisons
between these and other electricity-generation technologies.
Table 33. Summary of characteristic individual doses per unit of electricity generated integrated to
100 years due to atmospheric discharges from the coal cycle and the combustion of gas and oil(mSv/(GW a))Showing results from both older and modern coal plants

Electricity
generation
technology

ReleaseAfricaAsia and
Pacific

EuropeLatin
America

North
AmericaWest Asia- From mining

Radon

2.4 × 10−4

2.4 × 10−4

2.4 × 10−4

2.4 × 10−4

2.4 × 10−40

- From older
coal plants

Naturalradionuclides
includingradon

3.8 × 10−4

4.0 × 10−4

4.2 × 10−4

3.9 × 10−4

4.2 × 10−40

3.1 × 10−5

3.2 × 10−5

3.4 × 10−5

3.2 × 10−5

3.4 × 10−50

- From ash

Radon

1.5 × 10−4

1.5 × 10−4

1.5 × 10−4

1.5 × 10−4

1.5 × 10−4

0

Gas

Radon

6.3 × 10−5

6.3 × 10−5

6.3 × 10−5

6.3 × 10−5

6.3 × 10−5

6.3 × 10−5

Oil

Radon

1.7 × 10−7

1.7 × 10−7

1.7 × 10−7

1.7 × 10−7

1.7 × 10−7

1.7 × 10−7

Coal

- Frommoderncoal plants

VI. RADIATION EXPOSURES ARISING FROM ELECTRICITY
GENERATION FROM GEOTHERMAL ENERGY
154. Geothermal energy is derived from the heat of the earth at locations where magma is closer thannormal to the earth’s surface. The generation of electricity from geothermal sources presently depends
upon the use of steam or hot water from wells drilled into underground reservoirs. As shown in table 2[I17], geothermal energy in the reference year 2010 generated 7.77 GW a of electricity and accounted
for 0.3% of the total world’s electricity generation in 2010.
155. Electricity generation from geothermal sources first began in 1904 at Larderello, Italy, with anexperimental 10 kW generator. At the present time the installed capacity is in excess of 9,000 MW in
25 countries [L3]. The larger power plants are at The Geysers, California; Larderello, Italy; Cerro
Prieto, Mexico; Leyte, Philippines; Salton Sea, California; Hellisheidi, Iceland; Tiwi and Malitbog,
Philippines; and Wayang Windu and Darajat, Indonesia [K1].

200 UNSCEAR 2016 REPORT156. Different approaches can be taken to generate electricity from geothermal sources. If dry steam isavailable, the simplest is to run the steam through a turbine and exhaust to the atmosphere. Efficiency isimproved if the dry steam is condensed, with heat dissipated through cooling towers. Most reservoirs,
however, do not produce dry steam, but some combination of steam and hot water or hot water alone.
In this situation, it is necessary to separate steam from water, or, if hot water alone is produced, to flash
the water to steam through single, double, or even triple passes. A binary process is in use at some sitesthat use water of lower temperature; hot water is run through a primary loop and a fluid with a lower
boiling point through a secondary loop. As of the year 2016, most electrical energy generation isthrough the flash process with single or double flash; the other major technology is the use of dry steamwith condensation [M3].
157. During the 1970s there was substantial interest in evaluating the potential impact of the release of222
Rn from gases vented during the operation of geothermal electricity-generating plants. The more
complete evaluations were reported for The Geysers and the Larderello Plants. Anspaugh [A10]reported that 19.3 TBq/a were being released from The Geysers during the operation of 11 dry steam
condensing units with 502 MW of electrical power; the daily rate was noted to be equivalent to theaverage daily background emission of 222Rn from about 40 km2 of soil. For the Larderello geothermal
complex, the emission rate per unit energy was 38.5 TBq/(GW a) according to George et al. [G1]. Thiswas based on measurements performed by D’Amore [D1], who calculated that 300 GBq/d were being
released from the Larderello geothermal complex, which had an installed capacity of about 420 MW[P1]. Thus, the emission rate per unit energy at Larderello was 260 TBq/(GW a).
158. The above represent the two situations where data were complete enough to estimate emissions
per unit of electricity generated. Based on this limited set of data, the Committee assumed an averagevalue of 150 TBq/(GW a). There are measurements of radon in steam or hot water from other locations,and samples of such data are given in table 34; at the time of the assessment no data were available forthe amount of 222Rn that might be released from geothermal sites using binary circle systems. A reason
that data are scarce on radon related to geothermal energy is that, following the measurements in the
1970s, it was generally concluded that this source of radon was not significant compared to the natural
exhalation of radon from soil. Measurements of 222Rn and 220Rn in geothermal fluids do continue, but
are mainly related to understanding of the dynamics of resources [W6].
Table 34. Reported concentrations of 222Rn in fluids produced from geothermal wells
Location

Activity of 222Rn per unit ofproduced fluid (Bq/kg)

Reference

The Geysers, weighted average

520

[A10]

The Geysers

620

[K3]

Larderello

1 280

[K3]

Larderello, weighted average

3 100

[G1]

Salton Sea, California

110

[K3]

East Mesa, Imperial Valley, CA

1.1

[K3]

Wairakei, New Zealand

630

[W6]

159. The value of 150 TBq/(GW a) for the release of 222Rn normalized to electricity generation fromgeothermal energy was used with the electricity generation data by region from table 2 and withestimates of population density to give an estimate of the characteristic individual and collective doses.
There is substantial uncertainty regarding the population density around geothermal power plants, as itis necessary to locate the plants within the narrow confines of the resource. The Geysers and the Cerro

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 201

Prieto plants are in isolated locations, but large metropolitan areas are within 1,500 km. The otherextremes are geothermal power plants located on islands. Because of these discrepancies and the lack of
specific population data on each power plant, the Committee made two calculations for geothermal: one
with the assumption of low-density population, as in the assessments for uranium mining and milling,
and one with the default population density.
160. Table 46 in chapter VIII compares the collective doses to the public for the discharges from thenuclear fuel cycle, coal cycle, gas, oil and geothermal electricity-generating technologies. As explainedin the previous paragraph, two values are given for geothermal energy. The normalized discharge ofradon for geothermal electricity generation is the largest of the electricity-generating technologies
shown in table 46. However, the impact of geothermal electricity generation remains small because of
the limited use of this technology. Further, there is substantial uncertainty regarding the discharge ofradon from geothermal power plants; data are available for only two plants, and the estimateddischarges from these two plants vary substantially.
161. There are even fewer data on occupational exposure in the geothermal industry. Anspaugh [A10]reported no evidence of increased external gamma exposure, but did note increased levels of radon andshort-lived progeny at several locations normally inaccessible except during maintenance. In general,
moderate to severe disequilibria were found, and an average excess amount of exposure was
0.05 Working Levels 7 (WL) (1 µJ m−3). Razzano and Cei [R1] indicated that 36 persons work overthree shifts to operate the Larderello complex. From data in the same paper, the electricity generated in2013 at Larderello was 5,659 GW h with 767 MW of “efficient” installed capacity. The combination ofthese data leads to an annual occupational exposure of 5 man WLM/(GW a). According to [I11]
1 WLM is equivalent to an effective dose of 10–20 mSv, but the higher values are associated withmines and homes. A more reasonable value for outdoor exposure is 10 mSv, which would also be
consistent with [U10]. Thus, the normalized collective dose to workers in geothermal electricity
generation would be 0.05 man Sv/(GW a).

VII. ASSESSING DOSES FROM THE CONSTRUCTION PHASE OF
ELECTRICITY-GENERATING TECHNOLOGIES
162. All electricity-generating technologies considered by the Committee have a construction phaseduring which the facilities and infrastructure are established. Although in general there are no
significant radiation exposures associated with the construction, there may be radiation exposuresassociated with obtaining the materials for construction such as in mining activities.
163. Industries not directly linked to electricity generation, such as mining, milling and processing of
metal ores, contribute to occupational and public exposure because of the presence of naturalradionuclides [U11, U22]. While public exposures from the mining industry may be negligible on an
individual basis, typical occupational doses in the mining industry may be up to a few millisieverts in ayear. The collective dose from occupational exposures during the mining of metals was evaluated forthis study on the basis of the dose per unit mass of metal ore mined for each electricity-generating
technology, as discussed in the UNSCEAR 2008 Report [U11], and on recent data from 2012 onradiation exposure of miners of rare earth metals in China [W3].
7
A working level (WL) is a unit of potential alpha energy per unit of air. In the SI system of units, one WL is equal to 2.08 × 10−5 J m−3.
A working level month is a unit of exposure to 1 WL for a month. For occupational personnel one month is considered to be 170 hours.

202 UNSCEAR 2016 REPORT164. Aggregate is also used extensively in the construction of electricity-generating plants. Aggregate
in building and construction refers to the material used for mixing with cement, bitumen, lime, gypsum
and other additives to form concrete or mortar. Commonly used aggregates include sand, crushed or
broken stone, gravel, and blast-furnace slag. In some cases, the amount of aggregate used is similar to
the total mass of all other materials.
165. The relative volumes of concrete required to construct various nuclear and coal-fired power plants
are tabulated in tables 35 and 36. Typical radionuclide concentrations in raw and processed materials
and in wastes of the mineral processing industry, including the cement industry, were reviewed in the
UNSCEAR 2000 Report (see table 27 in [U9]). Typical radionuclide concentrations in metal ores werereported to be about three orders of magnitude higher than of materials in the cement industry. For thisassessment, only the occupational exposures from the mining of metal ores needed for metals used inconstruction of electricity-generating technologies were evaluated.
Table 35. Estimates of the amounts of construction materials used in nuclear power plants
Plant

a

Mass of metals (t)Volume of concrete (m3)

PWR, 1 000 MW(e) [D5]61 000169 000

BWR, 1 000 MW(e) [D5]66 000200 000

PWR, 1 000 MW(e) [B8]a38 000 (estimated)

75 000 (estimated)

Nuclear power plant [W5]36 00075 000

The amount presented for specific metals: Al, Cr, Cu, Ag, Fe, Pb, Mg, Mn, Mo, Ni, Sn, Zn.

Table 36. Estimates of the amounts of construction materials used in coal-fired power plantsaPlantaMass ofmetals (t)Comments

Volume ofconcrete (m3)Coal-fired power plant [S7]50 000 steel

Amount presented in generic terms as "steel"and expressed in kg/MW(e)380 000

Coal-fired power plant [W5]41 000 metals

Amount presented for specific metals: Al, Cr,
Cu, Mn, Mo, Ni, steel (low alloy and stainless)and V and expressed in t/GW(e)31 0001,000 MW(e) plant assumed.

166. The mass of metal ore required to produce the metals in the building materials has been estimatedfrom information on the total metal inventory required and from the metal ore grade used. It was
assumed that steel was the primary metal used. Iron is alloyed with carbon to produce plain carbon
steel. Alloy steels also contain various alloying metals. The category of steel used in the different
electricity-generating technologies and the contents of alloying metals were considered. Plain carbon
steels were assumed to contain iron and 0.16–0.59% carbon. Low-alloy steels were assumed to containless than 0.25% carbon, as well as small amounts of nickel, chromium, molybdenum, manganese andsilicon. High-alloy stainless steels were assumed to contain molybdenum, chromium and/or nickel andother elements. These assumptions are based on the information in [C1, G3].
167. For high-alloy stainless steel, type S316 was used as the basis for this assessment, because it is
widely used in industrial applications. Type S316 contains around 2% by mass of molybdenum [A3,S3]. Other steel types may contain different levels of molybdenum, or none at all, which could
influence the results. This is because molybdenum is a low-grade metal ore and therefore extraction ofmolybdenum may make a larger contribution to occupational exposures [A3, S3].

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 203

168. Based on the above, the composition of typical steels for each category (plain, low-alloyed andhigh-alloyed stainless) was assumed as shown in table 37. If the type of steel was not known, theproportions given in table 38 were assumed.
Table 37. Composition of different types of steel (based on [C1, G3])
Material

Composition (% by weight)
PLAIN STEEL

Iron

99.62

Carbon0.38

LOW-ALLOY STEEL
Iron

99.63

Molybdenum

0.33

Carbon0.05

HIGH-ALLOY STAINLESS STEEL
Iron

65.35

Chromium

17.00

Nickel

12.00

Molybdenum

2.50

Manganese

2.00

Other non-metallic elements

1.07

Carbon0.08

Table 38. Assumed mix of steel types used in construction where details were not specified(based on [C1, G3])
Steel type

Proportion of total (%)

High-alloy stainless steel

10

Low-alloy steel

30

Plain steel

60

169. Assumptions on ores. The amount of ore required for the construction of an electricity-generating
facility also depends on the ore grade. The assumptions made regarding ore grades are shown intable 39 [A8, C3, M1].

204 UNSCEAR 2016 REPORT

Table 39. Metals and their assumed ore grades
The data presented here are typical values taken from [A8, C3, M1], and the numbers have been rounded

Resource

Ore grade at mine (%)

Tin

79

Manganese

47

Zirconium

47

Iron

46

Magnesium

41

Tantalum

31

Chromium26

Aluminium

11

Zinc

7

Lead

6

Nickel2

Copper

1.4

Vanadium

0.3

Molybdenum

0.18

Silver0.01

Gold0.0005170. Recycling of metals. Recycling is an important part of the metal industry from both an economicand energy-saving point of view. In the United States, approximately two thirds of the steel produced in2008 was made from recycled material [U15]. Around 43% of the total crude steel production
worldwide is made from recycled steel, with the main sources for steel recycling typically being
discarded cars, household appliances and steel cans, as well as old buildings and structures [W4].
Typical building construction uses approximately 60–65% recycled metal [B1], while in the United
Kingdom about 94% of construction steel is recovered [T3].
171. Special rules apply to the re-use of scrap steel and other metals from nuclear power plants;
authorities establish limits on the activity concentrations of radionuclides in materials that can bereleased for recycling. There is, however, no indication that there are significant differences in the
degree of recycling between the various electricity-generating technologies.
172. The effect of recycling metals has not been included in the present study, because of insufficient
data for large portions of the world. By not explicitly considering recycling, the collective dose per unitof electricity generated due to the extraction and processing of the metals used in the construction of a
facility is likely to be overestimated by around a factor of two on average, and possibly a factor of fiveor more for some regions of the world with developed recycling processes. This should not affect therelative comparisons of occupational exposures from mining metals between technologies, but would
affect comparisons of the total exposures (public and occupational) among the electricity-generating
technologies.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 205

173. Dose assessment. To estimate the radiation exposure of workers in the mining industry(occupational exposure) from ore extraction and processing, the following assumptions were made:(a) The ores are extracted only through underground mining. The use of above-ground mining
would give rise to smaller radiation doses than underground mining [U11]. Consequently, thisassumption leads to an overestimation of collective dose.
(b) The presence of several metals in an ore was not considered. Each ore was assumed to contain
only one metal type, and thus the contribution to occupational doses from the extraction andprocessing of the various minerals was evaluated independently and summed. In reality somemetals will be extracted together and consequently the resulting collective doses may have been
overestimated.
(c) Occupational exposures from mining and processing of raw materials were inferred from dataon uranium mining and processing, and on mining other minerals. Mining data for undergroundcopper mines in Australia (Mount Isa), Canada (Kidd Creek), South Africa (Palabora), andPortugal (Neves Corvo) show annual production rates of 1–5 kilotonnes of ore per employee [I19].
The denoted employees included all workers at the mines. Based on these data, the Committee
selected a mid-range value of 3 kt of metal ore per employee for use in this assessment (table 40).
(d) The average annual effective dose to non-uranium miners, including those involved in mineralprocessing, given in the UNSCEAR 2008 Report [U11] was 3.0 mSv, with a range of 1.3 to
5.0 mSv. Using an annual average value of 3.0 mSv for all workers associated with the mining
industry, and the value of 3 kt of metal ore per employee given in (c), the collective dose per unit
mass of ore extracted was estimated at 1 × 10−9 man Sv/kg (table 40). The data for (c) and (d) wereobtained from different worker populations and this introduces some uncertainty in the estimationof the collective dose per unit mass of ore extracted.
(e) In a recent study in China [W3] data were collected on radiation levels, and dose assessments
were performed for workers of the rare earth industry in the Baotou area and Sichuan Province,
two large mining districts in China. These regions account for about 77% of the total annual outputof rare earth mines in China; the reported annual production rates were 0.01 and 0.02 kt of rareearth metal per mining employee, respectively.
(f) The Chinese study has estimated the collective dose per unit mass of rare earth metal
extracted. The typical value for the whole rare earth industry that was obtained in the Chinese
study on radiation exposure of Chinese workers in mining, crushing, beneficiation and refining isabout 150 × 10−9 man Sv/kg. This value relates to the mass of metal extracted, which differs fromthe value 1 × 10−9 man Sv/kg given in (d) which relates to mass of metal ore. The Committee used
the latter value in the analysis in order to account for the different metal requirements withassociated metal ore grades of each electricity-generating technology.
(g) As a check, an alternative estimate was obtained by comparing with the data on uranium
mining. The collective dose per unit mass of uranium extracted given in the UNSCEAR 2008
Report [U11] was 1 man Sv/kt (based on 2000–2002 data). Assuming (from information in [I4,O8]) an average grade for uranium ore of 0.3%, the collective dose per unit mass of uranium ore
extracted is 3 × 10-9 man Sv/kg.
174. Although the assumptions in paragraph 173 affect the absolute values of the assessed collective
dose, the relative importance of the different electricity-generating technologies with regards to theconstruction phase remains.

206 UNSCEAR 2016 REPORT

Table 40. Collective effective dose per unit mass of ore mined used in this study
Estimatebased onAnnual production rateof ore per person (kt)

Average annual
occupational dose (mSv)

Collective dose per unit mass ofmetal ore extracted (man Sv/kg)

[U11]

3

3

1 × 10−9175. Nuclear power plants. For the current assessment, the assumed design lifetime of a nuclear power
plant was 40 years and the capacity factor 8 was 90% (table 41).
176. Three different studies of the use of materials in nuclear power plants have been published. Doneset al. [D5] studied the material used in two Swiss nuclear power plants—the pressurized-water reactor(PWR) at Gösgen and the boiling-water reactor (BWR) at Leibstadt. Both were of the 1,000 MW(e)type, but the capacity of the Leibstadt plant was increased to 1,190 MW(e) following construction. For
the latter, the total amount of metals used was 66,000 tonnes, of which two thirds were unalloyedsteel/iron. In addition, 200,000 m3 of concrete were used. Bryan and Dudley [B8] considered thequantities of different materials used in a typical 1,000 MW(e) PWR with river water cooling. The total
mass of metals used was estimated at 38,000 tonnes but the proportions of each metal were notspecified. The total amount of concrete used was estimated as 210,000 tonnes (~87,260 m3). White andKulcinski [W5] gave the amount of material needed for four power plants, including a nuclear power
plant, where values of 36,000 tonnes of metals and 180,000 tonnes (~75,000 m3) concrete were given.
The study focused on the energy payback ratio 9 and CO2 emissions, and it was not in that sense a fulllife cycle assessment. Furthermore, the study did not consider potential radiation exposure. The datafrom the three studies are presented in table 35.
Table 41. Assumed plant sizes, life times and capacity factors for the different electricitygenerating technologies*

Assumed to be the same as coal

Electricity-generating
technology

Typical plant gross capacity(MW(e))

Lifetime(years)Capacity factor(%)

Nuclear [I3]

1 000

40

90

Coal [F3, I18]

1 000

40

80

Natural gas [O1, S8]

505

25

80

Solar PV [F3, N3]

—

30

9.4

Wind turbine [D2, N4, O10]

25

20

24

1 000*

20

75

NA

30

NA

Biomass [W5]Geothermal [R5, S9]8
Capacity factor (net) is the ratio of the net electricity generated, for the time considered, to the energy that could have been
generated at continuous full-power operation during the same period [U25].
9Energy payback ratio is the ratio of total electrical energy produced during a system’s normal lifespan, divided by the electricalenergy required to build, maintain and fuel it.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 207

177. Example of a dose calculation for a nuclear power plant. Specifications of the typical plant grosscapacity, lifetimes and capacity factors (can also be called load factor) used in the calculations areshown in table 41. Table 42 shows an example of the details of a dose calculation for a nuclear power
plant. The collective dose was estimated based on occupational exposures related to metal ore mining
and processing of the metal ore for metals needed to construct the power plants.
Table 42. Example calculation of collective dose from occupational exposure during construction
phase of a nuclear power plant
Type: PWR Gösgen, 970 MW(e), capacity factor 90% [D5]Resource/Elements

Metal needed forconstruction (kg/kWh)

Ore grade atmine (%)

Amount of ore neededat mine (kg/kWh)

Collective dose
(manSv/(GW a))

Iron

4.7 × 10−546

1.0 × 10−4

8.9 × 10−4

Manganese

1.4 × 10−647

3.1 × 10−6

2.7 × 10−5

Chromium

1.2 × 10−526

4.8 × 10−5

4.2 × 10−4

Molybdenum

1.8 × 10−6

0.18

9.9 × 10−4

8.7 × 10−3

Nickel8.6 × 10−62

4.3 × 10−4

3.8 × 10−3

Iron

1.8 × 10−546

3.9 × 10−5

3.5 × 10−4

Molybdenum

5.9 × 10−8

0.18

3.3 × 10−5

2.9 × 10−4

Iron

1.1 × 10−446

2.4 × 10−4

2.1 × 10−3

Copper

4.7 × 10−6

1.4

3.3 × 10−4

2.9 × 10−3

Aluminium

6.3 × 10−7

11

5.6 × 10-6

4.9 × 10−5

Stainless steel

Low-alloyed

Unalloyed steel

Other

Total

1.9 × 10−2

178. Coal-fired power plants. There are a number of different designs of coal-fired power plants.
However, the amounts of material used in their construction do not differ significantly. Spath et al. [S7]conducted a full life cycle assessment of a coal-fired power plant covering the construction anddecommissioning of the plant. The amount of steel used in the construction of a 1,000 MW(e) plant was
50,000 tonnes, but no data were given on the other metals. Also, the proportions of plain carbon steel,low-alloy steel and stainless steel were not specified. White and Kulcinski [W5] described a total
amount of 41,000 tonnes of metals in their study. In both studies [S7, W5], the analysed units had anassumed capacity factor of 80% and a lifetime of 40 years. The information from these studies issummarized in table 36.
179. Natural gas power plant. The designs of natural gas power plants also vary. Spath and Mann [S8]presented a full life cycle assessment for a 505 MW(e) natural gas power plant covering its construction
and decommissioning. The study is similar in approach to the one they used for coal-fired power plants[S7]. Again, potential radiation exposure was not considered and no specific data were provided on theparticular metals used to construct the plant. The amount of steel used was 31 t/MW(e) and the amountof concrete used was 98 t/MW(e). The Committee assumed that the average capacity of a natural gas

208 UNSCEAR 2016 REPORT

power plant was 505 MW(e) to estimate the amount of each metal using the assumptions shown intable 37 and table 38.
180. Solar energy power plant. Solar energy is by far the largest energy resource available on earth.
Two different technologies contribute to solar electricity generation: solar photovoltaics (PV) andconcentrating solar power (CSP).
181. Solar PV systems convert direct and diffused solar radiation into electricity through a
photovoltaic process using semiconductor devices. Solar PV systems can be used anywhere in theworld on suitable land and on buildings. Solar PV technology is also very adaptable to being used in amodular fashion, which means that systems can be installed close to centres of demand.
182. Silicon-based PV solar cells contain small amounts of rare earth elements. Crystalline-silicon PV
cells are the most common PV cells in use today [U14].
183. Concentrating solar power (CSP) systems are designed to produce high-temperature heat forelectricity generation or for co-generation of electricity and heat. CSP systems are capable only of
exploiting direct normal irradiation, which is the energy received directly from the sun (i.e. notscattered by the atmosphere) on a surface tracked perpendicular to the sun’s rays. Areas suitable for
CSP development are those with strong sunshine and clear skies, usually arid or semi-arid areas.
Parabolic mirrors or troughs are used in CSP 10 technology. The parabolic mirrors are designed toconcentrate solar radiation onto linear heat collection elements [N2].
184. A solar energy power plant does not have a fuel cycle in the same way that most other energy
technologies do, consequently any potential human exposure to ionizing radiation is related to the useof natural resources for manufacturing the centralized CSP plants or the PV solar cells. The largest
contribution to the collective dose is from the acquisition of raw materials, especially minerals used for
manufacturing the CSP plants or PV solar cells.
185. Silicon-based PV solar cells contain small amounts of rare earth elements, which are of particularinterest because the ores from which they are obtained have relatively high contents of uranium andthorium. Thus, the mining of the ores and their subsequent processing lead to occupational radiation
exposures.
186. Because relevant data on the use of metals and other natural resources were only readily available
for the PV solar systems, the assessment was only carried out for this type of solar energy production.
187. The metals required to construct a solar PV system and amount of ore required are shown intable 43. The results are based on data for a life cycle assessment of a 3 kWp 11 multicrystalline–silicon
PV solar panel mounted on a slanted roof in Switzerland [J1]. The solar panel efficiency was 9.4%. The
solar PV system lifetime was assumed to be 30 years, with the exception of the power inverter thatchanges direct current to alternating current. In this case a lifetime of 15 years was assumed. The parts
included were mainly PV panels, mounting structures, inverter and electric installations. Metals used
for the solar panels included small amounts of rare earth elements, which are derived from ores withhigh contents of thorium or uranium.
188. The amount of ore required per unit of electricity generated depends on the ore grade. In table 43,
the values were based on published ore grades [A8, C3, J1].
10

Also called concentrated solar power and concentrated solar thermal (CST).
The symbol, kWp, stands for the peak power (kW) of the solar PV system—the basic unit for the characterization of thecapacity of PV plants measured in a standardized test at a temperature of 25 °C and an irradiation of 1,000 W/m2.
11

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 209

Table 43. Ore usage at mine for the production of energy by a PV solar system mounted on aslanted roof in Switzerland
Figures are for production of 1 kWh by a 3 kWpa PV solar system; ore grade values provided in table 38

Resource

Amount of metals required per unit
electricity producedb(kg/kW h)

Amount of ore required per unit
electricity produced(kg/kW h)

Aluminium

1.7 × 10−3

1.5 × 10−2

Chromium

2.6 × 10−6

1.0 × 10−5

Copper

4.2 × 10−4

3.0 × 10−2

Iron

1.3 × 10−3

2.7 × 10−3

Lead

1.3 × 10−6

2.2 × 10−5

Magnesium

5.4 × 10−5—

Manganese

4.4 × 10−6

9.3 × 10−6

Molybdenum

4.2 × 10−6

8.3 × 10−3

Nickel

5.0 × 10−8

2.5 × 10−6

Silver

2.9 × 10−6

2.9 × 10−2

Tantalum7.8 × 10−7

2.5 × 10−6Tin

1.6 × 10−6

2.1 × 10−6

Zinc

3.1 × 10−6

4.4 × 10−5aThe symbol, kWp, stands for the peak power (kW) of the solar PV system—the basic unit for the characterization of thecapacity of PV plants measured in a standardized test at a temperature of 25 °C and an irradiation of 1,000 W/m2.

bCalculated based on data in Jungbluth et al. [J1].

189. Wind power plant. The kinetic energy of wind is exploited in wind turbines for electricity
generation. Wind speeds suitable for electricity generation range from 4 to 25 metres per second. These
are attainable practically all over the world, with the exception of some equatorial regions. Wind poweris exploited not only onshore but also offshore, where wind speeds are higher and the wind is typicallyavailable more regularly and for longer periods of time. The depth of water and distance from centres ofdemand onshore are major factors influencing the siting of offshore developments [O3].
190. An assessment of materials used to build an onshore wind power unit was conducted by Whiteand Kulcinski [W5]. The assessment was for a three-blade wind turbine and assumed a 25 MW(e) unit,
with a capacity factor of 24% and a lifetime of 20 years. The data on construction materials issummarized in table 44.
Table 44. Estimates of the amounts of construction materials used in wind power plants
Data in White and Kulcinski [W5] and references therein. Amount presented is for low-alloy and stainless steel and expressedin t/GW(e)PlantaWind power plant [S7]a25 MW(e) plant assumed.

Mass of metals (t)Volume of concrete (m3)85 000 steel

130 000

210 UNSCEAR 2016 REPORT191. Biomass power plant. Biomass fuel includes straw, wood or wood residues from forests, andwood waste from wood-processing plants, such as sawmills or pulp and paper mills. Biomass fuel cancontain varying amounts of radionuclides associated with past emissions or accidents. One example ofthis is fuel containing varying amounts of 137Cs in wood that has been obtained from forested regions
affected by fallout from past nuclear power plant accidents. The 137Cs can be emitted in the flue gases
or concentrated in the ash from the combustion of biomass containing 137Cs and possibly lead to
exposures. In some areas filters are currently in use to decrease the 137Cs in the flue gases, andtreatment or proper burial of the ashes can reduce or prevent exposures. Exposures via these pathways
are not treated further here since the 137Cs originates from nuclear power plant accidents and areconcentrated in the forested areas that have received significant fallout from past accidents.
192. In general terms, biomass fuel has a lower energy density than coal and is more challenging tohandle. As a consequence, the fuel handling equipment is heavier than that for a coal-fired power plant ofequivalent capacity. Technically, biomass power plants can be as large as coal-fired power plants in termsof installed capacity, but, in general, biomass units are not built as large as coal units, mostly because of
the significant difficulties with supply and storage associated with materials of lower energy density [R2].
193. The design of a biomass power plant is not very different from that of a coal-fired power plant.
The main difference is that the ash handling equipment is generally smaller in a biomass plant,
depending on the type of fuel used (biomass typically generates less ash than coal). However, thisdifference may be neglected to a first approximation, and therefore the Committee assumed that thematerials required per kilowatt to build a biomass power plant were the same as those given by Whiteand Kulcinski for a coal-fired power plant [W5]. For the biomass power plant, the capacity factor was
assumed to be 75% and plant lifetime 20 years.
194. Summary. Table 45 shows occupational collective effective doses normalized to energy
production due to mining for ores and the processing needed for construction of the electricitygenerating plants or devices, as estimated in this study. Electricity-generating technologies using coal,natural gas and biomass have the lowest collective doses normalized to electricity generated, followed
by slightly higher values for nuclear power. Wind power shows a larger collective dose by about afactor of 10 compared to the lowest values, and solar power shows the largest value by about a factorof 80. The differences between the various electricity-generating technologies are connected to various
plant facilities using different types and amounts of steel and metals. Another reason is the differencesin the capacity factors and lifetime of the plants, with a higher capacity factor and longer lifetime
implying a lower collective dose per unit of electricity generation.
Table 45. Collective effective dose normalized to unit of electricity generation for construction ofelectricity-generating plants or devices
Electricity-generating technology

Normalized occupational collective dose due to mining
and processing of ores needed for construction(manSv/(GW a))

Nuclear

0.02

Coal

0.01

Natural gas

0.01

Solar PV

0.8

Wind

0.1

Biomass

0.01

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 211

VIII. COMPARISON OF RADIATION EXPOSURES FROMELECTRICITY-GENERATING TECHNOLOGIES
195. Sources of radiation exposure from electricity-generating technologies based on the (a) nuclear
fuel cycle, (b) coal cycle, the combustion of natural gas, oil and biofuels, (c) geothermal energy,(d) wind power and (e) solar energy have been investigated in this annex. Two electricity-generating
technologies, the nuclear fuel cycle and the coal cycle have been substantially investigated using thesame methodology, the Committee’s revised methodology for estimating public exposures due toradioactive discharges (annex A). This same methodology was also used to investigate public exposuresderived with more rudimentary assessments based on less available data for the electricity-generating
technologies that employ combustion of oil and natural gas, and geothermal energy.
196. Occupational exposures for all of these technologies were also estimated, relying mainly on datafrom dosimetric records of worker exposures. In addition, new assessments on occupational exposuresfrom (a) decommissioning of nuclear power reactors and (b) the mining of rare earth metals needed forthe construction phase in different electricity-generating technologies have been presented, adding forthe first time solar energy, wind power and combustion of biomass to the electricity-generating
technologies assessed by the Committee.
197. This chapter compares the results for the different electricity-generating technologies investigatedby applying the revised methodology for estimating public exposures due to radioactive discharges(annex A). This is followed by a comparison of all results for both occupational and public exposures.

A. Comparison of public exposures due to radioactive discharges
from the electricity-generating technologies based on the nuclear
fuel cycle, coal cycle, combustion of natural gas and oil, and use ofgeothermal energy
198. Table 46 shows results from applying the revised methodology for estimating public exposuresdue to radioactive discharges (annex A). The results represent the sum of all discharges assessed in thisannex for the electricity-generating technologies based on the (a) nuclear fuel cycle, (b) coal cycle andcombustion of natural gas and oil, and (c) use of geothermal energy.
199. The table provides estimates of collective doses to the worldwide public, and associated collective
doses normalized to electricity generation in 2010, integrated to 100 years. Collective doses normalized
to electricity generation are the sum of the collective doses for each process in the nuclear fuel cycle, orthe coal cycle, divided by the total electricity generated in 2010 for that cycle. For the coal cycle, thedoses were estimated assuming all discharges were either from older coal plants or from modern ones.
Both results are shown and represent a range of values for the coal cycle. Although the assessments of
the nuclear fuel cycle and coal cycle were more substantive than those for the combustion of oil andnatural gas, and geothermal energy, the Committee considered that comparing the magnitudes of thetotal and normalized collective doses was still valid.

212 UNSCEAR 2016 REPORT

Table 46. Collective dose to the worldwide public, and associated normalized collective dose for 2010,
integrated to 100 years a, b
Except where otherwise specified, the collective doses given are for the local and regional components. Shown also is thepercentage of total world electricity generation in 2010 for each electricity-generating technology and the discharges for222

Rn normalized to the electricity generation in 2010

Electricity-generating
technology

Collective
dose(man Sv)

Normalized
collective dose(man Sv/(GW a))% of total world
electricity
generation in2010

Normalized 222Rn
discharges(TBq/(GW a))

13

Uranium mining – 66

NUCLEAR FUEL CYCLE
Nuclear, total from mining andmilling, power plants andreprocessing, excluding global
component

130

0.43

Milling – 3
Operational milltailings – 3Mill tailingsc – 10

Adding global component
integrated to
100 years

910

3.0

500 years

1 700

5.5

10,000 years

7 600

25

COAL CYCLE
Coal, older coal plants
Coal, modern coal plants

1 400
670

1.4

40

Coal mining – 2.8
Power plants – 0.07

0.7

Ashc – 1.8

OTHERSNatural gas

55

0.10

22

0.75

Oil

0.03

0.000 3

4.6

0.002

Geothermal (low-density
population – default
population)

5–160

1–20

0.3

150

a

Projections of any health effects using collective doses in the table are not recommended.

b

All estimates are calculated based on best estimates; site- and location-specific collective doses are not presented.

c

The values of the normalized 222Rn discharges (TBq/(GW a)) shown in table 46 for uranium mine mill tailings (Mill tailings)
and for coal ash deposits (Ash) were multiplied by 100 to account for radon emanating for 100 years from these surfaces. The
value for coal ash deposits was also multiplied by a factor of 0.6 since only 60% of the ashes produced are deposited.

200. Excluding the global component from the globally-circulating nuclides, electricity generationfrom the coal cycle gave the highest collective dose to the public integrated to 100 years—both for the
total generation in 2010 and when normalized to unit of electricity generated—for both older andmodern coal plants. When the global component resulting from the globally-circulating radionuclides
originating from the nuclear fuel cycle was taken into account, integrated to 100 years, the totalcollective dose from the nuclear fuel cycle was of the same order as from the coal cycle, about in the
middle range of the values for modern versus older coal plants. Because of ongoing global exposuresfrom the globally-circulating radionuclide 14C discharged from nuclear power plants and from

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 213

reprocessing facilities, the collective dose from the nuclear fuel cycle slowly increases over centuries asshown in figure V, in section IV. This is also evident in the doses shown in table 46 for the global
circulation integrated to 100, 500 and 10,000 years. The total collective dose normalized to electricity
generated for the nuclear fuel cycle, including the globally-circulating nuclides and integrated to
10,000 years, is 25 man Sv/(GW a) as shown in table 46. Comparing with earlier UNSCEAR
assessments, the UNSCEAR 2000 Report [U9] estimated the same quantity as 40 man Sv/(GW a). As
described earlier, the collective dose from globally-circulating radionuclides is the sum of very small
doses to the entire world’s population. The local and regional components of collective dose assessed
for the nuclear fuel cycle excluding globally-circulating radionuclides, and for the coal cycle, are forthe local and regional populations exposed to the respective discharges of the source (seeparagraph 34).
201. The magnitude of collective doses resulting from the coal cycle, for both the older and themodern coal plants, is due in large part to (a) the amount of electricity generated from coal (40% of
the world’s electricity generation in 2010), (b) the greater number of coal mines compared touranium mines and (c) the contribution to the collective dose from the discharge of radon and otherradionuclides from the 238U series arising from the combustion of coal at the power plants. Anotherimportant factor is that a larger population base lives near coal plants and coal mines compared tonuclear plants and uranium mines.
202. Natural gas accounted for 22% of the world’s electricity generation in 2010, which led to a larger
contribution to the collective dose than oil, but was still small in relation to coal. Combustion of oil forelectricity generation contributed only marginally to collective doses. These relatively small doses fromnatural gas and oil combustion were however the dominant contribution to collective dose for West
Asia where, as shown in table 2, natural gas and oil accounted for most of the electricity generation inthat region in 2010.
203. Discharges of 222Rn normalized to the electricity generated are also shown in table 46; the largest
value is estimated for geothermal energy. This finding is based on few existing data. However, becauseof the process for generating electricity from geothermal energy discussed in section VI, it is reasonable
that the impact of 222Rn discharges from gases vented during the operation of geothermal electricityproducing plants is relatively significant. The resulting doses from geothermal energy shown in table 46
are assessed using two different population densities, low-density and the default population density(see chapter III), where the available population data discussed in chapter VI suggest the low-density
population distribution may be more realistic. The range in results for the collective dose demonstrates
the strong dependence on the population data. However, the impact of geothermal electricity generation
remains small because of the limited use of this technology.
204. The next largest value for the normalized 222Rn discharge is for uranium mining, which is a factor
of about 20 larger than the value for coal mining. This is also reasonable because uranium mining
occurs where there are high levels of uranium and therefore of its progeny products such as 222Rn.
Shown in table 47 is a comparison of estimated doses to the public from the mining of coal and themining of uranium. The assessments resulted in a larger estimate for the characteristic individual dose
per unit of electricity generated for uranium mining (non-ISL) than for coal mining, consistent with thelarger normalized 222Rn discharge. However, values for both the worldwide collective dose (essentiallycomprising local and regional components alone) and of the associated values normalized to unit ofelectricity generated in 2010 are larger for coal mining than uranium mining. This is because of thelarger number of coal mines in the world and because the population density is higher around coalmines compared to uranium mines.

214 UNSCEAR 2016 REPORT

Table 47. Comparison of doses to the public from mining of coal and mining of uranium
Coal mining

Uranium mining
non-ISLISL−6Characteristic individual dose per unit of electricity generated(Sv/(GW a))

2.4 × 10

5.5 × 10

2.5 × 10−7

Worldwide collective dose (man Sv)

3.7 × 102

4.0 × 101

1.3 × 100

World-average collective dose per unit of electricity generated(man Sv/(GW a))

3.8 × 10−1

2.8 × 10−1

1.3 × 10−2

−7205. Figure VIII shows graphically the contribution each electricity-generating technology consideredmakes to the worldwide collective dose to the public. Two distinct values are presented for the nuclear
fuel cycle: the first is the local and regional component for the population exposed directly to thedischarges, and the second includes the global component for the population of the world as a whole
due to globally circulating radionuclides. The values for both the coal cycle and geothermal energy are
deemed to represent boundary values, assuming either all modern or all older coal plants for the coal
cycle, and for geothermal energy assuming two different population distributions. Figure IX shows thesame sources as figure VIII but for the total collective dose normalized to electricity generated. The
values for the nuclear fuel cycle without global circulation are slightly less than for the coal cycle,
which lies somewhere between the two values given for the modern and older coal plants. Including the
global circulation to the nuclear fuel cycle results in the larger value for these two technologies.
Comparing these data with figure VIII shows how the magnitude of electricity generated for eachtechnology results in the collective dose to the public. Figure IX also shows the magnitude of the upperbound for the collective dose normalized to electricity generation for geothermal energy.
206. Table 48 breaks down the comparison of collective doses—and collective doses normalized toelectricity generated in 2010—for each of the source components of the nuclear fuel cycle and coal
cycle technologies that have been assessed. These components are coal mining, discharges from coal
combustion during power plant operation and from coal ash deposits associated with the coal cycle forelectricity generation. For the nuclear fuel cycle, these components are uranium mining and milling
including mill tailings, discharges during electricity generation from the nuclear power plant anddischarges from reprocessing. Collective doses normalized to electricity generated in 2010 for each of
the source components for the nuclear fuel cycle shown in table 48 use the electricity produced as aresult of each process in the normalizations.
207. Although the differences in collective doses normalized to electricity generation for the various
sources are not very large, (with the exception of old-style coal plants, which have the largest value),
the total collective doses in 2010 show significant differences. For both sets of results, the largest
values are from old-style coal combustion power plants. Collective doses for the different components
associated with the coal cycle are all larger than collective doses for the components associated with thenuclear fuel cycle, except for modern coal plants compared to nuclear power plants. The categories
shown in table 48 make clear the importance of radon discharges from coal mining and from coal ash
deposits, which—with the assumptions made in this study—make significant contributions to the totalcollective doses even when the coal combustion power plant is modern. The differences in the dosevalues for the old-style compared with the modern coal combustion power plant are due to the nuclides
in the 238U decay chain other than radon, because the radon discharge from the flue-gas stack is thesame in both cases. Table 49 shows this difference between modern versus old-style coal combustion
power plants regarding dose contribution per radionuclide from the 238U decay chain.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 215

208. Table 49 shows local and regional components of collective doses assessed from discharges ofradionuclides in the 238U decay chain from the flue-gas stack of a coal plant during the combustion ofcoal for electricity generation in a modern compared to an old-style coal plant. The 222Rn discharges are
the same for both types of coal plant; consequently the doses from 222Rn are also the same in table 49. It
is clear from the breakdown of the radionuclides shown in table 49 that the large differences in thedoses assessed for the old-style versus modern coal plants are due to the other radionuclides shown,with 210Pb and 210Po giving the largest contribution.
Figure VIII. Local and regional components of the collective doses to the public integrated to
100 years, for the electricity-generating technologies based on the nuclear fuel cycle, coal cycle,
combustion of gas and use of geothermal energy
Data from table 46

216 UNSCEAR 2016 REPORT

Figure IX. Local and regional collective doses to the public integrated to 100 years, normalized toelectricity generated for the technologies based on the nuclear fuel cycle, coal cycle, combustion ofgas and use of geothermal energy

Table 48. Comparison of collective doses to the public, and collective doses normalized toelectricity generation in 2010, integrated to 100 years, to the world-average population within a
1,500 km radius of each source for the electricity-generating technologies based on the coal cycle
and the nuclear fuel cycle
Coal
Sourcea

NuclearCollective
dose(man Sv)

Normalized
collective dose(man Sv/(Gw a))

SourceCollective
dose(man Sv)

Normalized
collective dose(man Sv/(Gw a))

Coal mining

370

0.4

Uranium mininga
and milling

53

0.2

Older coal plants

780

0.8

NPP generation

68

0.2

Modern coal plants

60

0.1

From coal ash deposits

240

0.2

Reprocessing

7.6

0.03

Of the 53 man Sv for uranium mining and milling, 40 man Sv is from mining only.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 217

Table 49. Local and regional components of collective dose, integrated to 100 years, due todischarges from modern and old-style coal combustion power plants

Radionuclide

Collective dose from modern coal plants(man Sv)

Collective dose from old-style coal plants(man Sv)

Local

Regional

Local

Regional

Pb

2

9

28

135

Po

3

14

46

216

222

Rn

2

8

2

8

226

Ra

1

7

20

98

230

Th

2

8

26

122

234

U

0.5

2

7

34

238

U

0.4

2

6

29

Totala

11

50

135

641

210
210

a
Slight differences between the totals shown here and those in table 48 are due to rounding performed on the values in table 48.
The values in table 49 are left as generated in the assessment to avoid inconsistent rounding of individual radionuclides for the
purpose of attaining the same total rounded value as shown in table 48. It should also be recalled that in this table, as in all tablesin the annex resulting from assessments, the calculations are done to full precision and any discrepancies in the final sum of
numbers in the tables are due to rounding.

209. Doses assessed from the radon emanation from coal ash versus uranium mill tailings, using 222Rn
discharges normalized to electricity generated given in table 46, are compared in table 50. As describedin chapter IV and more extensively in annex A, a low population density was used in the assessments
on uranium mill tailings whereas a default population density was used in the assessment on coal ash.
The characteristic individual dose per unit of electricity generated is larger for uranium mill tailings
compared to coal ash, reflecting the greater activity concentrations of naturally occurring radionuclides,
including radon, in mill tailings. However, when the total electricity generation is taken into account,collective doses assessed from coal ash are greater than those from mill tailings. The results show theeffect of the continued emanation of radon centuries into the future.
Table 50. Collective dose per unit of electricity generated to world-average population within
1,500 km and integrated to 100 years, also integrated for two different times since disposal, andcharacteristic individual dose per unit of electricity generated from emanation of radon from coalash and mill tailings
Time sincedisposal

Collective dose per unit of electricity
generated (man Sv/(GW a))

Characteristic individual dose per unit ofelectricity generated (Sv/(GW a))

Coal ash

Mill tailings

Coal ash

Mill tailings

100 years

0.2

0.04

1.5 × 10

8.4 × 10−7

500 years

1.2

0.2

−7

210. Significant radionuclides from nuclear power plant discharges. The importance of 14C and theglobally circulating radionuclides, and 222Rn and the 238U series, is discussed in other sections. The
following concentrates on comparing other significant radionuclides discharged from different nuclear
power plant types.

218 UNSCEAR 2016 REPORT211. Table 51 shows the world-average collective dose per unit of electricity generated for the nuclear
power plant types assessed in this annex. The table illustrates the relative importance to the worldaverage collective dose per unit of electricity generated of 90Sr, 137Cs and 35S from the GCR-type
reactor, and tritium from the HWR reactor.
Table 51. Local and regional components of collective doses, integrated to 100 years, normalized
to electricity generation in 2010, shown for all nuclear power plant types considered
See chapter IV for details on the power plant types

Discharged
radionuclide

World-average collective dose per unit of electricity generated for each nuclear power plant type(man Sv/(GW a))
PWR

BWR

HWR

LWGR

AGR

GCR

FBR

H

7.2 × 10

8.6 × 10

1.7 × 10

1.2 × 10

1.9 × 10

3.3 × 10

2.3 × 10−1

14C

4.6 × 10−2

3.2 × 10−2

1.4 × 10−1

3.1 × 10−1

3.5 × 10−1

1.3 × 100

2.8 × 10−2

35

S

0

0

0

0

4.5 × 10−2

2.4 × 10−1

0

41

Ar

4.5 × 10

4.6 × 10

1.3 × 10

3.4 × 10

8.4 × 10

7.0 × 10

0

54

Mn

2.3 × 10−4

1.0 × 100

−5

4.0 × 1000

1.8 × 10−458

Co

3.7 × 10−5

4.0 × 10−50

5.9 × 10−600

1.7 × 10−460

Co

2.0 × 10−4

5.6 × 10−3

1.5 × 10−4

8.8 × 10−3

7.0 × 10−2

3.5 × 10−3

2.3 × 10−3

65

Zn

0

5.0 × 10−40

1.6 × 10−4

7.0 × 10−5

0

0

85

Kr

2.2 × 10−5

1.4 × 10

0

0

0

0

1.7 × 10−490Sr

7.0 × 10−5

1.2 × 10−30

1.2 × 10−2

3.4 × 10−5

4.7 × 10−1

0

106

Ru

3.7 × 10−5

1.0 × 10−6

0

0

0

0

0

131

I

6.0 × 10

3.2 × 10

1.7 × 10

7.5 × 10

2.5 × 100

1.5 × 10−4

133

Xe

1.2 × 10−4

1.5 × 10−40

1.1 × 10−2

0

0

0

135

Xe

3.4 × 10−5

3.4 × 10−40

2.4 × 10−3

0

0

0

138

Xe

1.5 × 10−9

3.1 × 10−70

1.1 × 10−6

0

134

Cs

3.2 × 10

2.6 × 100

3.4 × 10137

Cs

4.2 × 10−4

1.0 × 10−3

4.6 × 10−2

Total

1.2 × 10−1

5.0 × 10−2

1.9 × 100

3

−2

−5
−5

−6

−5

−4

−3

−5

−4

−4

0

−3

−5

−1

−3

−3

−2

−4

−5

−2

−2

0

0

7.4 × 10−2

3.8 × 10

1.1 × 10−2

6.6 × 10−3

6.6 × 10−4

6.2 × 10−1

7.7 × 10−3

4.8 × 10−1

4.9 × 10−1

2.8 × 100

2.8 × 10−1−4−4212. Figure X shows estimated collective doses by world region and totalled for the world by nuclear
reactor type, with the values taken from table 51. Both collective doses, and collective doses
normalized to electricity generated in 2010 are shown. The collective doses show a different profile
than that shown in the figure displaying the collective doses normalized to electricity generated. Twonuclear power plant types, GCRs and HWRs, dominate the collective doses normalized to electricity
generated. However it is the power plant types PWRs and HWRs that dominate the contribution to thecollective doses; this is because of a combination of the number of reactors that exist in each region
coupled with the discharged amounts of significant radionuclides for each reactor type.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 219

Figure X. Local and regional collective doses to the public integrated to 100 years, and associatedvalues normalized to electricity generation, by world region and nuclear reactor type
Results shown for collective doses from all discharges (man Sv) and for collective dose from all discharges normalized toelectricity generation (man Sv/(GW a)) in 2010. Note that the order of the individual reactor types is according to their world
share of electricity generated in 2010, as shown in table 9220 UNSCEAR 2016 REPORT

B. Comparison of public and occupational exposures
213. Table 52 summarizes the results from all occupational and public exposures that have beenpresented in this annex. The table shows collective doses normalized to electricity generated in 2010,
followed by collective doses for the same year. The local and regional components of doses to the
public are for world-average populations out to 100 km and between 100 km and 1,500 km from afacility respectively, integrated over 100 years. Other conditions are as stated earlier in this chapter.
This section considers the total exposure from each electricity-generating technology, i.e. public plus
occupational exposures.
214. The occupational exposures from coal mining dominate collective doses, so that the total
exposure of the public and occupational exposures combined is largest from the coal cycle. This is true
even considering the public exposures from the nuclear fuel cycle including the global circulation
integrated out to 10,000 years. The next largest is the total exposure from the nuclear fuel cycle, where
public exposure including global circulation gives the highest dose followed by occupational exposures.
The same trend is observed for the associated collective doses normalized to electricity generated, withthe exception of the public exposures from the nuclear fuel cycle normalized to electricity generated
integrated out to 10,000 years, which gave the largest value.
215. The comparisons of public doses shown in table 52 that have been assessed using the revised
methodology for estimating public exposures due to radioactive discharges (annex A) have been describedin section A of this chapter. They are provided in this table to facilitate comparison with occupational
exposures and view the entire exposure, public and occupational, from each electricity-generating
technology. One comparison of interest noticeable in table 52 is that collective doses to the public
resulting from radon discharges from natural gas power plants are about the same as collective doses tothe public resulting from the discharges from operation of nuclear power plants (55 versus 68 man Sv,
respectively). However, the collective dose to the public normalized to electricity generated in 2010 isabout a factor of four larger for the nuclear fuel cycle than for the combustion of natural gas (not including
global circulation). The relatively large actual collective dose assessed for combustion of natural gas
results from the larger amount of electricity produced by natural gas than nuclear power in 2010.
216. The assessments on collective dose provided in the annex can be put in perspective byconsidering natural exposure to background radiation from 222Rn. All doses to the public from naturalgas, oil, and geothermal energy are calculated from the release of 222Rn. A significant fraction of thedoses from the coal and nuclear cycles are also due to the release of 222Rn. Radon-222 occurs
naturally in nature and is a major fraction of the dose to man from all naturally occurring sources ofradiation. Wilkening et al. [W7] have estimated that the average exhalation rate of 222Rn from soil is
1.6 × 10−2 Bq (m2 s)−1. As a result of this exhalation from soil and the additional exhalation of 222Rn
from building materials, the Committee [U9] has previously estimated that the average effective dosefrom 222Rn is 1.15 mSv a−1. For the world population considered here, the collective dose due toexposure to naturally occurring radon is 11,500,000 man Sv a−1, a value much larger than any given
in table 46.
217. Total occupational collective doses are significantly greater than public collective doses for thecoal cycle and combustion of oil, from a factor of about ten for coal plants to about five hundred for
combustion of oil, although the values for combustion of oil are in comparison small. For natural gas
and geothermal energy, total public collective doses were assessed to be larger than total occupational
collective doses. Total occupational collective doses for the nuclear fuel cycle are of about the sameorder as the total collective dose to the public when including the global circulation to 100 years.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 221

218. The largest occupational collective dose normalized to energy generated in 2010 resulting fromthe mining for metals for construction materials was from solar photovoltaic (PV) technology, which
was a factor of forty and eighty larger than for the nuclear fuel cycle and coal cycle, respectively. Thiswas followed by the occupational collective dose for wind power, which was also larger than the values
for the nuclear fuel cycle and coal cycle. These differences come from the different metal requirements
for solar PV and wind power technologies, discussed in chapter VII.

Nuclear fuel cycle

Coal

Moderncoal plant

Older coalplant

Natural gas

Oil

Geothermal

Solar PVWind

Biomass

0.8

0.1

0.01

COLLECTIVE DOSES NORMALIZED TO ELECTRICITY GENERATION (man Sv/(GW a))
PublicU mining andmilling

0.2

Coal mining(Rn discharge)

0.4

0.4

Nuclear power
plant operation

0.2

Coal plant
operation

0.1

0.8

Reprocessing

0.03

Ash, radon
emanation

0.2

0.2

0.7

1.4

0.1

0.000 3

1–20

11

11

0.01

0.15

0.05

0.01

0.01

0.01

Total public (not including globally
circulating radionuclides)

0.43

Total public (including
globally circulating
radionuclides)

100 years

3.0

500 years

5.5

10 000 years

25

Occupational

2.7c

Occupational - decommissioning, nuclear

1.8

Occupational - mining for construction

0.02

COLLECTIVE DOSES (man Sv)
PublicU mining andmilling

53

Coal mining(Rn discharge)

370

370

Nuclear power
plant operation

68

Coal plant
operation

60

780

Reprocessing

8Ash, radon
emanation

240

240

670

1 400

Total public (not including globally
circulating radionuclides)

130

55

0.03

5–160

222 UNSCEAR 2016 REPORT

Table 52. Comparison of the public and occupational exposures assessed in this annexa,b

Nuclear fuel cycle
Total public (including
globally circulating
radionuclides)

Coal

100 years

910

500 years

1 700

10 000 years

7 600

Occupational

788

Decommissioning (occupational) per nuclear
reactor5

Occupational - mining for construction

6

Moderncoal plant

Older coalplant

Natural gas

Oil

Geothermal

11 000

11 0007

17

0.4–0.8

7

7

3

a

Projections of any health effects using collective doses in the table are not recommended.

b

All estimates are calculated based on best estimates; site- and location-specific collective doses are not presented.

c

From UNSCEAR 2008 Report.

Solar PVWind

Biomass

34

0.4

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 223

224 UNSCEAR 2016 REPORTC. Commentary on significance of accidents
219. While this annex has focused on comparing the exposures from normal operations of the various
electricity-generating technologies, a commentary is needed regarding the risk of serious accidents thatgive rise to radiation exposure. This is clearly only significant for nuclear power. While it is beyond thecompetence of the Committee to assess the probability of any future accident, the Committee has
assessed information on past accidents that have exposed the public and workers, notably the 1986
accident at the Chernobyl nuclear power plant in the former Soviet Union (UNSCEAR 2008 Report,annexes C and D [U12]). Moreover since that report, the Committee has conducted an assessment of
the levels and effects of radiation exposure due to the 2011 nuclear accident at the Fukushima Daiichi
nuclear power station (FDNPS). The UNSCEAR 2013 Report [U13] gave estimates of the collective
doses to the population of Japan due to the FDNPS accident. These estimates were compared with theprevious estimates by the Committee for populations of European countries exposed to radiation
following the 1986 Chernobyl accident in the former Soviet Union. This comparison is shown intable 53.
Table 53. Estimates of the collective effective doses from the FDNPS and Chernobyl accidents [U13]
Accident

Fukushima Daiichi nuclear power station
Chernobyl Unit 4aCollective effective dose (thousand man Sv)
Over first year

Over ten years

Up to age 80 yearsa

18

36

48
400

Summing the dose to all exposed individuals integrated from their age at the time of the accident until they reach age 80 years.

220. Comparing radiation doses alone, it is clear that serious accidents give rise to collective doses thatare very many times greater than collective doses due to normal operations. For example, the collective
dose from the single accident at Chernobyl Unit 4 is more than 400 times the annual global collective
dose to the public from all nuclear power. Such a comparison of collective doses can be made only inorder to gain perspective on the magnitude of the radiological impact. However, great care must be
taken when assimilating these comparisons, because there are obviously many major non-radiologicaldifferences between accidents and normal operations for electricity generation. While collective doses
from such accidents have been much larger than those from annual normal operations, the distributionof doses is more localized geographically, whereas the collective doses from normal operations forelectricity generation are population-averaged over geographical regions or the world as a whole.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 225

IX. RESEARCH NEEDS
221. The use of electricity is ubiquitous throughout the world and the demand for electricity continues
to grow. Because of the prevalence of different types of electricity-generating technologies coupledwith the diversity inherent in the countries of the world, standard data on exposures to radioactive
discharges or materials is non-uniform or non-existent. Thus, any study comparing different electricitygenerating technologies will suffer from inconsistent or incomplete databases. This annex has usedconsistent methodologies together with sound judgement on the use and interpretation of available datato assess public and occupational exposures from electricity-generating technologies. Some specificareas of interest that would improve understanding or that would increase the certainty of this work areincluded in this section.
222. Data. Consistent data on radioactive discharges across all electricity-generating technologies,
including how they change with time and changing practices across all electricity-generating
technologies, would improve the ability to assess and compare these technologies. Data on occupational
exposures and their change with time and practice for each technology would also help. The following
lists some specific needs that have been identified:(a) Decommissioning of coal power plants generates NORM wastes, the magnitude of whichshould be quantified (NORM wastes are naturally occurring radionuclides such as lead andpolonium which in this case become plated out in pipes and the boiler, and other mechanical partsof the plant [M5]). Occupational exposures received during decommissioning of coal power plants
need to be monitored in a representative manner in order to allow assessment of this exposure routefor workers. This is also true regarding occupational exposures of workers decommissioning
nuclear power plants.
(b) The combustion of natural gas and oil for electricity generation causes naturally occurring
radionuclides to accumulate in the pipes. Assessment of any occupational or public exposures fromthis source is dependent on representative data on the activity concentrations that occur. There aresome data available on this, however the measured external gamma exposure rates are of moresignificance.
(c) In order to assess the occupational doses resulting from the management of radioactive waste
generated in nuclear power plants, representative data on doses received by workers needs to becollected or compiled [M5].
(d) Information on the proportion of nuclear fuel reprocessed for peaceful (i.e. electricity
generation) versus non-peaceful (i.e. military applications) has only been available for the La
Hague, France reprocessing facility and therefore the assessment of the doses due to discharges
from reprocessing activities was compromised by the lack of open information. Also, some of theoccupational doses at reprocessing facilities, such as Sellafield in the United Kingdom, are fromthe reprocessing of historical wastes so it is very hard to relate occupational doses to just nuclear
fuel that has been reprocessed, or to electricity generation. Actual data on reprocessing activities
related to electricity generation are needed to improve the exposure assessments.
(e) Storage, transmission and distribution of electricity are not included in the assessments of dosefrom electricity generation presented in this annex. These infrastructures are used regardless of thetype of electricity generation (although there are differences between centralized and distributed
grid systems that are optimal for different technologies), and these kinds of dose assessments could
be included in any future update of the evaluations presented in the annex.

226 UNSCEAR 2016 REPORT223. Databases could be consistent and standardized; these types of efforts are currently conducted byvarious international organizations. The success of these efforts, however, depends on the volunteer
compliance of member countries submitting annual reports to these organizations, such as UNSCEAR,
IAEA and NEA, and on funding and interest to keep the programmes current. These programmes
would benefit from extending the data-collection schemes to include all relevant technologies, andinvesting in research to identify significant parameters to help design streamlined data-collection
schemes that would further decrease the burden on member countries and data-collecting organizations.
224. Uranium mining using the ISL process. Explanations were given in chapter IV of the assumptions
made about radon releases during uranium mining using the ISL process. Existing data areinconclusive, and radon could be released in varying quantities from the lixiviant used to dissolve theuranium in the underground ore body, depending on the specifics of the process. Determining more
precisely the amount of radon releases from ISL uranium recovery, and how to better generalize them,requires additional research and a more complete understanding of the variation in the global processes
and procedures currently in use for mining of uranium using the ISL process.
225. Recycling of metals. In chapter VII on assessing doses from the mining of metals needed in theconstruction phase, several assumptions were made in the assessments. Although the assumptions affectthe absolute value of the assessed collective doses, the relative importance when comparing thedifferent electricity-generating technologies with regards to the construction phase should remain
relevant. However, these assumptions make a difference when comparing to the other components inthe full cycle of electricity generation, such as in table 52. One area of importance that could be studied
further is the effect of recycling metals, which has not been included in the present study because of
insufficient data for large portions of the world. By not explicitly considering recycling, the collective
dose per unit of electricity generated due to the extraction and processing of the metals used in theconstruction of a facility is likely to be overestimated by around a factor of two on average, andpossibly a factor of five or more for some regions of the world with developed recycling processes(assuming the recycled metal is not contaminated with radioactive material). Also, recycling may affectthe different technologies somewhat differently because it is more difficult and costly to recycle
electronics and components (such as in solar panels) than larger bulk metal objects (such as in moreconventional power plant components).
226. Reference year. This annex concentrates on the reference year 2010 for the assessments. Lifetime
commitments would be relevant for comparing the full impact of the different electricity-generating
technologies. Confounders in this type of extrapolation are associated with the changes in the respective
electricity-generating technologies over time, caused by, for example, improving effectiveness of eachtechnology in both economic and environmental terms and in radiation protection practices.
227. Suess effect. The displacement of 14C in atmospheric CO2 discharges during the coal cycle andother fossil fuel releases of CO2 (which are depleted in 14C) is known as the Suess effect. Research into
this effect would improve knowledge of the global carbon cycle and could help clarify the role of 14C inhuman exposures in the modern world.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 227

X. CONCLUSIONS
228. This annex provides estimates of exposures from the various electricity-generating technologies
that (a) may be used by researchers and policy-makers in their own more comprehensive assessments
for developing energy policy; (b) can be used to help inform the media and the public on these mattersin a balanced perspective; and (c) can highlight possible emerging issues or opportunities for
improvement that may warrant more attention and scrutiny, or future research. The following
summarizes the findings.
229. The Committee has updated its methodology for estimating public exposures due to radioactive
discharges, which is now more flexible for use in evaluating radiation exposures to the public fromdiverse electricity-generating technologies. This methodology along with extensive data collection andanalyses has provided the Committee with a sounder basis for comparative studies than was possible
earlier. The Committee has also re-evaluated occupational exposures arising from different electricitygenerating technologies relying on data mainly from dosimetry records of worker exposures. Theseevaluations comprised the basis for the current comparative study on radiation exposures of both the
public and of workers from electricity generation. To compare exposures, the Committee has focusedon two metrics. These were (a) the collective doses to defined population groups integrated over
specific time periods resulting from one year’s electricity generation by each technology in eachgeographical region and for the world as a whole, and (b) the relevant collective doses divided by theamount of electricity generated by each technology. The year 2010 was used as the reference year forthe comparisons.
230. The Committee has conducted this comparative study by investigating sources of radiation
exposure from electricity-generating technologies based on the (a) nuclear fuel cycle, (b) coal cycle, thecombustion of natural gas, oil and biofuels, and (c) geothermal, wind and solar power. Two electricitygenerating technologies (nuclear fuel cycle and coal cycle) were investigated in detail because a morerobust database existed for these technologies. The Committee evaluated the main sources ofradioactive discharges from the life cycle of these electricity-generating technologies. For the nuclear
fuel cycle, these were uranium mining, milling and mill tailings, power plant operation, spent fuelreprocessing, and decommissioning activities. For the coal cycle, they were the mining for coal, power
plant operation for both a modern coal plant and an older-style coal plant, and coal ash deposits.
231. The Committee estimated that, excluding long-lived globally circulating radionuclides, thecontribution from the coal cycle, assuming discharges from a modern coal plant, was more than half of thetotal collective dose to the global public from the discharges due to a single year’s global electricity
generation, while the nuclear fuel cycle contributed less than a fifth. The contribution from the coal cycle
comes from discharges of natural radionuclides (primarily radon and its radioactive progeny) during coalmining, combustion of coal at the power plant and from coal ash deposits. Similarly, almost half of thecontribution to public exposures from the nuclear fuel cycle also comes from discharges of naturalradionuclides during uranium mining and milling activities. These values depend on the amount ofelectricity generated by each technology; in 2010, the coal cycle produced the largest amount of total
electricity generated, about 40%. Although radon and its progeny are relatively important contributors tocollective doses to the public for both the nuclear fuel cycle and the coal cycle, the associated individual
doses are small compared with doses due to inhalation of naturally occurring radon [U9, W7].

228 UNSCEAR 2016 REPORT232. The Committee found, however, that the contribution to the collective dose to the public fromeach electricity-generating technology was not only because of how much electricity each technology
generated. There were differences due to the collective doses per unit of electricity generated. In normal
operations, the coal cycle gives a higher collective dose per unit of electricity generated than electricity
generation from nuclear power plants, and significantly higher than the other technologies evaluated—with the exception of geothermal energy. Based on the limited information on radon discharges fromgeothermal power plants, the collective dose per unit of electricity generated by geothermal energy
could be significant. However, because the use of geothermal technology is not widespread, its
contribution to radiation exposures of the global public is smaller than that of the coal cycle.
233. Previous investigations on electricity generation from the nuclear fuel cycle have examined thecontribution to human exposures made by long-lived radionuclides, such as carbon-14, which aredischarged, circulate globally and continue to contribute to radiation exposures of the public centuries
into the future, albeit as extremely small individual doses. The Committee found that public exposuresfrom these globally-circulating radionuclides, for one year of discharge and integrated to 100 years,
result in a contribution to exposures from the nuclear fuel cycle that are about the same as the coal
cycle. Over long integration times such as hundreds of years, the contribution from these radionuclides
results in larger collective doses to the global public from the nuclear fuel cycle than the coal cycle.
234. The Committee also assessed the occupational exposures for these technologies. The largest
collective dose to workers per unit of electricity generated resulted from coal mining, because of
exposures to naturally occurring radionuclides. Of all the collective doses evaluated, to both the public
and workers, the exposure of workers from coal mining gave the largest contribution, although it has
fallen over time because of better mining conditions. Regarding the mining of rare earth metals needed
for construction, by far the largest collective dose to workers per unit of electricity generated assessedin this study came from solar power, followed by wind power. This is because the workers are exposed
to natural radionuclides during mining, and the amount of low-grade ore required to be mined for thesetechnologies is high.
235. The total collective dose (i.e. to the global public and all exposed workers combined) per unit ofelectricity generated by the coal cycle was larger than that generated by the nuclear fuel cycle, evenwhen considering the long-lived globally-circulating radionuclides integrated out to 500 years. Whenconsidering the amount of electricity generated in the year 2010 by each technology, the coal cycle
resulted in the largest collective dose to the global public and workers combined, followed by the
nuclear fuel cycle. Of the remaining technologies, geothermal energy and combustion of natural gas
were the next largest contributors.
236. Great care should be taken when interpreting and using these results, because this analysis only
gives a perspective on the magnitude and differences of radiation exposures, and cannot be used todetermine whether one form of energy generation is preferable to another. As stated earlier, a numberof factors determine why a certain mix of energy generation technologies may be selected by countries.
Radiation exposure is only one of them.
237. Moreover, when comparing collective doses for the various electricity-generating technologies, it
is important to note that the collective dose from serious accidents, such as those that occurred at the
Chernobyl and Fukushima-Daiichi nuclear power stations, were orders of magnitude larger than thecollective doses to the world population from one year’s normal operation of all the technologies ofelectricity generation that were assessed in the annex. More significantly, the distribution of doses afteran accident is more localized geographically (local populations receive higher doses than the average),whereas the collective doses from normal operations for electricity generation are often averaged over
populations within each geographical region or the world as a whole.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 229

XI. ACKNOWLEDGEMENTS
The Committee wishes to acknowledge with gratitude the experts directly involved in compiling data,
conducting analyses and quality control checks, and drafting text and conclusions, and the support staff
in their institutes who assisted them. In particular, the Committee would like to thank: the consultant,
L. Hubbard; the expert group, who provided valuable oversight and advice for the work; staff,
T. Anderson and K. Jones, from Public Health England in the United Kingdom, who conducted manyof the analyses; L. Anspaugh, H. Grogan, V. Holahan and J. Simmonds, who provided helpful advice,drafted text, and reviewed and checked the results. The views expressed in this scientific annex andrelated material remain those of the Committee and do not necessarily represent the views of individual
experts, the United Nations or its Member States.
Expert group
Co-chairs: E. Waller (Canada)
Members: T. Anderson (United Kingdom), L. Anspaugh (United States), H. Grogan (United States),
G. Hirth (Australia), V. Holahan (United States), K. Jones (United Kingdom), L. Hubbard (Sweden),
R. Michel (Germany), J. Simmonds (United Kingdom)
Observer: G. Proehl (IAEA)The Committee would also like to acknowledge with gratitude: E. Rochedo, who conducted some ofthe preparatory work to compile data; B. Lauritzen, who compiled some of the data and conducted
some of the analyses on occupational exposures due to mining of metals for construction purposes; andM. Tzivaki for conducting some specialized quality checks of the calculations.

ANNEX B: RADIATION EXPOSURES FROM ELECTRICITY GENERATION 231

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ANNEX CBIOLOGICAL EFFECTS OF SELECTED INTERNALEMITTERS—TRITIUM

241

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

243

CONTENTSI.

INTRODUCTION ................................................................................................................................. 245

II. SOURCES AND LEVELS .................................................................................................................... 248
A. Natural sources .......................................................................................................................... 248
B. Artificial sources ........................................................................................................................ 248
III. PHYSICAL, RADIOLOGICAL AND BIOCHEMICAL CHARACTERISTICS .............................. 252
A. Physical characteristics ........................................................................................................... 252
B. Radiological characteristics................................................................................................... 252
C. Biochemical characteristics ................................................................................................... 253
IV. HUMAN EXPOSURE .......................................................................................................................... 255
A. Exposure of the public ............................................................................................................ 255
B. Occupational exposure .......................................................................................................... 258
V. BIOKINETICS AND DOSIMETRY .................................................................................................... 259
A. Information on biokinetics and dosimetry of tritiated compounds ............................ 259
B. Overview of current biokinetic models for tritium ....................................................... 272
C. Intakes of tritium in relation to pregnancy and breast-feeding .............................. 281
D. Uncertainties in dose coefficients for tritium ................................................................. 284
E. Summary of biokinetic and dosimetic models .............................................................. 285
VI. BIOLOGICAL AND HEALTH EFFECTS .......................................................................................... 287
A. Non-radiological effects of tritium in biological systems .......................................... 287
B. Deterministic effects................................................................................................................ 288
C. Stochastic effects of HTO in mammals ............................................................................. 295
D. Effects of tritiated biochemical substrates ...................................................................... 299
VII. RELATIVE BIOLOGICAL EFFECTIVENESS .................................................................................... 304
A. Track structure considerations ............................................................................................ 305
B. RBE literature reviews and experimental studies.......................................................... 307
C. Factors affecting RBE values ................................................................................................. 308
D. Summary of RBE value determinations ............................................................................ 309
VIII. EPIDEMIOLOGICAL STUDIES ......................................................................................................... 316
A. Studies of occupational exposure ...................................................................................... 316
B. Studies of environmental exposure ................................................................................... 320
C. Summary of epidemiological studies................................................................................ 321

244 UNSCEAR 2016 REPORTIX. RESEARCH NEEDS ............................................................................................................................. 322
X. GENERAL CONCLUSIONS ............................................................................................................... 324
XI. ACKNOWLEDGEMENTS .................................................................................................................. 325
APPENDIX A: TABLES SUMMARIZING STUDIES OF OCCUPATIONAL AND ENVIRONMENTALEXPOSURE TO TRITIUM ........................................................................................................................... 327
REFERENCE ................................................................................................................................................. 341

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

245

I. INTRODUCTION
1.
The Committee has conducted an independent review of the scientific literature on thecharacteristics of tritium, its biokinetics and dosimetry within the human body for various physical andchemical forms and routes of intake into the body, radiobiological effects of tritium exposure, andepidemiological data relating to its impact on the health of workers and members of the public.
2.
Tritium is a radioactive isotope of hydrogen (symbol 3H, but commonly represented by T).
Chemically, it behaves like other isotopes of hydrogen (protium, 1H, the principal stable isotope, anddeuterium, 2H, the other stable isotope). The word tritium is used here to mean the particular isotope ofhydrogen irrespective of the chemical form in which it occurs.
3.
Tritium occurs both naturally, mainly as a result of the interaction of cosmic-ray particles with theatomic nuclei of air molecules in the upper atmosphere, and as a consequence of the operation of
nuclear reactors and other industries. Tritium in the environment and workplace is encounteredpredominantly as tritiated water (HTO) in liquid or vapour form.
4.
Tritium emits low-energy beta particles with a short range in body tissues and, therefore, poses arisk to health as a result of internal exposure only following ingestion in drinking water or food, or
inhalation or absorption through the skin. Unlike external penetrating radiation, such as X-rays andgamma rays, internal exposure to tritium has the potential to result in heterogeneous dose distribution
within tissues and cells. Other factors that may affect the potential radiotoxicity of tritium include
transmutation and isotopic effects. Transmutation is the term used for the formation of a new elementby radioactive decay, which has the potential to adversely affect metabolic processes. Isotopic effects
apply to low atomic mass elements such as hydrogen, for which tritium atoms with larger mass may
replace the stable protium in cellular processes. Both effects are judged to be minor contributors toradiotoxicity when compared to the predominant effect of the energy deposition from beta particles
emitted by tritium decay.
5.
Five main chemical forms are of interest when considering the biological and health effects ofinternal exposure to tritium: HTO, organically bound tritium (OBT), tritiated biochemical substrates(including DNA precursors), insoluble compounds, and tritiated gases. OBT is the general term used todescribe tritium that is non-exchangeably bound to carbon atoms within organic constituents of cells
and tissues (e.g. proteins, polysaccharides, lipids).
6.
Absorbed doses arising from the intake of tritium cannot be measured directly and recourse has tobe made to the use of bioassay (such as the determination of tritium in urine) or to assessments basedon environmental monitoring. Biokinetic models of the behaviour of tritium in the body are used todetermine intake from such measurements and are also used together with dosimetric models to relateretention of tritium in body tissues to the time-course of dose delivery within tissues. For intake oftritium as HTO, distribution between organs and tissues and within cells is quite uniform, depending on
their water content, and so the dose is uniformly delivered despite the short range of the low-energy
beta particle emissions.
7.
However, some organic substrates containing tritium concentrate in specific organs and tissues,and even within specific regions within cells. In such cases, the pattern of dose distribution is very
different from that experienced following uniform exposure to external penetrating radiation or246 UNSCEAR 2016 REPORT

incorporation of HTO, with heterogeneity of dose between organs and tissues, and potentially withinorgans and even within cells. For intake of tritiated nucleotides and nucleosides, for example, a small
proportion has been shown to reach cells intact and may then be incorporated into cellular DNA,
resulting in localized energy deposition [D5, N1].
8.
There is also some tritium-containing radioactive material with low solubility in aqueous media,
such as tritides of metals (e.g. Ti, Zr, Hf), tritiated luminous compounds, micro fragments of glass andcarbon and beryllium particles contaminated with tritium. Such inhaled particles exhibit long-term
retention in the lungs, leading to prolonged exposure of lung tissue to beta radiation.
9.
The International Commission on Radiological Protection (ICRP) has used three main biokinetic
models in the estimation of doses from compounds that contain tritium for protection purposes [I8, I9,
I10, I14, I15, I18]:(a) A model for tritium absorbed to blood as HTO following either ingestion or inhalation, appliedalso to other tritiated compounds, including elemental hydrogen and methane, that partially convert
to HTO after being taken into the body;(b) A model for tritium absorbed to blood as OBT, mainly following ingestion in food, but also
applied to inhalation of non-specified organic material and to ingestion or inhalation of somespecific tritiated organic compounds;(c) The generic ICRP models for the human respiratory tract, specifying absorption parametervalues for inhalation of insoluble forms of tritium used in industry, including metal tritides.
10. The existing ICRP biokinetic and dosimetric models for tritium are currently being upgraded on
the basis of recent biokinetic data, especially for recently developed physical and chemical forms oftritium. This work includes models for tritium as gases, HTO, organic substances and OBT, andmaterial with low solubility.
11. Electrons with very low energy, including beta particles from 3H, have higher linear energy
transfer (LET) values than electrons generated by the interaction of higher energy photons (e.g. fromexternal gamma rays). This higher LET may result in greater effectiveness in causing cancer. The
assessment of the effectiveness of different radiation in causing health effects relies on data on theirrelative biological effectiveness (RBE). RBE is an empirical quantity that depends on the biologicalsystem, the observed end points, the dose and the experimental conditions. In recent decades, several
tens of experiments have been conducted using mammals (mostly mice) and their cells to determine
RBE for tritium under various experimental conditions and considering a range of biological end
points. However, only a small number of studies were performed to directly measure cancer induction
in mammals.
12. Laboratory studies using animals have demonstrated that tritium, like other sources of radiation,
can interfere with the development of the embryo or fetus, and can induce carcinogenic, heritable andreproductive effects and cell death. The use of high doses of tritium, for example, in the form of HTO
or tritiated thymidine, has also been shown to induce acute radiation syndrome.
13. The dose and risk from some tritiated biochemical substrates and OBT is greater than that from
HTO due to their longer residence in the body. However, there are few studies looking specifically at
biological effects related to tritiated biochemical substrates and most of them use DNA precursors andamino acids. There is no appropriate ICRP biokinetic and dosimetric model for use in human risk
assessment and radiation protection for tritiated nuclear acid precursors and there is a practical need forANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

247

the development of such models for intake of tritiated biochemical substrates, including nucleotropic
forms even though the number of workers dealing with such forms of tritium is limited.
14. Most experimental studies on tritium were performed 20 to 30 years ago. While this work wascompetently performed at the time, it did not use modern scientific approaches and procedures that areoften more sensitive and can use multiple approaches to test a single question. The application ofmodern techniques would be helpful in reinvestigating aspects of tritium dosimetry and effects,
including fetal and embryo studies, and DNA damage analyses.
15. Workers may be subjected to wide-range occupational exposure to tritium in various chemical andphysical forms. Usually, occupational exposure to tritium is low relative to other sources of exposure.
However, historically there have been several cases of occupational exposure of workers (RussianFederation, Germany), mostly following accidents but also following chronic exposure to considerable
quantities of tritium, resulting in haematological radiation syndrome [M15, O3, S12], including a few
cases of radiation-induced death.
16. The principal source of quantitative information on radiation-induced cancer and other healtheffects in humans remains the epidemiological follow-up studies of the Japanese survivors of theatomic bombings exposed to external radiation [P12, P13, U11]. An important question is the extent towhich these risk estimates are applicable to exposure from internal emitters, including tritium with its
low dose-rate and low-energy beta radiation with heterogeneity of exposure between and within organs
and tissues. Currently, little information on tritium-specific risks can be derived from epidemiologicalstudies of tritium workers or members of the public potentially exposed to tritium, beyond theconclusion that tritium-specific risks have not been substantially underestimated.
17. The Committee has agreed to undertake a comprehensive review of the biokinetics, dosimetry andeffects of selected internal emitters. The first radionuclide to be considered is the radioisotope of theelement hydrogen, tritium. The main reasons for this selection are as follows:−

The potential for large scale production of tritium in connection with civilian and military fusion
activities, as well as its creation as a by-product from operation of nuclear fission reactors,
especially heavy water reactors;−

Exposure of workers and the public to various physical and chemical forms of tritium, includingorganic and substrates with low solubility, with a wide range of radiotoxicity that requires
comprehensive scientific analyses;−

Professional and public concerns expressed between 2006 and 2010 regarding the radiotoxicity oftritium, which led to extensive review and data analysis in a number of countries, including
Canada, France and the United Kingdom.

248 UNSCEAR 2016 REPORT

II. SOURCES AND LEVELS

A. Natural sources
18. Tritium was discovered in 1934 by Oliphant, Harteck and Rutherford [O4] and isolated in 1939
by Alvarez and Cornog [A2]; the production of tritium by natural processes was reported by Libby[L10]. There are three main sources of natural tritium: production in the atmosphere by galactic cosmic
rays, production in the atmosphere by solar flare accelerated particles, and accretion from the sun. Thisnatural production of tritium is estimated to occur at a rate of about 0.12 to 2.0 tritium atoms per squarecentimetre of earth surface per second, with the most probable values being close to 0.2 to 1.0 tritium
atoms per square centimetre per second [J1, N2].
19. Tritium produced by natural processes is rapidly converted into HTO, which then joins the water
cycle. Its concentration in continental surface water and throughout the oceans is about 400 Bq/m3 and100 Bq/m3, respectively. Humans, on average, ingest about 500 Bq of tritium each year, with aresulting average annual effective dose of about 0.01 μSv [U11].

B. Artificial sources
1.

Nuclear weapon tests

20. In the mid-1950s and early 1960s, tritium was widely dispersed during the above-ground testingof nuclear weapons. Especially large quantities of tritium in elemental form and as tritium oxide werereleased in the environment in a series of hydrogen-bomb tests that started in 1952; their total explosive
fusion yield was 328 Mt. The total amount of tritium released into the atmosphere from the testing of
nuclear weapons from 1945 to 1980 was estimated to be 186,000 PBq [U11]. The quantity of tritium inthe atmosphere from weapon testing peaked in 1963 and has since been decreasing.
21. Tritium is readily recycled in the biosphere and becomes homogenously disseminated in thehemisphere where it has been released. The International Atomic Energy Agency (IAEA) runs a global
network of 155 stations to measure tritium in precipitation [I1]. Measurements of tritium in drinking
water in the United States in the early 1960s showed concentrations over two orders of magnitude
higher than background levels that decreased with a half-time of about three years (figure I).

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249

Figure I. Environmental tritium in surface water (pCi/L) in the United States in 1951–1975 [B15]1 pCi/L = 0.037 Bq/L

2.

Production of tritium

22. In countries with developed nuclear technologies, tritium is produced in large quantities formilitary and peaceful purposes by means of irradiation with neutrons of lithium enriched with isotope
6
Li at industrial nuclear reactors. Tritium can be released into the environment from operational tritium
production facilities in the form of elemental hydrogen or tritium oxide with high specific activity.
23. Elevated levels of HTO were measured in lakes located in the area of the Mayak facility(Ozyorsk, Russian Federation) in 1982, 1986 and 2001–2003 by Chebotina and Nikolin [C12] and in
2009–2012 by Kazachenok et al. [K5, K6]. In the former measurement series, tritium concentration inlake water in 2001–2003 was inversely proportional to the distance from the Mayak facility (figure II).
During the observation period 1982–2003, the HTO concentration in lake water decreased by a factorof 2 to 16 while during the same period its mean concentration in major Russian rivers went down by afactor of about 3, from 8 to 3 Bq/L [S20].

250 UNSCEAR 2016 REPORT

Figure II. Dependence of tritium concentration in lake water (Bq/L) in 2001–2003 on the distance (km)from the Mayak facility [C12]3.

Operation of nuclear facilities

24. Tritium is produced in nuclear reactors by ternary fission, one triton per around 1 in 104 fissionsof 235U induced by thermal energy neutrons and by neutron reaction with light elements such as boron,
lithium and hydrogen (deuterium) [N2]. Tritium is produced in much larger quantities in heavy-watermoderated nuclear reactors through neutron capture by deuterium atoms.
25. Tritium is released into the environment from nuclear reactors, especially heavy water reactors,
and spent fuel reprocessing plants, including waste storage and waste disposal sites. In the future, there
is potential for significant releases during the operation of fusion reactors. Tritium is released
predominantly as HTO or elemental hydrogen, partially converted by environmental biota to OBT.
From 1998 to 2002, the global annual average releases of tritium to the atmosphere and to the aqueous
environment from nuclear facilities were estimated to be 11.7 PBq and 16.0 PBq, respectively. The
resultant average annual collective effective doses from these releases were estimated to be 25 and10.5 person-Sv, respectively [I2, I3, U11].
26. In the vicinity of nuclear installations, especially near heavy water reactors, tritium activity inenvironmental compartments can be above background values. For example, while tritium (HTO)
activity concentrations in air at background locations in Ontario, Canada, range from 0.01 to
0.08 Bq/m3, tritium in the vicinity of CANDU nuclear power plants (NPPs) range from 0.05 to
31 Bq/m3. Fish caught in the vicinity of NPP effluent discharges have HTO activity concentrations up
to 50 Bq/L while in fish from background locations, it was less than 9 Bq/L [C23].

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

4.

251

Incidental releases from nuclear facilities

27. Large incidental releases of tritium from tritium production facilities have been reported to occur
from Lawrence Livermore Laboratory, United States in 1970 and from Savannah River Plant, UnitedStates in 1974–1984 shown in table 1 [O2]. The released activity decreased with time from 11 to
18 PBq in early 1970s to 0.3 PBq in 1984. The chemical forms of the released tritium werepredominantly elemental hydrogen (gas) or tritium oxide or their mixture. Monitoring has shown thatelemental tritium was gradually converted in the environment to tritium oxide.
Table 1. Large incidental releases of tritium in the United States [O2]Lawrence Livermore Laboratory (LLL) and Savannah River Plant (SRP)

Site

Year

Tritium release (PBq)

HTO (%)

LLL

1970

11

<1

SRP

1974

18

<1

SRP

1975

6.7

0.6

SRP

1981

1.2

>99

SRP

1983

2.1

1

SRP

1984

0.3

70

28. Elevated levels of tritium in the environment were also observed following major nuclear
accidents—the Chernobyl accident in the USSR in 1986 and the Fukushima accident in Japan in 2011.
In May 1986, tritium concentrations in precipitation collected in the Ukraine and the European part of
the Russian Federation had increased by a factor of two–three compared with 1985, as had tritium
concentration in river water [S20].
29. After the Fukushima accident, tritium concentrations in precipitation collected 170–700 km
southwest from the Fukushima Daiichi Nuclear Power Station and in plant water collected in its
vicinity were substantially elevated (up to a factor of a few tens) compared with pre-accident levels[K2, M4]. According to Povinec et al. [P11], the amount of tritium released and deposited over thenorth-west Pacific Ocean was in the range of 0.1–0.5 PBq.

5.

Other tritium-bearing facilities and commodities

30. Elemental tritium and tritiated luminous compounds are widely used in the luminizing industry,e.g. for illuminating watch and compass dials and as permanent warning lights. Metal plates withincorporated tritium are used in nuclear physics as targets for nuclear reactions, e.g. neutron production.
Other metal plates with incorporated tritium are used as sources of air ionization in industry andagriculture. Tritiated biochemical substrates are produced at radiopharmaceutical facilities and thenapplied for diagnostic health examinations in hospitals, and research activities in medicine and biology.
31. Tritium used in such industrial, health care, research and other applications is partially released
into the working environment, human habitat and natural environment and becomes incorporated in thebodies of workers and members of general public in various physical and chemical forms.

252 UNSCEAR 2016 REPORTIII. PHYSICAL, RADIOLOGICAL AND BIOCHEMICAL CHARACTERISTICS

A. Physical characteristics
32. Tritium (3H or T) is the heaviest radioactive isotope of hydrogen. The tritium atom has one proton
and two neutrons in its nucleus and one electron. The binding energy of nucleons is 8.4 MeV, and thediameter of a tritium atom is 1.1 Angstroms. The dissociation energy, T2 to 2T, is 4.59 eV; ionizationenergy, T to T+ e−, is 13.55 eV.
33. Tritium’s physical properties are similar to those of common hydrogen (1H), which dominates innature over tritium and the intermediate by mass, stable deuterium (2H or D). Under ambientconditions, tritium is a colourless highly flammable diatomic gas with the molecular formula T2. It is
possible to make liquid tritium at atmospheric pressure by cooling it to below 25 K (−248 °C). Liquidhydrogen can be stored in insulated containers under pressure.
34. Tritium has a high coefficient of diffusion. It readily diffuses through porous substances such as
rubber and can also diffuse through metal. Tritium, like common hydrogen, easily undergoes various
chemical reactions depending on physical and chemical conditions. The prevailing form of tritium innature is tritium oxide (T2O) or HTO.
35. Tritium figures prominently in studies of nuclear fusion because of its favourable reaction crosssection and the large amount of energy (17.6 MeV) produced through its reaction with deuterium:
D + T = 4He2 + n + 17.6 MeV

B. Radiological characteristics
36. The nucleus of the tritium atom is unstable and decays with the emission of a beta particle and an
antineutrino to stable 3He. The antineutrino is of no biological significance because it does not interact
with matter:H → 3He+ + e‒ + 𝑣𝑣̅ e

3

37. Tritium has a physical half-life of 12.3 years and, in the pure elemental state, a specific activity of
3.56 × 1014 Bq/g. Emitted beta particles are very low energy, mean 5.7 keV (90 fJ) and maximum
18.6 keV (300 fJ).
38. In water, the average track length of the beta particle is 0.56 µm and the maximum track length is
6 µm, which compares with a typical cell nucleus diameter of 6–15 μm, while a cell has a diameter of
10–100 μm [V1]. Tritium beta particles are completely absorbed by sheets of plastic, glass or metal.
They do not penetrate dead layers of skin. However, following intake of tritium, beta radiation canirradiate internal organs. Within the body it gives a relatively low absorbed dose per disintegrationcompared with other beta-emitting radionuclides, but the ionization density of the electron is greater.

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

253

39. Radioactive decay of tritium atom also results in transfer of some recoil energy to a daughter 3He+
positively charged ion. This energy depends on the random dispersion angle between the emitted
electron and antineutrino and comprises 1.0 eV as average and 3.3 eV as maximum [F2, G8]. This
energy is insufficient for either daughter atom self-ionization (required energy of the order 10 keV) or
tissue ionization (about 30 eV). Besides the recoil energy, the daughter 3He+ ion also carries excitationenergy of about 11 eV that can influence the fate of the molecule to which the tritium atom was boundand result in its chemical transmutation and modification of its chemical properties.

C. Biochemical characteristics
1.

Tritiated water

40. Tritium is most commonly found in natural and working environments in the form of HTO, which
has the same chemical properties as ordinary water. Water with a tritium activity of 1 Bq/L contains
less than one tritium atom among 1017 molecules of water. HTO can enter the human body byinhalation, skin absorption (liquid and vapour) [D3, P9], or ingestion of water or food [B14, I13]. Onceinside the body, HTO diffuses freely and rapidly across cellular membranes, and reaches equilibrium
throughout the total body water pool [H12]. HTO is excreted via urine, faeces, sweat, and breath [N1].
Since HTO quickly reaches equilibrium with the water in the body and is distributed uniformly among
all soft tissues, the concentrations of HTO in sweat, sputum, urine, blood, perspiration, and exhaledwater vapour are considered to be equal [H12].

2.

Tritiated gases

41. Tritiated elemental hydrogen (HT or T2) is relatively inert in biological systems and has a very
low uptake into body fluids and tissues [H12]. Humans are mostly exposed to HT by inhalation or skincontact with contaminated surfaces. A small fraction of inhaled HT is converted to HTO in the humanbody. The primary sources of HT are tritium production and processing facilities (such as those
involved in making gaseous tritium light sources), tritium recovery facilities, and nuclear fuelreprocessing facilities. HT is readily converted to HTO in the environment, with soil microorganisms
playing an important part in this process [A3].
42. Tritiated methane (CH3T) is relatively inert in biological systems. Humans can be exposed to
CH3T by inhalation in workplaces or in the public domain following biochemical degradation of OBT
in the environment. Because of the low solubility of methane in body fluids, the radiologicalimplications of inhaling CH3T are mostly determined by its oxidation to HTO and biochemical
conversion to OBT in the human body [P5].

254 UNSCEAR 2016 REPORT

3.

Organically bound tritium

43. Because tritium atoms are exchangeable with normal hydrogen atoms, a fraction of the tritium
absorbed by plants or animals can become incorporated into organic compounds such as carbohydrates,
fats, proteins, and collagen: this is referred to as OBT 1. Animals, including humans, ingest OBT andform OBT from HTO within their tissues [D5, K14].
44. A tritium atom in OBT attached to a carbon atom is essentially fixed until the compound ismetabolized (i.e. the tritium is non-exchangeable). However, a tritium atom attached to an oxygen,sulphur, nitrogen or phosphorus atom is readily exchangeable with hydrogen in water and is notconsidered as OBT in this annex [D5, R11, S1] or specifically qualified as exchangeable OBT [K14].
OBT exhibits longer retention times in the body than HTO.

4.

Tritiated organic substances

45. A broad spectrum of organic substances labelled with tritium, including biochemical substrates,
are produced and used widely in research and for other purposes. Workers may be exposed byinhalation or through skin contamination, and also by inadvertent ingestion. In the human body,labelled biochemical substrates (e.g. amino acids, DNA precursors, glucose, hormones) may be
metabolized with partial loss of tritium label converted to HTO, or be incorporated into biologicalmacromolecules as OBT [B11, H12]. Foreign organic compounds (such as organic solvents) are usually
rapidly excreted from the body in urine and faeces [B8].
46. Labelled DNA precursors (e.g. 3H-thymidine, 3H-deoxycytidine) belong to a special group that, inthe mammalian body, are partially degraded to HTO and partially incorporated into the DNA of dividing
cells, and thereafter selectively expose the nuclei of proliferating cells to beta radiation [D5, N1].

5.

Metal tritides and other low soluble forms of tritium

47. Tritiated compounds with low solubility, which are widely produced and used in industrial or
research facilities, include luminous compounds (as powders), particles of metal tritides that are used as
accelerator targets or as ionization sources in industry and agriculture, and carbon, beryllium andtungsten dust, micro-fragments of glass contaminated with tritium used in fusion experiments. Airborne
particles with insoluble tritium are inhaled by workers and, depending on their dimensions andrespirability, deposited in the respiratory tract. Following inhalation, such substances can remain in thepulmonary region of the lungs and expose tissues to both beta radiation and, to a lesser extent, tobremsstrahlung.
48. Although insoluble particles are largely retained in lung tissues, transported by macrophages toregional lymph nodes, or escalated from the lungs by mucociliary clearance, some dissolution will occur
and a proportion of their tritium content will be removed and absorbed to blood as HTO. When luminous
powder produced from zinc sulphide granules coated with a thin layer of high activity tritiated polymer isinhaled, some tritium is detached from the polymer and absorbed to blood as low molecular organic
foreign compounds that are rapidly excreted from the body in urine and faeces [B8, B11, B14].
1

OBT in biological tissues is carbon-bound tritium that was originally formed in living systems through natural environmental or
biological processes from HTO (or HT via HTO). OBT is not exchangeable with hydrogen in water.

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

255

IV. HUMAN EXPOSURE
A. Exposure of the public
1. Tritiated water in global water cycle
49. The Committee [U9] has estimated worldwide average annual individual effective doses mainly
from data on ingestion of the globally dispersed HTO created as a result of fallout from above-groundnuclear weapon testing (figure III) [B25, U9]. The doses received from the inhalation of 3H are
negligible in comparison with those received from ingestion. The derivation of these doses is largely
based on environmental measurements and is described by the Committee in its UNSCEAR 2000
Report [U9], Bouville et al. [B25] and Bennett [B16]. The background concentration of tritium inhumans is calculated from an average of the concentrations in the sources of water ingested, assumed tobe 33% from the atmosphere, 53% from fresh water, 13% from groundwater and 0.7% from ocean
surface water (through fish) [B25, N2].
50. The largest annual doses from tritium in fallout were received by the world population during theperiod of intense nuclear weapon testing during the late 1950s and early 1960s, before the Limited Test
Ban Treaty of 1963. The peak global average annual effective dose from tritium in fallout was 7.2 µSv in
1962. Since the majority of atmospheric nuclear weapon tests during that period took place in the
Northern Hemisphere, average doses from tritium were greater in the Northern Hemisphere than in theSouthern Hemisphere. Generally, tritium follows the global water cycle: a large proportion is transferredto the oceans within a few years of production and a very small fraction is ingested by humans [U9].
Figure III. Worldwide average annual individual effective dose in 1950s–1990s from the ingestionof tritium produced in atmospheric nuclear weapon testing [B25, U9]
Dashed line is for dose from natural tritium256 UNSCEAR 2016 REPORT

2.

Local exposure of the public from nuclear facilities

51. Tritium released from nuclear facilities, especially from those operating with large amounts oftritium (i.e. tritium production facilities, heavy water reactors or reprocessing plants) may enter thebodies of people residing in the vicinity in drinking water or via inhalation. In these conditions, humandoses caused by local intake of environmental tritium are usually larger than those caused by global
tritium levels common in neighbouring areas with no facilities releasing tritium.
52. Measurements of tritium (HTO and OBT) in environmental media are carried out routinely in thevicinity of Canadian nuclear facilities. These measurements allow estimation of doses to members of
the public from all exposure pathways (e.g. inhalation, skin absorption, ingestion of food and drinking
water). In 2006, annual tritium doses to members of the public living in the vicinity of NPPs were lessthan 2.5 µSv whereas they were slightly higher in the vicinity of two facilities manufacturing gaseous
tritium light sources (GTLS). The annual tritium doses of respective critical groups for the two GTLS
facilities were 15 and 67 µSv [C23].
53. Kim and Han [K13] studied environmental radiation conditions in 1992–1993 around the
Wolsong NPP, Republic of Korea, for a CANDU-6 heavy water reactor that had been operational since1983. Activity concentrations of HTO and OBT were analysed in food samples collected within
1-15 km of the reactor site and HTO was collected from the air in some locations. In water samples
extracted from rice, Chinese cabbage, radish and pumpkin, HTO concentration was in the range of
3-100 Bq/L and that in combustion water obtained from organic parts of vegetables was 4−130 Bq/L,
both inversely proportional to the distance from the site. The ratio of tritium concentration per unithydrogen mass in OBT to that in free HTO was in the range of 1.0–2.8, with an average of 1.35. On thebasis of monitoring data, the authors assessed the annual effective dose for adult members of the public
to be in the range of 1.3 µSv in a radius of 0–1.6 km to 0.15 µSv at 8–16 km. Although both values aremuch lower than the background dose, they are substantially higher than the annual dose fromenvironmental tritium to the Korean public residing away from NPPs, assessed by Yoon et al. [Y7]from urine samples of 50 persons to be about 2 nSv.
54. In 2008, Chebotina and Nikolin [C13] measured elevated tritium concentrations in the urine of
45 residents of five towns located in the vicinity of the Mayak facility (figure IV). The average
concentrations, in the range of 100–800 Bq/L, correlated inversely with the distance between the town
and the Mayak facility. Those values are much higher than tritium concentrations in potable watermeasured in the area in the 2000s, indicating inhalation as a possible intake pathway. The measured
concentrations of tritium in urine correspond to average annual effective doses incurred by the residents
of five towns in 2008 in the range of 3–14 µSv.

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

257

Figure IV. Average concentration of tritium in 2008 in urine of residents of towns located in thearea of the Mayak facility [C13]3.

Organically bound tritium in human tissue

55. Very few authors have directly measured tritium content in human tissue. Bogen and Welford[B21] for example have summarized results of tritium measurements in the United States’ environment
carried out in 1960s to early 1970s in non-equilibrium conditions caused by termination of nuclear
weapon tests in 1963 and continued radioactive fallout (decreasing with time) from the stratosphere.
They sampled water vapour from the air, tap water, soil, vegetation, food, animal and human tissue andthey measured HTO in water distilled from the samples and OBT in water from combustion of driedsamples. Both sets of data were presented in terms of HTO activity concentration in water. In all linksof the ecological chain up until 1973, the tritium specific activity in the OBT fraction was higher thanthat found in the free water fraction. The ratio of specific activities OBT/HTO decreased by a factor of
about 1.5–2 from each trophic level: soil 6–8, vegetation 3–4, animal 2–3 and human 1.5–2. Thosepatterns can be interpreted by slower clearance of OBT compared to HTO from various organisms ofthe trophic chain and their residues (soil organic matter) in conditions of decreasing HTO concentrationin the environment, as was the case in the 1970s.
56. Ujeno et al. [U3] measured tritium in tissue water distilled from samples of human organs and tissues(brain, lung, liver, kidney and muscle) collected by staff of Kyoto University during forensic autopsy of eightdead bodies. The tritium concentration in water from various organs and tissues was similar and not affected258 UNSCEAR 2016 REPORT

by sex or age. The average HTO concentration in tissue water was 2.5±0.7 Bq/L, which was similar to thatmeasured in tap water, rain water and water distilled from local food.
57. Hisamatsu et al. [H13, H14] presented the results of tritium measurements in organs and tissues
collected from 11 human cadavers (10 males, one female, mean age±SD=46±16 year), who diedsuddenly in 1986 at Akita Prefecture in northern Japan. They measured tritium concentrations in freewater (HTO) distilled from human samples (brain, liver, lung, heart, kidney, blood serum and whole
blood) and in combustion water obtained by combustion of dried samples in oxygen atmosphere. The
mean tritium concentrations in free water from seven diverse organs and tissues were similar, with arange of 1.5–1.9 Bq/L and average of 1.6 Bq/L. Mean tritium concentration in combustion water ofvarious organs/tissues varied in the same range, average 1.7 Bq/L. The ratio of OBT/HTO tritium
concentrations in human tissue varied between 0.95 and 1.3 with an average of 1.1. Tritium
concentrations in local food sampled in 1985–1987 were also reported. Free water concentration in six
samples of the total human diet varied between 1.4 and 2.2 Bq/L and combustion water varied in therange of 1.7–2.2 Bq/L. Their ratio varied between 0.9 and 1.6 with an average value of 1.2. Thus, inequilibrium conditions of low-level intake of environmental tritium, no differences in tritium
concentrations were revealed between various human organs and tissues, between diet and humantissue, and between free water tritium and OBT.

4.

Measurements of environmental tritium

58. The measurement of environmental tritium in its various forms as gases or vapours (HT, HTO,
organic molecules), liquids (HTO or OBT in solution) and solids (OBT, hydrides) is a key step for dose
assessment and evaluation of health and environmental risks. Sampling, storage and treatment areimportant points in the analytical procedure for tritium. The final form of tritium for analysis is usuallywater, and low concentrations of tritium in water (few Bq/L) are currently measured either by gas
proportional counter or by liquid scintillation counter. They can also be determined indirectly using asensitive mass spectrometer, measuring the amount of the decay product, helium-3, formed in a water
sample in a closed vessel during a given period [W8].
59. Reference water that is virtually tritium-free is used as calibration blanks for the analytical system
and a recent comparison of these water sources gave results ranging from 0.004 to 0.17 Bq/L [F9].
Analytical procedures have been developed to measure OBT [B2] and respective standards are underdevelopment. Recent studies of the speciation of tritium as OBT are investigating variations in thehydrogen content of different forms and identifying compounds solubilized in the samples during labile
exchange [B1].

B. Occupational exposure
60. In working environments, tritium is present in various physical and chemical forms depending on
the production processes. Working environments may contain tritium in a variety of different chemical
forms, including HTO, elemental hydrogen, organic solvents, airborne particles of metal tritides (e.g.
Ti, Zr, Er), tritium contaminated glass and dust particles, luminous compounds, and labelledbiochemical substrates (e.g. amino acids, DNA precursors, glucose, hormones) [B11, H16, I2, I3, J1].
61. Occupational exposure to tritium is usually low relative to other sources of exposure. For example,
in 2006 average annual tritium doses for NPP workers in Canada ranged between 0.07 and 0.26 mSv,
which represented between 14 and 29% of the total effective dose. In the same year, workers employed at
two Canadian GTLS manufacturing facilities had annual tritium doses of 0.19 and 0.3 mSv [C23].

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

259

V. BIOKINETICS AND DOSIMETRY
62. Biokinetic models describe the time-dependent deposition and translocation of radionuclides inthe body and the rates at which they are removed from the body. Biokinetic models are used to
calculate the number of nuclear transformations of radionuclides in each source organ during a
specified period following an intake. Dosimetric models are then used to calculate the absorbed doses
to specific organs and tissues (referred to as target organs) per nuclear transformation of radionuclides
in each source organ (i.e. each site of radionuclide deposition or transit in the body).
63. For protection purposes, ICRP calculates values of committed effective dose as a doubly weighted
sum of organ and tissue doses—first, adjustment to take account of the relative effectiveness ofdifferent radiation types in causing stochastic effects using radiation weighting factors (wR), andsecond, adjustment for differences between organs and tissues in their contribution to total detrimentfrom stochastic effects using tissue weighting factors (wT) [I23].
64. Because the range of the beta radiation emitted by tritium is short, all the radiation energy is
generally assumed to be absorbed in the tissues and organs in which the tritium decays. Therefore,
organ or tissue dose from tritium radiation is entirely determined by a relevant biokinetic model and notby radiation transport in the body. The biokinetic model for tritium depends on the type of tritiatedcompound taken into the body, as this dictates its deposition, translocation, retention and excretion.
65. In animal studies, the absorbed dose in tissue resulting from an acute administration of tritium canbe calculated using information on the initial activity concentration of tritium in the tissue and the rateof removal of tritium from that tissue arising from biological processes and radioactive decay. Thissections present the basis for the biokinetic models used for the intake of tritiated compounds byinhalation, ingestion or absorption through the skin. Examples of dose coefficients (committed effective
dose per unit activity taken into the body) for tritium calculated with the various biokinetic models areprovided.

A. Information on biokinetics and dosimetry of tritiated compounds
1.

Tritiated water(a) Early biokinetics of HTO in mammals
66. Tritiated water can enter the human body via ingestion of food and drink and—mostly foroccupational exposure—by inhalation of HTO vapour or direct absorption through skin exposed towater or water vapour. Following ingestion, absorption from the alimentary tract into the bloodstream iscomplete within a time range from a few minutes to some tens of minutes. Following inhalation, almost
the entire amount of inhaled HTO vapour is absorbed very rapidly from the respiratory tract into thebloodstream [B7, P9]; absorption through skin provides an additional common route of entry into thebloodstream [D3, O6, P9].
67. Following uptake to blood from the alimentary or respiratory tracts or through the skin, HTO istransported by the circulatory system to all the body organs and tissues and diluted uniformly in body260 UNSCEAR 2016 REPORT

water. This process takes from a few hours to some tens of hours. A small fraction of tritium from HTO(0.5 to 4%) exchanges very rapidly with hydrogen in organic molecules as OH, NH and SH bonds
throughout body tissues. Another small fraction (from less than 1 to 3% in humans) is gradually converted
to OBT as a result of biochemical processes, i.e. CH bonds in organic molecules [B8, H12, P9].
68. The absorption of HTO through the skin from either the vapour or liquid phase has been
investigated by several authors. DeLong et al. [D3] exposed mice, rats and human adult volunteers to
HTO vapour in air. Animals were sacrificed following exposure. Urine and blood samples werecollected from human subjects for 48 hours after the end of exposure. Absorption rates of HTO,
calculated from measurements of tritium in blood and total body water, suggested that a delay occurredin the distribution of the absorbed HTO. The absorption rate of HTO through the skin was the same
whether the skin was covered with a cloth (cotton) or uncovered. It was also proportional to the water
vapour pressure. This suggested a single diffusion mechanism for percutaneous absorption. However,
the absorption rate for the vapour phase was larger than could be accounted for by diffusion due tovapour pressure alone, perhaps as a result of capillary action. The absorption rate increased withincreasing skin temperature. DeLong et al. [D3] and Pinson and Langham [P9] concluded similarly thatthe quantity of HTO entering the body through the total skin, when exposed to an atmosphere
containing a given activity per unit volume, would be about equal to that entering through the lungs.
Osborne [O6] exposed volunteers to HTO in air and measured the tritium activity in urine, showing acorrelation between skin absorption rate and skin temperature.

(b) Long term biokinetics of HTO in mammals
69. Several studies have examined the biological half-time of HTO in a total of about 400 adults bymeasuring tritium activity concentrations in urine. Butler and Leroy [B28] found this parameter to varywith the amount of water ingested (decreasing with increasing water intake rate), with the ambient
temperature (decreasing with increasing ambient temperature) and with age (decreasing with increasing
age in adults). Their study, based on 310 cases of HTO intake, showed that the biological half-time of
HTO varied from about 4 to 18 days, with a mean of 9.5 days. During the warmer months, the averagehalf-time was lower; the difference being attributed to increased water intake. Other studies, based on
fewer cases, showed similar results [H12]: from 6 days for 8 cases [R16], to 12 days for 5 cases [B7].
In a few cases of accidental intake of large amounts of HTO, the excretion rate was accelerated with
diuretics and increased intake of fluids (e.g. [S2, T14]). The ICRP model uses a biological half-time of
10 days for HTO and 40 days for OBT formed from HTO in the body of adults [I8, I9, I10, I14, I17].
70. A number of studies have reported evidence for the presence of a second exponential componentof tritium activity concentration in human urine associated with the formation of OBT from HTO andits subsequent removal. The biological half-time for OBT removal mostly ranged from 23 to 104 days
(mean 59 days) and its contribution to total excretion in 17 study subjects varied between 0.01 and
0.7% (mean 0.2%) [B7, H9, L4, S2, S17, T14]. A value of 40 days, based on carbon turnover in thebody, was adopted by ICRP [I13]. This was derived from the ratio of reference values for the bodycontent of carbon (16 kg) and daily carbon intake (0.3 kg) for adults [I7].
71. Some studies have reported an even longer component for the removal of tritium from the body(e.g. [M15, M16, S2]). However, the parameters of this component as derived from data obtained on
five subjects occupationally exposed to either HTO or tritiated luminous compounds are very uncertain.
Such a component would contribute in only a minor way to tissue doses, as it represents less than 1% of
total OBT. The biological half-times of tritium following acute intake of HTO by adults reported byvarious authors are shown in table 2.

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Table 2. Biological half-times of tritium in humans following acute HTO intake
Numberof cases

Study

Compartment 1(body water)

Compartment 2(organically bound)Compartment 3(organically bound)

Pinson and Langham [P9]

9

11.3

—

—

Foy and Schnieden [F10]

10

5–11
(mean: 7.5)——

Wylie et al. [W9]7

6.4–12.1(mean: 8.5)——

Butler and Leroy [B28]

310

4–18
(mean: 9.5)——

Osborne [O6]30

6.4–14.4(mean: 10.5)——

Snyder et al. [S17]

1

8.7

34

—

Sanders and Reinig [S2]

1

6.1

23

344

Minder [M15]

1

—

10–30

139 to 230

Lambert et al. [L4]

1

9.1b

36

—

Moghissi et al. [M16]3—21 and 26

280, 550±140,
350±190

Henry [H9]

1

7.5

63

—

Balonov et al. [B7]

5

11–13

39 to 76—c

Trivedi et al. [T14]aBiological half-time (days)

8a(mean: 11.9±0.3)(mean: 51±7)

6.2–12.8(mean: 8.4)

58–104(mean: 74±18)d—Oral diuretic administered from day 3–35 post-intake.

b

HT/HTO acute intake.

cData for three tritium luminous dial painters collected 6–10 months after termination of employment.

dDuring the initial period when the exposed individuals increased fluid intake one month post-intake, the biological half-time
varied from 5.0 to 8.1 days with a mean of 6.3 days.

72. The partitioning of HTO and OBT after intake of HTO was examined by Takeda and Kasida [T1]
as part of a study of the biokinetics of HTO in rats. These investigators found that “initially, the ratio oftissue-bound tritium to total tritium was about 3% in the kidney and 1–5% in other tissues.”
Measurements of tritium in human tissue are generally unavailable, but information can be derived bybiophysical modelling using long-term measurements of tritium in urine or, in some cases, blood
samples. The published data from the six studies [B7, H9, L4, S2, S17, T14] are available for suchanalysis (table 2). Additional information can be derived from tritium measurements for organic
components of urine (urea) or blood (proteins, dehydrated cells). However, the number of reliablemeasurements is limited [L4, R16, T16] and their representativeness with regard to tritium content inorgans and tissues is questionable.
73. The long-term retention of tritium in the human body as shown in figure V was modelled byseveral authors as linear recurrent two-compartment models with transfer of tritium from HTO to OBT(synthesis of biomolecules with OBT) and subsequent catabolic loss (degradation of biomolecules withOBT) with all tritium excretion as HTO [B7, S17, T14]. Using this approach, the peak OBT activity

262 UNSCEAR 2016 REPORT

was calculated by Balonov and Chipiga [B6] from the data of the six studies referred to above [B7, H9,
L4, S2, S17, T14]. The peak OBT activity was estimated to be reached in 16–38 days (mean 26±5 days)
after a single HTO intake and accounted for 0.1–1% (mean 0.4±0.2%) of the initial intake of tritium.
Balonov and Chipiga estimated the contribution of OBT to total soft tissue doses after intake of HTO
using 17 available data sets and obtained values of 1.8–4.6% (mean 3.0±0.9%) [B6]. It is notable thatthe contribution of OBT to the total dose was estimated to be similar in nine subjects with relatively
low previous occupational tritium intake [H9, L4, S2, S17] or in five volunteers with no previous
tritium intake [B7] (3.0±0.9%) and in eight workers at a Canadian heavy water reactor (3.0±0.8%),
indicating that the contribution of previous chronic HTO intake by the workers was insignificant.
Figure V. Schematic two-compartment recurrent model of tritium biokinetics following HTO intake
in the body [B7, S17]
I(t) is HTO intake rate (Bq/day); W(t) is tritium activity in the body as HTO (Bq); B(t) is tritium activity in the body as OBT (Bq);
U(t) is tritium activity excreted from the body as HTO (Bq): α, β and γ are transfer rate constants (day-1)

74. Trivedi et al. [T14] assessed the contribution of OBT to effective dose following acute intake of
HTO by calculations based implicitly on a simpler linear two-compartment model with no transfer oftritium from HTO to OBT. Instead, it was assumed that some fraction of HTO taken in was instantly
converted to OBT that was gradually degraded to and excreted as HTO (figure VI). The contribution of
OBT to the effective dose assessed with this model was 3–9% (mean 5.3±2.1%) for 15 subjects [B7,H9, I14, S17, T14] excluding data from Rudran [R16] obtained on workers occupationally exposed totritiated luminous compounds. In another study by Trivedi et al. [T16] the tritium concentrations inurine and blood samples both as HTO and OBT for six workers chronically exposed to low levels oftritium were measured. The activity concentration of OBT per gram of hydrogen in OBT from urine
and blood samples (their ratio was 0.95±0.25) was assumed to be equal to that in body tissues. Bymeans of a simple equilibrium model, they calculated the contribution of OBT to the total dose to beequal to 4.7–9.9% (mean 6.9±3.1%) for the six workers. This is in good agreement with the previous
results from 15 subjects with acute intake of HTO.
75. The estimate of Trivedi et al. for the contribution of OBT to dose following acute intake of HTO(figure VI) is 1.8 times greater than that obtained by Balonov and Chipiga using a more physiologicallyrealistic model (figure V); it does not account for the gradual bioaccumulation of OBT and assumes its
instant formation from HTO. A similar conservative approach is used by the ICRP model for HTO (seealso figure VIII) which is applied for radiation protection purposes [I14]. According to this model, 97%
of HTO absorbed to blood is distributed in body water (T1/2=10 days in adults) and 3% is instantly

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263

converted to OBT (T1/2=40 days in adults); the contribution of OBT to the effective dose in the ICRP
model is about 9% under assumption of uniform OBT distribution in organs and tissues.
Figure VI. Schematic two-compartment model of tritium biokinetics following HTO intake in the
body [T14]
I(t) is HTO intake rate (Bq/day); x is fraction of tritium instantly converted to OBT (dimensionless); W(t) is tritium activity in the
body as HTO (Bq); B(t) is tritium activity in the body as OBT (Bq); U(t) is tritium activity excreted from the body as HTO (Bq):β and γ are transfer rate constants (day−1)

76. Other authors [H12, J3, N1, S2, T9] have used the limited available human data to develop athree-compartment model of tritium retention following intake of HTO. The results showed
considerable variation, both in the observed biological half-times and in the proportions of tritium
entering OBT compartments. The model parameters suggested by Taylor [T9] for adults are shown intable 3. The resulting committed effective dose per unit intake of HTO by adults, on the basis of thismodel, is 1.7 × 10-11 Sv/Bq. The current ICRP dose coefficient for HTO is 1.8 × 10-11 Sv/Bq [I15].
Table 3. Parameter values for HTO model proposed by Taylor [T9]Model component

Distribution (%)

Biological half-time (days)

HTO

99

10

OBTa

0.98

40

OBT

0.02

350

b

a

Short-term OBT compartment.

bLong-term OBT compartment.

264 UNSCEAR 2016 REPORT

2.

Inhalation and skin absorption of elemental tritium gas

77. Tritiated elemental hydrogen (HT) is only slightly soluble in body fluids and has a much lower
uptake into biological systems than HTO. After inhalation, most of the HT is exhaled, but a small
fraction is dissolved in body fluids and is then oxidized to HTO by anaerobic bacteria in thegastrointestinal tract, the only known biological site of HT oxidation [I6].
78. The ICRP based its assessment of the effective dose resulting from the inhalation of HT primarilyon exposure of the lungs, as opposed to exposure of the skin. Oxidation to HTO in vivo was notconsidered [I9]. The absorption of HT through the skin appears to be negligible and it does not convert
to HTO on contact with the skin [H12].
79. The studies of Pinson and Langham [P9] and Peterman et al. [P3, P4] found that exposure to HT
resulted in excretion of HTO in urine, and that the HTO, formed from the oxidation of HT, was retainedin the body and excreted with the usual biological half-time of about 10 days. About 0.01% of the HT
inhaled by human volunteers was converted to HTO in the body [P3, P4, P9].
80. Using their biokinetic models for HT and HTO and data from studies with human volunteers,Peterman et al. [P3] concluded that the effective dose from inhaled HT was dominated by two roughly
equal contributions: the effective dose resulting from exposure of the lungs to HT in inhaled air, and theeffective dose resulting from exposure to HTO caused by the oxidation of HT. The current dosecoefficient for inhalation of HT given by the ICRP Publication 68 [I15] is based on the work ofPeterman et al. [P4] and its human respiratory tract model.
81. Most of the energy of the tritium beta particles is not deposited in the target cells of therespiratory tract due to their short range. The average depth of the nuclei of these cells range from about
10 to 50 μm in the extrathoracic, bronchial and bronchiolar regions of the respiratory tract [I16]. In thecurrent ICRP model of the human respiratory tract [I16], the tritium beta particles are assumed todeliver a dose only in the alveolar-interstitial region after inhalation of HT. Trivedi and Gentner [T15]
noted that the nuclei of target cells within the alveolar-interstitial region, at depths of less than 10 µm,
are assumed to receive some dose.

3.

Contact of skin with tritium-gas-contaminated surfaces

82. Eakins et al. [E1] applied tritium-gas-contaminated metal surfaces to the forearms of fourvolunteers. The estimated average body content of HTO and OBT were about 0.5 and 0.3%,
respectively, of the applied activity of tritium gas. The results did not depend on the type of material
tested and the initial body content of organic tritium was less than that of HTO for each volunteer.
Urinary excretion of tritium, initially primarily OBT, reached a peak about 24 hours after exposure. Up
to 50% of the OBT was excreted via urine with a biological half-time of about one–two days; the rest of
the OBT was excreted with a biological half-time of 0.1–0.2 days. One–three weeks after exposure, thetritium was excreted predominantly in the form of HTO with a biological half-time of about 14 days.
The effective dose resulting from intake of tritium from contact with a contaminated surface wasestimated to be about 9 × 10−12 Sv/Bq [J4].
83. Similarly, Trivedi [T13] used hairless rats and showed that when HT-contaminated stainless steel
was brought in contact with intact skin, tritium was fixed as OBT and HTO in the skin anddemonstrated biphasic excretion of OBT and HTO.

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4.

265

Ingestion of organically bound tritium

84. While HTO diffuses freely in body tissues and enters into equilibrium with body fluids in amatter of minutes or hours [H12, P9], several factors determine the distribution in the body of OBT
ingested through the diet. These include the biochemical composition of OBT as a mixture oftritiated carbohydrates, fats and proteins, the oxidation rates of the dietary constituents, absorption
from the alimentary tract, and the synthesis and retention of organic forms and the HTO and OBT
excretion rates [D5].
85. Experiments indicate that about 3–20 times more tritium is bound to organic compounds in moremetabolically active tissue after ingestion of food containing OBT than that after ingestion of the sameactivity of tritium in the form of HTO [P7, R11, T3, T6]. This ratio depends on the animals studied(rabbits, rats, mice), the kind of tritium-labelled food (e.g. alfalfa, wheat, meat, shrimp) and theduration of tritium intake (single, few weeks, three generations). Takeda and Kasida [T1] found that,
following the intake of HTO by animals, 1–5% of the tritium was incorporated into organic constituentsof rat tissue. Therefore, several tens of per cent of the tritium ingested as OBT would be expected to beincorporated into organic molecules in mammal tissues. The assumption made in the ICRP
Publication 60 [I13] that 50% of tritium is incorporated into OBT in tissues after an intake of OBT
stems from this reasoning.
86. On the basis of all available data, several authors [I12, K15, M1, P6, R11, R12] suggested—as arounded value—that approximately nine times more OBT may be present after intake of OBT than after
intake of HTO. This corresponds to a range of 9−45% for the proportion of tritium reaching blood thatwas retained as OBT, the remainder being converted to HTO (the wide range resulting from thediffering metabolic roles of different OBT molecules, e.g. as an energy source or a structuralcomponent). Further, it was also suggest that the use of a value of 50% would be suitable for generalradiological protection applications.
87. While the biokinetic parameters used by ICRP are supported by a range of data, there have beensuggestions that the contribution to dose from OBT may be greater for intake of either HTO or OBT
than predicted by ICRP models. However, the conclusions of Takeda [T4], Komatsu et al. [K21] andRodgers [R12] are reasonably consistent with the ICRP conclusion that the contribution to dose from
OBT after intake of HTO will be small (<10%) and that the overall dose from intake of OBT will begreater than from HTO by about a factor of two (the data in table 4 show the ratio in ICRP dosecoefficients to be about 2.5).
88. Some experimental data suggest that after chronic intake of HTO, equilibrium tissue concentrationsof HTO and OBT are similar [C25]. Etnier et al. [E3] used a four-compartment model of hydrogenmetabolism to show theoretically that OBT in food can increase the cumulative total body dose by a factor
1.7−4.5 times the free body water dose alone. This model is regarded as providing a reliable representationof tritium biokinetics. The predictions in the model were demonstrated by Takeda et al. [T3] using rats fed
with tritiated wheat and HTO. After chronic ingestion of tritiated wheat (22 days), the amount of OBT in
the tissue of rats exposed to tritiated wheat was about 6–11 times higher than when exposed to HTO.
89. Hunt et al. [H19] measured the retention time of tritium in volunteers who had eaten fish that
contained OBT or HTO as a consequence of discharges from the General Electric Healthcare Ltd. plantin the United Kingdom. The excreta from five volunteers were screened for a period of up to 150 days
after intake. The results suggested biological half-times ranging from four to eleven days for theretention of the total amount of tritium, with no evidence for a statistically significant contribution froma component with a longer half-time.

266 UNSCEAR 2016 REPORT

90. Animal studies have shown a non-uniform distribution of OBT in soft tissue, and also a nonuniform distribution of tritium in the various organic compounds in the body [D5, I13]. However, withthe exception of the special case of DNA precursors, discussed later, any non-uniformity of distribution
within cells appears to be small.
91. Richardson and Dunford [R9] conducted a literature review of the studies of exposure to OBT. They
proposed two biokinetic models governed by the overall metabolic reactions of the principal nutrients:
carbohydrates, fats, and proteins. The parameters for two models of differing complexity—called the
HCNO-S and HCNO-C models—were evaluated on the basis of biochemical reactions. The simpler
model has a single compartment representing the principal nutrients. The more complex model includes
compartments representing the longer-term retention of carbohydrates as glycogen, fats as adipose tissue,
and proteins in bone and soft tissue. The differences in the water and organic contents of tissues andorgans and the different biokinetics for the different organic components were considered [R8, R9].
92. Melintescu et al. [M10] and Galeriu and Melintescu [G1] developed other physiologically based
multicompartment models where OBT exchange rates were associated with energy metabolism ofmajor groups of human organs and tissues. Model predictions for HTO intake were successfully tested
against available human data, e.g. from Trivedi et al. [T14]. Final model solutions were presented asdose coefficients for both HTO and OBT for all ICRP age groups separately for males and females andsex-average values.
93. Table 4 shows values of committed effective dose per unit intake for adults obtained fromRichardson and Dunford [R8], an earlier paper of Richardson et al. [R6], other physiologically basedmodels of Melintescu et al. [M10] and Galeriu and Melintescu [G1] and the corresponding values
calculated using the ICRP models for HTO and OBT. The dose coefficients from the various studies are
similar to the ICRP values [I17]. Richardson and Dunford [R8] also provided dose coefficients for tritiatednutrients, with the lowest value for tritiated carbohydrates equal to 3.3 × 10−11 Sv/Bq, which is 21% lower
than the ICRP value for OBT, and the highest value for tritiated protein being 8.4 × 10−11 Sv/Bq, which is
twice the ICRP value for OBT [I17, R8].
Table 4. Adult effective dose coefficients from physiologically-based biokinetic models and ICRP
models for ingestion of HTO and organically bound tritium
Biokinetic model

Dose coefficient (10−11 Sv/Bq)
Ingestion of HTO

ICRP [I14]1.8Richardson et al. [R6]

1.8 (males); 2.2 (females)

Melintescu et al. [M10]

1.7 (males); 2.4 (females)

Galeriu and Melintescu [G1]2.0

Ingestion of OBTICRP [I14]Richardson et al. [R6]
Richardson and Dunford [R8] and ICRP [I7]
Melintescu et al. [M10]Galeriu and Melintescu [G1]4.2
4.2 (males); 6.1 (females)
5.0–7.4
3.9 (males); 5.7 (females)
4.9

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267

94. Physiological models of OBT biokinetics proposed by Richardson et al. [R6] and Melintescu et al.
[M10] provide sex- and age-differentiated biokinetics and dose coefficients. This analysis was based on
estimates of daily intake rates of carbon in females (228 g carbon) and males (303 g carbon) andcalculated biological half-times of carbon of 51 days and 40 days for female and male adults,
respectively. Table 4 presents the dose coefficients from physiologically-based OBT biokinetic models
by sex. The dose coefficient is higher in females by about 20–40% for HTO and up to 50% for OBT.
Further, the authors that the modelling of age dependence of the dose coefficients both for HTO andOBT is in good agreement with ICRP data [I17, M10].

5.

Intake of tritiated biochemical substrates

95. Tritiated biochemical substrates, such as glucose, amino acids, hormones, and DNA and RNA
precursors, are produced and widely used in biomedical research, leading to possible exposure byinhalation, skin contamination and possible inadvertent ingestion. A general metabolic feature is thatsuch biochemical substrates may be directly incorporated into organic molecules in body tissue if
absorbed to blood and transported intact to sites of active metabolism within cells. The extent ofincorporation of tritium into specific forms of OBT is determined by such factors as the chemicalcompound containing tritium, its isomeric form, the position of the label in the molecule, and theamount of carrier [B11, B12, F2, L1, T2, T4, T8, T10]. Catabolism of labelled compounds will result intritium being partially oxidized and entering the body water as HTO or catabolized and excreted as low
molecular weight organic substances.

(a) Tritiated glucose and amino acids
96. Studies by Takeda [T2, T4] and Balonov et al. [B8, B11, B12], have shown that the bound fractionof tritium administered intraperitoneally to rats as biochemical substrates varied from 3–5% (tritiatedD,L-alanine, glucose) to 50−80% (tritiated L-tyrosine, L-lysine). The bound fraction of tritium fromtritiated biochemical substrates was substantially larger than that of HTO. The low retention of tritium as
OBT after administration of tritiated glucose is consistent with its rapid catabolism while the highretention of tritiated lysine is consistent with its incorporation into protein as an essential amino acid.
97. Retention of tritium administered as the amino acids, glycine, leucine and methionine, wasintermediate between that of glucose and lysine. D-isomers of 3H-amino acids were assimilated
considerably less than L-isomers. Thus, the level of binding of 3H2-L-leucine-2,3 in different tissues,
except kidneys, was 2–2.5 times higher than that of the D-isomer, and 1.5–2 times higher than that ofracemic mixture. Similar results were obtained with 3H-lysine isomers [B8, B12].
98. Figure VII shows the retention curves for OBT in rat spleen after intraperitoneal injection of HTO
and some biochemical substrates [B8, B12]. Tritium was rapidly excreted after administration of3
H-glucose with a half-time of 3–4.5 days. Retention functions for the majority of 3H-amino acids
showed two components: one with a half-time from 0.7 to 2 days and the second with a half-time from
7 to 16 days, presumably reflecting metabolism of two groups of functionally different proteins.
Retention times were shown to be tissue specific. In actively proliferating tissues—bone marrow, small
intestine, testis—the observed kinetics of OBT is influenced by the processes of cell differentiation andmovement of labelled cells out of the organ. Due to high demand for L-lysine in tissues and its
reutilization, the kinetics of its excretion from tissues is slowed two–three-fold relative to other aminoacids [B8, B12, T2, T4].

268 UNSCEAR 2016 REPORT

Figure VII. Specific activity of bound tritium in rat spleens normalized to intraperitoneally injected
activity of tritium (per g of body weight) after injection of HTO, D-glucose-6-3H, glycine-2-3H2 andL-lysine-6-3H [B8, B12]

99. Table 5 presents estimated values of absorbed dose (D) in four organs and tissues and thecontribution of OBT to dose for HTO and three biochemical substrates derived from the same studies of
intraperitoneal administration to rats [B8, B12]. Injection of 3H-glucose did not create a dose
significantly higher than the dose from an equal amount of HTO, and differed only in the increased
contribution to dose from OBT (7–23% against 2.5−4%). The dose following administration of3
H-glycine was 5–40% greater than for HTO, and contribution of OBT to the dose reached about
20-40%. Notably, for administration of L-lysine-3H, the dose was two–eight times that for HTO, andthe contribution from OBT was >90%.

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269

Table 5. Absorbed dose, D, in rat organs and tissues (mGy) and OBT contribution to dose (%) after
intraperitoneal injection of 37 kBq/g of HTO and tritiated biochemical substrates [B8, B12]

Organ, tissue

D-glucose-6-3H

HTO

Glycine-2-3H2

L-lysine-6-3H

D (mGy)OBT (%)

D (mGy)OBT (%)

D (mGy)OBT (%)

D (mGy)OBT (%)

Bone marrow

21

4

21

19

26

38

92

97

Small intestine

–

–

22

23

–

–

59

95

Testis

20

3

19

7

21

24

40

92

Muscle

20

2.5

–

–

28

43

150

98

100. Absorption coefficients for ingestion of and skin contamination by tritiated biochemical substrates
were obtained in experiments in which preparations were administered to rats. The absorption
coefficient was defined as the average ratio of OBT levels in tissue after ingestion or skin application to
those after intraperitoneal injection. While 3H-glucose and 3H-amino acids were completely absorbedby the gastrointestinal tract [B12], absorption of 3H-thymidine was considerably lower (10–20%)
because of its degradation to 3H-thymine in the gastrointestinal tract. In contrast, 3H-deoxycytidine wasabsorbed almost completely (60–100%). In rats fed HTO, tritiated amino acids, glucosamine, or
tritiated DNA or RNA precursors for 22 days, the greatest concentrations of OBT were found after
exposure to amino acids, with intermediate concentrations found after exposure to DNA and RNA
precursors [T4]. Studies using rat everted gut sacs showed only about 2% of tritiated thymidine crossed
the intestinal epithelium [L2].
101. During six hours after the application of preparations on rat skin, 1–4% of tritiated substrates wereabsorbed in blood, and 0.01–0.5% of tritium activity was measured in the skin layer at the place of
application after its decontamination [B8, B12].

(b) Tritiated nucleic acid precursors
102. The DNA precursors, deoxythymidine and deoxycytidine labelled with tritium, have mostcommonly been used in studies of cell kinetics. For work involving RNA, tritiated uridine and adenine
are the precursors that have been used [H16]. After tritiated thymidine has been orally administered tohumans or animals, about 2% is incorporated into DNA during the synthesis stage of the cell cycle[L2], and the remainder appears as HTO. Tritiated thymidine is available for only a short time after
intake and primarily for uptake by rapidly cycling cells such as those of the bone marrow or gut.
However, prolonged administration of 3HTdR throughout gestation results in labelling of slower
cycling cells [K11].
103. According to Feinendegen et al. [F5], nuclei of 5–30% of proliferating cells in mammals arelabelled by tritium. As the average range of tritium beta radiation is considerably less than thedimensions of the nuclei of mammal cells, and distribution of DNA-bound tritium is extremely
inhomogeneous both on the scale of organs and tissues and also within cells, the concept of the averageorgan or tissue dose in the case of incorporation of 3H-nucleosides requires care in its interpretation. An
alternative is to estimate dose to radiosensitive nuclei in rare cases of tritiated DNA-precursor
incorporation by workers.

270 UNSCEAR 2016 REPORT104. Especially in embryo,3H-deoxynucleosides are actively included in intensively proliferating cellsystems of bone marrow and small intestine, and considerably less in tissues with small frequency of
mitoses (e.g. muscle, liver). A high concentration of 3H-CdR in bone marrow of rats is noteworthy[B12], contrasting with lower values obtained in mice [F5]. The dose in cell nuclei of bone marrowlabelled by 3H-CdR estimated in studies by Balonov et al. [B8, B12] according to the methodology
suggested by Feinendegen and Cronkite [F4, F5] is larger by two orders of magnitude than the average
tissue dose from an equal amount of injected HTO.
105. Taylor [T10] reviewed the biokinetics of 11 xenobiotic tritiated organic compounds and estimatedthat the clearance half-time in humans was less than 40 days in all cases. Some organic compounds may
be incorporated directly into structural components and retained for longer periods.

6.

Metal tritides and other forms with low solubility

106. Tritiated compounds with low solubility include luminous compounds (powder and paint), tritidesof metals (e.g. Ti, Zr, Hf), microfragments of glass and carbon and beryllium particles contaminatedwith tritium. Such compounds are produced and widely used in industry (painting wrist watches andcompasses with luminous paint, ionization sources) and research (e.g. accelerator targets, tritium carrierin fusion physics). Tritiated compounds with low solubility are considered as sources of internal
exposure in workplaces.

(a) Metal tritides
107. When tritides of metals (e.g. Ti, Zr, Hf) are used for industrial or research purposes, the tritium
chemically bound in the crystal matrix is gradually desorbed from the metal surface in the form of HTO
or HT. Due to external exposure of the device surface to radiation fluxes and heat, metal tritides canalso be released into the work environment in particulate form. The main pathway of radiation exposureof workers dealing with devices containing metal tritides is the inhalation of HTO vapour and airborne
particles of metal tritides.
108. Balonov et al. [B8, B11] reported that, following short-term inhalation by rats, titanium tritide(TiT) showed slow lung clearance during one month after exposure. Similar experiments of Cheng etal. confirmed slow clearance of a fraction of TiT (1 μm count median diameter (CMD)) after
intratracheal instillation of TiT suspension into rats [C18, Z8]. Similar results were obtained inexperiments on rats with hafnium tritide (1 μm CMD) [Z7] and zirconium tritide (0.3 μm CMD) [Z9].
The size of the slow cleared fractions and uptake rates of tritium dissolved from articles depend on themethods of TiT particle production and particle size.
109. In a series of in vitro experiments with powders of titanium, hafnium and zirconium tritides aimed
at simulating time-dependent absorption functions of tritium in the respiratory tract of rats and humans,
the dissolution of tritium in synthetic serum ultrafiltrate was studied during 30–200 days by Cheng etal. [C17, C19, Z9]. The dynamics of tritium release from particles was well described with two
exponentials, one with a half-time in the range from one day (TiT) to about 50 days (HfT and ZrT) andthe second with a half-time in the range from one month (TiT) to one year (ZrT) and several hundredyears (HfT). The long-term dissolution half-time from coarse powder (~100 μm) was larger than fromfine powder (~1 μm) [C17].

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271

110. In vitro and in vivo experiments were complemented with dosimetric considerations of beta
radiation self-absorption in particulate material [C19], which became important for particle sizes of
0.1 μm and more. From the available data, it was concluded that airborne ZrT and HfT should beconsidered in human internal dosimetry as material of slow solubility (ICRP Type S), and TiT as
material of medium solubility (ICRP Type M).

(b) Components of nuclear fusion reactors
111. Tritium is present in nuclear fusion facilities in various physical and chemical forms, some ofwhich have radiological properties different from HTO and HT. In the 1980s, features of tritiatedmicroballoon glass fragments that could be potentially inhaled by workers were studied both in vitro
and in vivo [C27, C28]. The median diameter of glass fragments was initially estimated to be about
4 µm but, accounting for fragment shape, the mass median diameter was specified as 20 µm.
Experiments in vitro have shown that 98% of tritium is released from glass with a half-time of 3–9 daysand 2% with longer half-time of 23–280 days. After intratracheal instillation of tritiated glass fragments
in rats, 93% of the tritium activity was removed from glass with half-time of 6±0.5 days and 7% with alonger half-time of 43±3 days, consistent with the results of the in vitro study. The resultant dose to thelung was three orders of magnitude greater than the dose to other tissues, and approximately 40 times
greater than the dose incurred in the lung from the inhalation of a similar quantity of HTO. However,
because of the low inhalability of tritiated glass fragments of such sizes, this ratio would be much lower
in a direct comparison of inhalation doses.
112. Since 1999, when the Joint European Torus (JET) fusion tokamak started its operation withtritium, tritiated dust and flakes were observed predominantly during maintenance operations. Thosecarbon, beryllium and tungsten particles were formed due to the interaction of plasma with the carbonbased first internal wall of the fusion reactor. The activity median aerodynamic diameter (AMAD) of
the dust particles was assessed to be ~4 µm with high specific activities of up to 3 GBq/g and AMAD
of flakes as about 100 µm [D4]. After inhalation, the former can deposit in different parts of therespiratory tract and expose its tissue to beta radiation. The in vitro dissolution tests with tritiated dust
have shown that 1–5% of tritium activity was dissolved in lung serum simulant within one minute and afurther 1–20% of tritium were dissolved over the next 100 days [H15]. Slow excretion of tritium
deposited in lungs may complicate its individual monitoring [R13]. From the available data, it can beconcluded that carbon and beryllium dust particles from fusion reactors can be classified as material of
medium or slow solubility (ICRP Type M or S).

(c) Luminous compounds
113. The basis for self-luminous tritium paint is fine zinc sulphide powder (~10 µm) coated with a thinlayer (0.01−0.1 µm) of tritiated polymer with high specific activities. In order to reach higher lightintensities, tritiated polymers (polystyrene, silicon rubber) with specific activities up to 20 TBq/g havebeen used [B8, I3]. During the paint luminizing process, workers may be exposed to T2, HTO, andvapours of tritiated organic solvents via inhalation and skin contact. Particles deposited in the lungs will
irradiate tissues by emission of beta particles and bremsstrahlung. HTO and unknown organic
compounds have been detected in urine [B8, R15], formed by degradation of polymers by radiolysis,oxidation and isotopic exchange.
114. In the past, some workers were exposed to high internal doses that were fatal in a few cases [M15,
S12]. Lambert and Vennart have shown that radiological control in the workplace usually necessitates
continuous biological monitoring of workers for tritium in urine at relatively short intervals [L5].

272 UNSCEAR 2016 REPORT115. In vitro experiments have shown that 0.5–5% of tritium from tritiated polystyrene-based “Soviet
luminous powder (PS-A)” was gradually released in buffer mixture as HTO and low molecular organic
foreign compounds [B8, B11]. In vitro studies of the dissolution of commercial luminous powder madefrom tritium-labelled polystyrene in bovine serum over five days [R15] showed that an average of 12%
dissolved on the first day, and about 2% of the remaining activity on subsequent days.
116. Similar uptake (0.5–5% depending on compound age) of both HTO and low molecular organic
compounds was observed in experiments with PS-A luminous powder orally administered to rats and ina similar study with three volunteers [B8, B11]. Substantial fractions of the tritium compounds wererapidly excreted in urine and faeces of rats, a proportion was retained for some days in the liver/kidney,
and the remainder was catabolized to HTO. The half-time of organic tritium in human urine was about
one day and the remaining tritium was excreted as HTO with a mean half-time of about 16 days.
117. Experiments on cats showed that absorption of tritium from luminous paints depended on theplastic substrate involved, with values of 0.007 for polystyrene, about 0.03 for silicone rubber and
0.8 for polyester [H12, W6]. Balonov et al. [B8, B11] also reported that following intratracheal
instillation of PS-A luminous powder into rats, the lung specific activity showed essentially no decrease
within five months, demonstrating its very low solubility.
118. In summary, from the available experimental and human data, it can be concluded that the
specific characteristics of tritiated luminous compounds differ substantially from those of HTO. Special
attention should be paid to individual bioassay monitoring.

B. Overview of current biokinetic models for tritium
119. Tritium may enter the body by inhalation, absorption through skin, and ingestion. The first two
are the more frequent routes of intake in the workplace, while the latter mostly contributes to exposureof members of the public. Furthermore, skin contact with tritium-contaminated surfaces, such as metaland glass, has been shown to result in the formation of OBT in the body [E1]. This has been shown to
be a route of tritium intake in the workplace [H12].
120. The fate of tritium once taken into the body is determined mostly by its chemical form. One can
expect to find HTO in most workplaces and environmental media where tritium is present. In general,consideration of the biokinetics of OBT refers to the non-exchangeable component resulting fromtritium atoms bonding directly to carbon in organic molecules and exhibiting retention times relating tocarbon turnover; the exchangeable component of tritium in organic molecules, in the form of hydroxyl
and sulphydryl groups, for example, has the same metabolism and distribution in the body as HTO.
121. Tritiated compounds may also exist as airborne particles, e.g. metal tritides or luminous powder. The
retention and clearance of these particles from the respiratory tract depend on several factors, such as
particle size and chemical composition. For dose assessment purposes, tritium absorbed into blood after
tritiated particles have been inhaled is usually treated as being in the form of HTO [C18, C19, C20, I17].
122. Tritiated compounds are categorized by ICRP for radiation protection purposes according to themetabolic model that best describes their dynamics after intake and subsequent uptake. Three primarymetabolic models are currently used by ICRP either separately or in combination to calculate
committed effective dose for an intake of tritiated compounds [I8, I9, I10, I12, I14, I17]:

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

273(a) A model for tritium absorbed to blood as HTO following either ingestion or inhalation, appliedalso to other tritiated compounds that partially convert to HTO after being taken into the body;(b) A model for tritium absorbed to blood following intake of OBT, mainly by ingestion in food,but also applied to inhalation of non-specific organic molecules, and to ingestion or inhalation ofsome specific tritiated organic compounds;(c) The generic ICRP model for the human respiratory tract, specifying absorption parametervalues for inhalation of poorly soluble forms of tritium.
123. All three models and respective sets of dose coefficients are widely used by the Committee in its
main reports, by the IAEA in its international standards (e.g. [I4]) and also in documents of WHO[W7], FAO and the CODEX Alimentarius Commission [C24]. The ICRP models—and especially dosecoefficients—are included in numerous national regulations on worker and public protection againstinternal exposure with different forms of tritium.

1.

Dosimetry of tritium using model for HTO(a) Biokinetic model for HTO
124. The ICRP biokinetic model for systemic HTO [I12, I14, I17], illustrated in figure VIII, is used to
calculate the committed effective dose from:(a) Intakes of (HTO);(b) HTO formed following inhalation of elemental hydrogen (HT, T2);(c) HTO formed following inhalation of tritiated hydrocarbon vapours and gases (e.g. CH3T).
Use of this model involves assumptions about the conversion of the other forms of tritium to HTO.
125. For modelling HTO intake, an instantaneous translocation to blood is assumed for both inhalationand ingestion. The ICRP model further assumes that HTO is transferred from blood with a biologicalhalf-time of six hours and distributed uniformly throughout the body. It is also assumed that 97%
remains as HTO, while 3% is instantly converted to OBT. In adults, HTO is assumed to be retainedwith a biological half-time of 10 days, and OBT with the biological half-time of carbon, calculated as
40 days (figure VIII). ICRP values for the partitioning between HTO and OBT after acute intake of
HTO in various age groups, and the corresponding biological half-times, are given in table 6 [I12, I14,
I17]. Age dependence of biological half-time for HTO was derived by ICRP from available humanobservations and physiological data [I7].
126. The calculation of committed effective dose per unit intake (dose coefficient) for adults
resulting from the intake of HTO, as given by ICRP [I14, I17], is based on the model presented infigure VI. The current value for intake of HTO by adults both by inhalation and ingestion computed
by ICRP is 1.8 × 10-11 Sv/Bq [I14, I15, I17].
127. The contribution of the OBT fraction to the committed effective dose in the HTO model has beenshown to be about 10% [I8, I9, I10, J3]. This is adequate for estimates of dose from both acute andprolonged intake [T16].

274 UNSCEAR 2016 REPORT

Table 6. ICRP parameter values for distribution and retention of tritium after acute HTO intake[I12, I14, I17]
Age

Initial distribution (%)

Biological half-time (days)

HTO component

OBT component

HTO component

OBT component

3 months

97

3

3.0

8

1 year

97

3

3.5

15

5 years

97

3

4.6

19

10 years

97

3

5.7

26

15 years

97

3

7.9

32

Adult

97

3

10.0

40

Figure VIII. ICRP model for biokinetics of HTO [I12, I14, I17](b) Absorption of HTO through skin128. Immersion in a vapour or liquid containing HTO also results in its absorption through the skin[D3, O6, P9]. In the context of occupational exposure to HTO vapour, ICRP [I8, I9, I10] referred to thework of Osborne [O6], Hill and Johnson [H12], and the review of Myers and Johnson [M26] andconcluded that about 1% of the HTO activity per cubic metre of air may be assumed to be absorbed
through the skin in a minute. On this basis, the amount absorbed through skin contributes about one
third of the total HTO intake for a given HTO concentration in air when the exposed individual is active
during exposure (i.e. breathing in more air than when at rest). The work of Osborne [O6] was based on
direct measurements of the absorption rate of HTO vapour through the skin of the whole body.

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

275(c) Inhalation of tritium gas
129. Following inhalation of HT, a small fraction (about 0.01%) is dissolved in body fluids andoxidized to HTO [P4]. The latter is the predominant contributor to the committed effective dose. HT is
not significantly absorbed through the skin and does not readily convert to HTO on the skin. Irradiationof the lungs by inhaled HT does not significantly increase the committed effective dose [I16] becauseof the short range of the tritium beta particles in lung tissue. The dose coefficient for inhalation of HT is
therefore about 0.01% of the dose coefficient used for inhalation of HTO.

(d) Inhalation of tritiated hydrocarbons
130. Tritiated methane is the only tritiated hydrocarbon for which ICRP recommends a dose coefficientfor inhalation based on the HTO model. Tritiated methane is known to be formed as a result of
microbial degradation within tritiated waste. About 1% of the inhaled tritiated methane is assumed to beconverted to HTO [P5]. The current ICRP dose coefficient for tritiated methane is therefore 1% of thatfor HTO [I17]. The ICRP approach is considered to give a conservative dose coefficient [P5].
131. Carlisle et al. reported in a more recent study on rats that the fraction of tritiated methane retainedin the body as HTO and OBT after acute inhalation to be about 0.06 to 0.13%, which is lower than thatassumed by the current ICRP model [C6]. However, the observed conversion of tritiated methane to
OBT in rats was greater than that estimated by the ICRP model in all human tissue examined [I17], in
particular in the liver, where the conversion was observed to be 22 times greater than that of HTO
estimated by ICRP. The authors further suggested that some of tritium taken in as tritiated methane isconverted directly to OBT. They concluded that the committed dose to some organs is one third to one
tenth of that estimated by ICRP and that the ICRP value of effective dose may be conservative.

(e) Summary of dose coefficients based on HTO model
132. Table 7 presents the current ICRP committed effective dose coefficients for tritiated compounds
based on the biokinetic model for HTO and table 8 illustrates the effect of age on the computed
committed effective dose per unit intake for the inhalation of HTO [I17].
Table 7. ICRP effective dose coefficients based on HTO model for various tritiated compounds,
modes of intake and age groups [I14, I17]
Tritiated compound

Mode of intake

Dose coefficient (Sv/Bq)
Infants (1 year old)

Adults−11

HTO

Inhalation

4.8 × 10

1.8 × 10−11

HTO

Ingestion

4.8 × 10−11

1.8 × 10−11HTInhalation

4.8 × 10−15

1.8 × 10−15

CTH3

Inhalation

4.8 × 10−13

1.8 × 10−13

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Table 8. ICRP effective dose coefficients for HTO inhalation by various age groups [I17]AgeDose coefficient (Sv/Bq)

Ratio of dose coefficient to that of an adult

3 months

6.4 × 10

3.6

1 year

4.8 × 10

2.7

5 years−11

3.1 × 101.7

10 years

2.3 × 10−11

1.3

15 years

1.8 × 10−11

1.0

Adult

1.8 × 10

1.0

2.

−11
−11

−11

Dosimetry of OBT and tritiated biochemical substrates using model for OBT

133. Figure IX illustrates the ICRP model [I14] used to calculate committed effective dose for theingestion of OBT. It is assumed that OBT, once taken into the body, is translocated to blood completely
and instantaneously, and is transferred from blood to tissue with a biological half-time of six hours,with 50% retained in tissue as OBT and 50% transformed into HTO. The uptake and retention oftritium in various tissues depends on the metabolic activity of the individual tissues and the constituent
chemical forms of OBT. However, the ICRP model assumes uniform distribution of doses from OBT to
all soft tissue in the body while a greater metabolic activity leads to greater uptake and more rapid loss.
Figure IX. ICRP model for biokinetics of OBT [I14]134. The ICRP model for the ingestion of OBT [I12, I14] is intended to represent the biokinetics of theaverage dietary content of the different chemical forms of OBT. The model was developed in theabsence of information about the exact proportions of the various chemical components of OBT in thehuman diet and the turnover of these components [I12, I14]. On the basis of the turnover of hydrogen in
Reference Man, less than 10% of tritium taken in daily as OBT in diet is assumed to be excreted daily
in the form of OBT (mostly in the form of urea, with about 3% in faeces) while the rest is assumed to
be excreted as HTO [R8].

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277

135. Table 9 gives the assumed partitioning between HTO and OBT and the corresponding biologicalhalf-times after dietary intake of OBT in various age groups [I14]. The ICRP committed effective dosecoefficients for OBT are shown in table 10 [I17]. Table 11 shows the effect of age on the committed
effective dose per unit intake for the ingestion of OBT.
136. The ICRP model for OBT and relevant dose coefficients can also be used for prospective dose
assessment in cases of intake of tritiated biochemical substrates if more specific biokinetic information
is not available. For most biochemical compounds, this approach will lead to some overestimation ofeffective dose [T10]. Considering a range of biochemical compounds for which biokinetic data are
available from animal experiments, only the essential amino acid, L-lysine-3H, resulted in estimatedtissue doses greater than those for HTO (by a factor of 2–8) [B11, B12, T2].
137. The ICRP model for OBT and the corresponding dose coefficients are not applicable for dose
assessments in cases of occupational intake of tritiated nucleic acid precursors. After tritiated DNAprecursors are taken in by humans or administered to animals, some fraction of tritium is incorporated
into DNA during the synthesis stage of the cell cycle [L3], and the remainder appears as HTO or
metabolized biochemical substrates. As distribution of DNA-bound tritium and its radiation energy isextremely inhomogeneous in organs, tissues and inside cells, the concept of average organ or tissue
dose in this case requires careful consideration. An alternative approach is to calculate dose taking the
localization of tritium in cell nuclei into account [F2, F3, F4, F5, N1].
138. The National Council on Radiation Protection and Measurements (NCRP) [N1] examined theabsorbed dose resulting from the ingestion of 3HTdR based mainly on theoretical considerations. For
acute intake, it concluded that the absorbed dose to stem cells and bone marrow (per unit intake)resulting from the ingestion of tritiated thymidine is greater by an order of magnitude than that resulting
from the ingestion of HTO. This conclusion may need to be revised when additional data on stem cells
and 3HTdR distribution and incorporation rates become available. It was noted that biokinetics oftritiated nucleic acid precursors is strongly dependant on mammal species [B11, B12]. Without
appropriate human biokinetic data, the modelling uncertainty remains unspecified.
Table 9. Partitioning between HTO and OBT after dietary intake of OBT [I14]AgeInitial distribution (%)

Biological half-time (days)

HTO component

OBT component

HTO component

OBT component

3 months

50

50

3.0

8

1 year

50

50

3.5

15

5 years

50

50

4.6

19

10 years

50

50

5.7

26

15 years

50

50

7.9

32

Adult

50

50

10.0

40

278 UNSCEAR 2016 REPORT

Table 10. ICRP effective dose coefficients based on OBT model for various modes of intake and age
groups [I17]
Tritiated compound

Dose coefficient (Sv/Bq)

Mode of intake

Infants (1 year old)

Adults

OBT

Inhalation

1.1 × 10−10

4.1 × 10−11

OBT

Ingestion

1.2 × 10−10

4.2 × 10−11

Table 11. ICRP effective dose coefficients for OBT ingestion for various age groups [I17]AgeDose coefficient (Sv/Bq)

Ratio of dose coefficient to that of an adult

3 months−10

1.2 × 10

2.9

1 year

1.2 × 10−10

2.9

5 years

7.3 × 10−11

1.7

10 years

5.7 × 10

1.4

15 years−11

4.2 × 10

1.0

Adult

4.2 × 10−11

1.0

3.

−11

Revised ICRP dosimetry models for HTO and OBT

139. ICRP has developed revised models with improved physiological realism focussing on
occupational intakes of radionuclides. Intakes by members of the public will be considered by ICRP at
a later stage.
140. The HTO systemic model includes compartments representing blood, extravascular body waterthat exchanges rapidly with blood, and two components of retention of tritium converted in vivo to
OBT. The revised ICRP model structure is shown in figure X. The transfer coefficient from blood to
excreta is set to yield an initial removal half-time from the body of 10 days. The transfer coefficientsfrom compartments OBT-1 and OBT-2 back to extravascular HTO correspond to half-times of 40 days
and one year, respectively; the net retention half-times in these compartments are slightly longer than
40 days and one year due to recycling of activity. Excretion pathways from blood are not shown infigure X but the following division is assumed on the basis of reference data for water balance ICRP
Publication 89 [I19]: urine, 55%; faeces, 4%; exhalation, 12%; and loss through skin (sweat plus
insensible loss), 29%.

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279

Figure X. Revised ICRP systemic model for HTO
Reproduced with the permission of ICRP

141. The revised model for systemic tritium applied to intake of OBT is a modification of the model for
OBT applied in ICRP Publication 56 and is shown in figure XI. In relation to occupational exposures, thismodel is assumed to apply to “biogenic organic compounds” of tritium for which specific information is
not available. It is assumed that 50% of tritium initially entering blood transfers immediately tocompartment OBT-1 and 50% is converted immediately to HTO within the blood compartment. Tritium
entering OBT-1 or blood subsequently follows the HTO model defined in figure X.
142. The revised model for HTO predicts that OBT would represent about 5–6% of total body tritium
following chronic exposure to HTO. The OBT model with the default initial division of activity between
OBT-1 (50%) and blood (50%) predicts that OBT would represent about 65–70% of total-body tritium in aworker who is chronically exposed to a biogenic form of tritium. The adult dose coefficient calculated using
the revised model for HTO is 1.9 × 10−11 Sv/Bq compared with the current value of 1.8 × 10−11 Sv/Bq.
Values for intakes of biogenic organic compounds are given as 3.5 × 10−11 Sv/Bq for inhalation (Type F) and5.1 × 10−11 Sv/Bq for ingestion (complete intestinal absorption, fA=1). The value for ingestion compares withthe current value for OBT of 4.2 × 10−11 Sv/Bq.
143. The ICRP will also publish revised dose coefficients for inhaled tritiated methane and particulate
forms for which the HTO systemic model is applied. The revised value for inhalation of tritiated methaneby adults is 5.8 × 10−14 Sv/Bq compared with the current value of 1.8 × 10−13 Sv/Bq, reflecting theassumption of 0.3% deposition and absorption to blood as HTO compared with 1% in the current model.
Revised values for inhaled particulate forms are 1.3 × 10−11 Sv/Bq for Type F and 2.4 × 10−11 Sv/Bq forType M, compared with current values of 6.2 × 10−12 Sv/Bq and 4.5 × 10−11 Sv/Bq, respectively. The
value for Type S materials is unchanged at 2.6 × 10−10 Sv/Bq.

280 UNSCEAR 2016 REPORT

Figure XI. Revised ICRP systemic model for OBT
Reproduced with the permission of ICRP

4. Dosimetry of tritiated compounds using model for low solubility
particulate tritium
144. Dose coefficients for inhaled particulate tritium were introduced by the ICRP Publication 71 [I17]in 1995 following publication of a number of scientific papers in which those forms of tritium wereidentified and partially characterized [B8, B11, H12, I3, R15, T8, W6]. The papers described conditionsof occupational exposure in research and industry during contact of workers with neutron generatorsand particle accelerators and with luminous powder.
145. Calculation of dose coefficients for particulate tritium was based on the generic ICRP model ofthe respiratory tract [I16], categorizing tritium aerosols according to the absorption types specified in
the model, i.e. fast (F), medium (M) and slow (S). It was assumed that: (a) beta radiation of tritium
located in the alveolar part of the respiratory tract was fully absorbed in lung tissue, and (b) tritium
separated from the particles deposited in the respiratory tract, due to both fast and slow processes,
behaved as HTO. Therefore, the ICRP generic lung model was combined with the specific ICRP modelfor HTO (see figure VIII). The clearance of tritium from the respiratory tract following deposition oftritium particles included the escalation of particles through the bronchial tree to the alimentary tractand dissolution and absorption to blood as HTO.
146. Data on the metabolism of inhaled particles of metal tritides are scarce [B8, B11, C17, C18, C19,
C20]. In the absence of specific information, ICRP has considered these particles as Type M [I17],
meaning that their rate of absorption from the respiratory tract to blood is moderate. Richardson andHong [R7] conducted dosimetric modelling of inhaled tritiated particles and reported that, while theANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

281

dose to the alveole can be up to two orders of magnitude higher than that from the same activity of
inhaled HTO, taking account of self-absorption of beta radiation in particulate material can reduce doseby up to an order of magnitude, depending on particle size, in the range of 0.1–10 µm.
147. Following inhalation of tritiated organic compounds such as luminous ones, uptake of HTO is
accompanied by uptake of low molecular compounds of tritium originating from degradation oftritiated polymer. The contribution of the tritiated low molecular compounds to the effective dose isgenerally low because of their rapid excretion in urine and faeces. However, these elevatedconcentrations of organic compounds of tritium in excreta should be recognized when interpreting theresults of individual monitoring.
148. The available dose coefficients for particulate tritium are applicable mostly for occupationalradiation protection of workers in research and industry dealing with neutron generators, particle
accelerators, and air ionizers and the manufacture of luminous products. However, for the substantiallylower probability of exposure of members of the public, ICRP has provided committed effective dosecoefficients for different age groups (table 12).
Table 12. ICRP effective dose coefficients (Sv/Bq) for inhalation of particulate tritium of various types [I17]AgeType F

Type MType S

3 months

2.6 × 10

−10

3.4 × 10

1.2 × 10−9

1 year

2.0 × 10−11

2.7 × 10−10

1.0 × 10−9

5 years

1.1 × 10−11

1.4 × 10−10

6.3 × 10−10

10 years

8.2 × 10−12

8.2 × 10−11

3.8 × 10−10

15 years

5.9 × 10−12

5.3 × 10−11

2.8 × 10−10Adult

6.2 × 10−12

4.5 × 10−11

2.6 × 10−10−11149. Dose coefficients for Type F aerosols are lower than those for HTO despite the assumption of
rapid dissolution and absorption of deposited particles. This is because a proportion of inhaled particles
is subsequently exhaled while deposition of inhaled HTO vapour is assumed to be complete. The main
contribution to effective dose from inhalation of Type M and S aerosols is from lung dose. The dosecoefficients presented in table 12 can be applied for internal dose assessment in cases of inhalation ofvarious airborne tritium particles, e.g. metal and graphite tritides and iron hydroxide used in tritium
facilities, luminous powder and other particulate forms.

C. Intakes of tritium in relation to pregnancy and breast-feeding
1.

Pregnancy and tritium intake

150. A few studies have investigated the transfer of tritiated compounds to the embryo and fetus andtheir distribution and retention in fetal tissues. In prenatal dosimetry, the term embryo refers to thedeveloping human offspring up to the end of the eighth week of pregnancy, from the initial stages ofgrowth up to the end of organogenesis. At this time, the embryo weighs less than about 10 g.

282 UNSCEAR 2016 REPORT151. Harrison et al. [H6] measured the transfer, distribution and retention of tritium in rats and guinea
pigs following the administration of HTO, tritiated glucose and tritiated food (i.e. liver and cress).
Transfer and retention of tritium in fetal tissue were similar for HTO and tritiated food. However, theretention of tritium in both maternal and fetal tissue for tritiated glucose was lower. The ratio of theconcentrations of tritium in the fetus (CF) to that in the mother (CM) at the end of organogenesis in rats
ranged from 0.4 to 1.0. At the end of pregnancy, the ratio ranged from 1.3 to 1.5 in the case of rats andfrom 1.1 to 1.3 in the case of guinea pigs.
152. Takeda et al. [T5] exposed female rats to single oral administrations of HTO, tritiated thymidine
or tritiated lysine. No significant differences were observed between the tritium concentrations in thefetus and in maternal tissue. Tritium concentrations in the fetus were greater following the ingestion oftritiated lysine than following the ingestion of HTO or tritiated thymidine. Tritiated lysine also resulted
in greater prenatal and neonatal absorbed doses than tritiated thymidine and HTO by factors of 1.5(when tritium was administered on the thirteenth day of gestation) and 6 (when tritium wasadministered on the first day of nursing).
153. Pietrzak-Flis et al. [P7] studied tritium incorporation in rats chronically exposed to tritiated food(activity concentration, 48.1 kBq/g) or HTO (activity concentration, 37.0 kBq/mL) for three successive
generations (from three weeks before mating of the parents to the delivery of the F3 generation). The
analysis of tissue at various ages showed that the absorbed dose rates were higher in rats exposed to
HTO. However, the amount of tritium incorporated into the organic fraction of tissue was several times
higher after exposure to tritiated food.
154. Kowalska [K22, K23] reported further results from the same study showing that the amount oftritium incorporated into the amino acids of rat-brain proteins and the main rat-brain phospholipids andgangliosides was higher after the ingestion of tritiated food than after the ingestion of HTO. The highest
tritium concentrations in rat-brain proteins followed in utero exposure while the highest tritium
concentrations in phospholipids and gangliosides were found in 21-day-old rats exposed during
pregnancy and lactation.
155. The ICRP Publication 88 [I18] has provided biokinetic and dosimetric models and dosecoefficients for the embryo, fetus and newborn as a result of intake of radionuclides by the mother. The
term fetus refers to the developing human offspring after the eighth week of pregnancy [I18]. The
equivalent dose to the embryo is assumed to be the same as that to the uterus wall and proportional tothe concentration of HTO in maternal body water. For the calculation of the equivalent dose to thefetus, the HTO concentration in fetal body water was assumed to be equal to that in the mother. The
ICRP uses a simple approach to the calculation of fetal doses for the majority of elements and theirradioisotopes, including tritium, considering data collected from studies of animals and humans [I18,
I21]. These fetal doses from tritium are calculated on the basis of relative concentrations of tritium
averaged over the whole body of the fetus (CF) and that of the mother (CM). The CF:CM ratios are taken
mainly from studies presenting data obtained at short times post-intake.
156. HTO rapidly crosses the placenta after it is inhaled or ingested by the mother. The resulting
equivalent dose to the fetus depends on the water content of the fetus over the course of its
development. As gestation progresses, the water content decreases relative to the amounts of protein,
fat and minerals—from about 95% of the total body mass at the sixth week of pregnancy to about 70%
at birth. For comparison, the percentage of total body water in a non-pregnant woman is about 50%.
The total water content of the mother’s body is known to increase during pregnancy. This is mainly due
to the increase of about 50% in the blood volume in order to carry additional nutrients and othersubstances needed for the fetus. Calculations have shown that the biological half-time of HTO in the

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mother varies from about 10 days at the start of pregnancy to about 12 days at term. This variation does
not significantly affect the ICRP fetal dose coefficients for HTO [I18].
157. On the basis of an average body water content of 80% for the fetus and 50% for the mother, ICRP
used a CF:CM ratio of 1.6 in the derivation of the dose coefficients for HTO. This ratio is applied to boththe HTO and OBT components for intake of OBT and is assumed to remain constant throughout
pregnancy. Tritium is assumed to be uniformly distributed throughout all tissues of the fetus. Harrison
et al. [H7] showed that the CF:CM ratio ranged from about 1.4 to 1.8, on the basis of data for the relative
water content of the mother to that of the fetus.
158. Following birth, the ICRP biokinetic model for the three-month-old infant [I18] is used to
calculate the committed effective dose per unit intake by the newborn child. Table 13 presents thecommitted effective dose coefficients for the unborn and the three-month-old child. The ICRP [I18] haspublished a range of dose coefficients for in utero exposure of the child following maternal intake byinhalation or ingestion, before or during pregnancy, considering acute and chronic exposure. Account istaken of the retention of activity in body tissue at birth and dose delivered post-natally. The purpose of
the ICRP fetal dose coefficients is to allow comparison with doses to other age groups in order toensure that protection does not neglect doses received in utero [C29, I22].
Table 13. ICRP prenatal and infant effective dose coefficients [I18]
Tritiated compoundand mode of intake

Prenatal dose coefficient (Sv/Bq)
Acute maternal intake

Prolonged maternal intake

Dose coefficients for
3-month-old child (Sv/Bq)

HTO inhalation−11

3.6 × 10

3.1 × 10−11

6.4 × 10−11

HTO ingestion

3.6 × 10−11

3.1 × 10−11

6.4 × 10−11

OBT ingestion

7.6 × 10−11

6.3 × 10−11

1.2 × 10−10

a

b

a

Values are for acute intake by the mother at the end of the tenth week of pregnancy. Acute intake occurring at other times yields
lower dose coefficients.

b

Prolonged intake by the mother, beginning at the start of pregnancy and continuing for the duration of the pregnancy.

2.

Intake of tritium from maternal milk

159. The ICRP [I21] has developed dose coefficients for newborns for intake of tritium from maternal
milk. The approach involved the estimation of the activity transferred to milk as a function of maternal
intake for various intake scenarios (acute or prolonged intake regimes, and for various maternal intake
times relative to birth). Nursing was assumed to continue for six months after birth, and the ingestion
dose coefficients for infants were applied [I14]. Maternal intake (by ingestion and by inhalation) during
pregnancy and during lactation was considered.
160. The HTO and OBT models were modified by ICRP [I21] to account for the transfer of tritium tomilk fed to infants (up to one year). The rate of OBT transferred to milk was taken to be that of carbon.
The ICRP developed dose coefficients on the basis of the assumption of six feeds per day and an
average daily milk intake by the infant of 0.8 L. Table 14 shows selected committed effective dosecoefficients for nursing infants resulting from maternal intake of tritium.

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Table 14. ICRP effective dose coefficients for nursing infant [I21]
Tritiated compound and mode of intakea

Dose coefficient (Sv/Bq)
Acute maternal intakea

Prolonged maternal intakeb

HTO inhalation

2.2 × 10−11

2.0 × 10−11

HTO ingestion

2.2 × 10−11

2.0 × 10−11

OBT ingestion

3.5 × 10−11

3.0 × 10−11

Values are based on acute maternal intake at one week after birth. Acute intake occurring at other times yields lower dose coefficients.

b

Values are based on prolonged intake during the lactation period (up to six months after birth). These dose coefficients are
greater than those for prolonged intake during pregnancy.

D. Uncertainties in dose coefficients for tritium
161. Dose coefficients applied for assessment of internal dose for purposes of human radiation
protection are defined by ICRP as regulatory parameters without any uncertainty [I23]. They arepresented in ICRP publications for reference persons [I7, I19] as values depending on the radionuclide,its physical and chemical form, exposure pathway (inhalation or ingestion) and a person’s age [I12, I14,
I15, I17, I18, I21]. Uncertainties of ICRP dose coefficients for some radionuclides, including tritium as
HTO, are systematically considered in a specific NCRP publication [N4] and by Leggett et al. [L8].
The need for individual values of metabolic parameters arises mostly in cases of emergency intake oflarge radionuclide activities when probabilities of radiation-induced acute health effects and risk ofstochastic effects require consideration on an individual basis in the context of possible decorporationor other medical treatment.
162. For the most common form of tritium, HTO, dose coefficients for adults are determined on thebasis of numerous human observations and are, therefore, associated with low uncertainty. Greater
uncertainties apply to dose coefficients for children exposed to HTO from the environment. Leggett etal. [L8] considered the sources, quality and completeness of data underlying the biokinetic models fortritium (as HTO) and concluded that the dose coefficient for HTO is known within a factor of two foran adult and between a factor two and three for a five-year-old child.
163. Exchange model parameters for HT inhalation are also derived from human observations, which
can generate moderate uncertainty of corresponding dose coefficients. More uncertain are dosecoefficients for OBT based on animal experiments and modelling and for tritium aerosols of various
types based on both in vitro and in vivo animal experiments [P4, P9].
164. Harrison et al. [H7] reviewed the assumptions used by ICRP in the derivation of dose coefficientsfor tritium along with their associated uncertainties. The assumptions considered were those related toabsorption to blood, the biological half-time of HTO and OBT in adults and children, the transfer to thefetus, the heterogeneity of the distribution in tissues and cells, and the RBE value of the beta particles
emitted by tritium.
165. Harrison et al. [H7] estimated uncertainties using ranges on the central values for theincorporation of tritium into OBT in body tissue of 0.01−0.1 and 0.15–0.75 of the tritium activity
reaching blood after intake as HTO and OBT, respectively. Biological half-times in adults were taken
to vary from about 5 to 20 days for HTO and from about 20 to 200 days for OBT, these ranges being

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the 2.5 and 97.5 percentiles of the log-normal distributions. They also considered CF:CM ratios ranging
from 1.4 to 1.8 based on the relative content of water in the mother and the fetus. The 2.5 and 97.5
percentiles of the log-normal distribution for the OBT transfer and distribution in fetal tissue were taken
to be 1.2 and 2. A range of 1–2.5 was used for the RBE value of tritium beta particles compared to
gamma rays. This analysis gave median (50%) values for the dose coefficient of 2.3 × 10−11 Sv/Bq for
HTO and 5.6 × 10−11 Sv/Bq for OBT for ingestion or inhalation by adults, excluding consideration of
RBE value. Table 15 shows probability distributions of the committed effective dose coefficients for
adults and for the fetus after ingestion by the mother during pregnancy, including the range of RBE
values. Dose coefficients for other age groups were estimated to vary by a factor of two to three for
HTO and OBT.
Table 15. Probability distributions of effective dose coefficients from ingestion of HTO or OBT byadults and for fetus after ingestion by mother during pregnancy [H7]Intakes during pregnancy assumed to take place at 10 weeks after conception

Age
Adult

Fetus

Form

Distributions of effective dose coefficients (10−11 Sv/Bq)
5 percentile

50 percentile

95 percentile

HTO

2.1

3.9

6.6

OBT

3.9

8.7

20

HTO

3.7

7.6

14

OBT

6.9

17

40

166. Hamby [H2], using Monte Carlo sampling, calculated that the dose coefficient for intake of HTO
by adults varied by a factor of 15 from its highest to lowest modelled values, with a median value of
2.2 × 10−11 Sv/Bq and a geometric standard deviation of 1.6. This range was most sensitive to the
biological half-time, the linear energy transfer (LET), and the reference radiation (i.e. whether X- or
gamma radiation). When the quality factor was set to unity, the geometric mean of the dose coefficientwas 1.3 × 10−11 Sv/Bq with a geometric standard deviation of 1.4.
167. Melintescu et al. [M10] also assessed the effect of uncertainty on the RBE value of thetritium beta radiation, the retention of HTO and OBT and the presence of tritium in the DNA
hydration shell. For the adult male, setting the radiation weighting factor to unity (based on the
RBE value) resulted in dose coefficients ranging from 1 to 2.9 × 10 −11 Sv/Bq for the intake of HTO
and from 5 to 6.7 × 10 −11 Sv/Bq for the ingestion of OBT. Accounting for the variability in theradiation weighting factor, the resulting dose coefficient range was from 1.5 to 5.1 × 10 −11 Sv/Bq
and 5 to 11.3 × 10 −11 Sv/Bq for HTO and OBT, respectively. Similarly, the dose coefficient for a
one-year-old was found to vary from 2.9 to 6.6 × 10 −11 Sv/Bq and 14 to 18 × 10 −11 Sv/Bq for HTO
and OBT, respectively, when the radiation weighting factor was set to unity.

E. Summary of biokinetic and dosimetic models
168. HTO behaviour following intake into the mammalian body is well understood with regard to both
early distribution in the body followed by excretion as part of water exchange (half-time of adulthumans 4–18 days) and simultaneous conversion of a small fraction (of the order of 1%) of tritium into
OBT with subsequent longer term excretion (half-time of adult humans 23–104 days). The averagecontribution of OBT to effective dose of adults derived from 17 human observations was 3.0%. The

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data on longer term processes with half-times of more than 100 days are highly uncertain and needfurther study. The ICRP biokinetic model for HTO for all age groups is moderately conservative, withan OBT contribution to dose equal to 9% for adults. The biokinetics of HT following inhalation is also
well understood. However, the current ICRP model considers only HTO formed in vivo and does nottake account of an additional component of dose to cells of the alveolar-interstitial region of therespiratory tract.
169. The ingestion of food containing OBT by members of the public results in the degradation of a
substantial fraction of the OBT to HTO in the gastrointestinal tract and in body tissue by catabolic
reactions, and uptake and conversion of the rest of the OBT into tissue macromolecules. Both thefractions and subsequent excretion rates of OBT from tissue depend on food origin and composition.
The ICRP model for OBT ingestion by humans is generally consistent with experimental results.
However, it does not consider the non-uniform deposition of OBT between various organs and tissues.
Physiologically-based OBT models, such as the ones proposed by Richardson and Dunford [R8] andGaleriu and Melintescu [G1], consider non-uniform deposition. Dose coefficients calculated using thesemodels are generally similar to ICRP values.
170. Tritiated biochemical substrates, such as glucose, amino acids, hormones, DNA and RNA
precursors, may be directly incorporated into organic molecules in body tissues if absorbed to bloodand transported to sites of active metabolism within cells. The extent of incorporation of tritium into
specific forms of OBT is determined by such factors as the chemical compound containing tritium,
its isomeric form, position of the label in the molecule, and the amount of carrier. OBT formed fromtritiated precursors of biological macromolecules is retained in tissue longer than HTO. Catabolismof labelled compounds will result in tritium being partially oxidized and entering the body water as
HTO or catabolized and excreted as low molecular weight organic substances. Following intake oftritiated precursors of biological macromolecules, the internal dose to mammal tissue is usually up to
ten times larger than the dose from intake of an equal amount of HTO, and the contribution of OBT
to dose may dominate.
171. Labelled DNA precursors (e.g. 3H-thymidine, 3H-desoxycitidine) entering the mammalian bodyby various routes are partially degraded to HTO and partially incorporated into the DNA of dividing
cells, and thereafter selectively expose nuclei of proliferating cells to beta radiation. 3Hdeoxynucleosides are preferentially incorporated into the proliferating cell systems of embryo andfetus, bone marrow and small intestine at any age and, to a substantially lesser extent, in tissue withlower frequencies of mitoses (e.g. muscle, liver). For both acute and prolonged intake of tritiated DNA
precursors, the absorbed dose to nuclei of proliferating cells may be larger by one to two orders ofmagnitude than the dose from intake of equal amounts of HTO. As the average range of tritium beta
radiation is considerably less than the dimensions of the nuclei of mammal cells, the use of averageorgan or tissue dose in the case of incorporation of 3H-nucleosides requires careful consideration. The
ICRP dose coefficients for OBT should not be directly applied to intake of tritiated DNA precursors.
172. Tritium particles of low solubility (e.g. metal tritides, luminous powder) inhaled by workers in
occupational conditions partially deposit in the respiratory tract and may be retained in the lungs forlong periods. Material-specific labile fractions of tritium segregate from particles as HTO or as organic
molecules that are absorbed to blood and excreted. Effective doses from inhalation of low soluble
tritium particles can be assessed by means of the ICRP human respiratory tract model combined with
the HTO model. Tritium as HTO or biochemical substrates is easily transported through the placenta to
the embryo and fetus and secreted in maternal milk. In conditions of chronic tritium intake by themother, committed internal doses incurred by the embryo and fetus and by the suckling infant correlatewith the dose of the mother and are not substantially different in magnitude.

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VI. BIOLOGICAL AND HEALTH EFFECTS

A. Non-radiological effects of tritium in biological systems
1.

Transmutation

173. Transmutation is the conversion of one element into another through radioactive decay. Whentritium undergoes decay, it becomes helium-3 (3He), a stable, inert gas. Helium is chemically very
different from hydrogen and, therefore, this could make a significant contribution to the effect oftritium when organically combined. If a tritium atom is bound to a DNA molecule when it decays, mostof the kinetic energy will accompany the beta radiation as it is ejected from the nucleus, but someenergy will provide a kick-back to the 3He atom as recoil energy. Kacena [K1] determined that therecoil energy was too small (up to 3 eV) to cause ionization of the DNA molecule on its own. However,
the resultant 3He atom would break free from the DNA molecule. In complex molecules, the effect of
this conversion to a positively charged carbonium ion would be difficult to distinguish because of theproximity to the deposition of energy and associated events from the beta emission.
174. Myers and Johnson [M26] and Gracheva and Korolev [G8] performed comprehensive reviews oftransmutation effects. They noted that the degree of damage caused by transmutation of tritium into
helium could theoretically vary significantly, depending on the position of the tritium atom in specific
DNA nucleotides. The studies covered several test systems in the S13 virus, in two strains of thebacterium E. coli, in Drosophila Melanogaster (fruit fly) and in cultured mammalian cells. On the basisof the position of the tritium in the nucleic acid, varying degrees of damage were observed inexperiments. The most pronounced mutagenic effect was detected with cytosine-5-3H, which in the caseof tritium decay converted to uracil. In some simple biological systems (virus, E. coli), thattransmutation resulted in an elevated mutation rate that was 3–400 times larger than the mutation ratecaused by beta radiation only [G8]. The reviewers argued, however, that the increase in the mutation
rate in mammals (resulting from transmutation) would not likely exceed 5% of the normal rate and that
this was too small to be detected.
175. Carsten [C11] discussed the possibility that such effects would be manifest in humans afteringesting HTO or OBT as food. He suggested that the risk was small enough to pose no significant
hazard, primarily because only 2% of the hydrogen atoms in DNA were located at the 5-position of thecytosine ring and damage would be minimal. Feinendegen and Bond [F3] reached the sameconclusion—that “the effects of intracellular tritium are overwhelmingly due to beta irradiation of thenucleus” and “transmutation effects do not produce a measurably increased effect under mostconditions.” The United Kingdom Advisory Group on Ionising Radiation (AGIR) [H16] came to asimilar conclusion. If DNA damage in mammals did occur from transmutation, it is unlikely that it
could be distinguished from radiation-induced damage and thus would be already accounted for inmeasured RBE values.

2.

Isotopic effects

176. While chemically similarly to hydrogen, tritium has slightly different physical properties due toits increased mass. Diabaté and Strack [D5] noted that synthetic reaction rates decreased as atomic mass

288 UNSCEAR 2016 REPORT

increased, causing a significant isotopic effect of OBT depletion with tritium compared with HTO fromwhich biological macromolecules are synthesized. In contrast, once fixed to carbon, C-T bonds are cleft
more slowly than C-H bonds. In equilibrium conditions of OBT synthesis in plants grown in a medium
with HTO, the isotopic ratio of tritium specific activity in hydrogen of OBT in bulk organic matter tothat in water was in the range of 0.6–0.8 [D5].
177. Intracellular discrimination between tritium and hydrogen atoms due to isotopic differences (inparticular a mass ratio of 3) have been considered in the past (see e.g. [C11, M2]) and concluded to beof little consequence for the risk of tritium. However, contrary to the traditional view of OBT inbiomatter, experiments that used denaturing agents have suggested that a proportion of tritium may bedesignated as “buried tritium”, a tightly bound HTO [B13]. In such a fraction, the buried tritium inbiomacromolecules, such as proteins, is in positions where the exchange rates are substantially reduced
as a consequence of the three-dimensional structure that arises upon the “folding” of thesebiomacromolecules. The hydrogen bridges between the molecules of water are stronger than between
organic configurations, resulting in accumulation of tritons both inside the biopolymers and within theirprimary hydration shields. There is an enrichment of tritium in the newly identified buried hydrogenbonds compared to the free water in the cell. In most biomolecules, the enrichment may be 1.4-fold but
in DNA, where the hydration shell consists of 11 molecules per nucleotide and is not readily permeable
to ions, the enrichment in the water trapped in the core may be twofold. While this will certainly resultin slightly more beta tracks originating from HTO within and around the DNA, it remains true that thevast majority of beta tracks encountered by the DNA will have originated from HTO outside the DNA
since that is where most of the HTO is situated. The effect on radiation dose to the DNA will therefore
be small but may increase the RBE value in experimental determinations.

B. Deterministic effects
1.

Lethality

178. Brues et al. [B26] investigated the lethality of tritium by giving mice single injections of
0.126-8.4 GBq of HTO. The dose that killed 50% of the population within 30 days (LD50/30) was about
9 Gy, which corresponds to an initial activity concentration of tritium of 37 MBq/g of body weight(BW). Furchner [F13] reported an LD50/30 of 8 Gy, for HTO given in a single intraperitoneal injectionof 33 MBq/g to CF1 female mice. Yamamoto et al. [Y2] (see also [Y6]) reported an LD50/30 of about
8 Gy after a single injection of 0.56 GBq in C57BL/6N female mice and about 13 Gy with an injectionof 0.93 GBq in female (C57BL/6N × C3H/He) F1 mice. Overall, these data suggest a LD50/30 dose of
the order of 10 Gy.
179. Yamamoto et al. [Y2] also investigated the lethality of a continuous oral administration of HTO
as drinking water in (C57BL/6N and C3H/He) F1 female mice. The tritium concentration reached aplateau in organs and blood after about seven days. Haematopoietic death typically occurred after two
weeks following ingestion of HTO with activity concentrations from 0.15 to 0.6 TBq/L in drinking
water. The lowest absorbed dose to cause death was estimated to be about 11 Gy from the continuous
ingestion of 0.15 TBq/L of HTO.
180. The dose from tritium necessary to cause death in mice appears to be similar to that from acuteexternal irradiation by X-rays or gamma rays, with tritium LD50/30 values of about 6–9 Gy,
corresponding to an acute intake of the order of 1 GBq of HTO. In the literature, there are somepublished historical cases of radiation sickness of workers caused by tritium [M15, O3, S12], including
two lethal cases.

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181. In 1963, a worker in a radiological laboratory in the former Soviet Union had an intake of about
350 GBq of HTO [O3] due to violation of safety rules. Monitoring of tritium in urine started 25 days
later, when the worker had visited a doctor because of deteriorating health. Tritium concentration in
urine declined with a half-time of seven days and the committed equivalent dose in soft tissues wasassessed to be about 12 Sv. Treatment started immediately after visiting the doctor. Pronounced
radiation sickness symptoms were observed during 1.5–2 months after the accident. After about
1.5 months, regeneration of the haemopoietic system began but substantial recovery of leukocyte andneutrophil concentration in blood took more than three months. The patient returned to work after
six months. Medical surveillance over 2.5 years showed recovery of working capacity and the absenceof substantial adverse effects in internal organs.
182. In the late 1950s and early 1960s, several workers at two European facilities for production oftritiated luminous compounds used large quantities of tritium (4–10 TBq per annum) [M15, S12]. In thecourse of chemical operations, each worker incorporated substantial amounts of tritium. Measurementsof tritium in urine started in the 1960s and were used for dose assessment. In case A-1, tritium
concentrations in urine varied in the range of 2–40 MBq/L. From these data, the equivalent dose in soft
tissue was assessed as 3–5 Sv over the four years preceding death [S12]. An assistant to A-1 (A-2)
received a dose that was about half of that received by A−1 and stopped working with tritium after she
showed moderate anaemia. Workers A-3 and A-4, who succeeded A-1 and A-2, worked underimproved conditions and incurred lower doses without health effects. In case B-2, the dose incurred
over three years before death was assessed from tritium measurements in urine in the range from a few
sieverts up to 20 Sv. In both cases A-1 and B-2 there was substantial contribution to dose from OBT, as
indicated by measurements made on autopsy samples, and both workers died with clinical signs of anaplastic pan-myelocytopenia complicated by pulmonary or other symptoms [S12].

2.

Effects of HTO on animal embryo and fetus

183. Straume and Carsten [S23] reviewed the current literature on exposure to tritium during fetal
development. During particular periods of development, some fetal cells will be dividing rapidly anddifferentiating to form tissues and organs, while other cell types may be showing very little or no cellproliferation. Tritium that is incorporated into OBT of low proliferation cells could result in larger
integrated doses to these cells, since the tritium would not be diluted by further cell proliferation.
However, uptake will be dominated by those cells undergoing rapid division. Commerford et al. [C26]reported that the dose from tritium incorporated from HTO into macromolecules, such as DNA andhistones, of rapidly dividing cells was small compared with that from the same activity of tritium in theform of HTO.
184. In a study by Laskey et al. [L7], rats were continuously fed HTO with activity concentrations of
0.37–370 kBq/mL from conception of the first generation until delivery of the second. The
corresponding dose rates were 0.03–30 mGy/d. Exposure to HTO with an activity concentration of
370 kBq/mL resulted in a 30% weight reduction of the testes in the first generation of adult males, butthere was no impairment of overall growth or reproductive ability. In the second generation of
newborns, decreases in the weight of the brain, overall body weight and litter size were found. An
increase in resorption for rats exposed to HTO with an activity concentration at 370 kBq/mL was also
found. However, Laskey et al. [L7] did not observe any effects on litter size or resorption at or below anactivity concentration of 37 kBq/mL (which corresponds to a dose rate of about 3 mGy/d).
185. Bursian et al. [B27] assessed the effects of continuous exposure of rats to HTO from conception
to birth. The activity concentrations used were 0, 37, 370 and 3,700 kBq/mL. In utero exposure to doses

290 UNSCEAR 2016 REPORT

as low as 0.66 Gy (corresponding to the highest concentration used) produced measurable andpersistent decreases in brain weight and increases in norepinephrine concentrations at 21 and 45 days
after birth. No differences from the controls were observed in the rate of turnover or the concentrationsof dopamine, acetylcholinesterase or monoamine oxidase.
186. Jones et al. [J8] gave pregnant squirrel monkeys water with tritium levels ranging upwards from
2 kBq/mL throughout gestation. No effects on body weight, body dimensions, organ weights,
haematological patterns, or the histology of organs or tissues, with the exception of ovaries, wereobserved in newborn progeny. However, the number of primary oocytes in female progeny decreased
markedly with increasing levels of HTO in maternal drinking water.
187. Yamada et al. [Y1] studied the effect of prolonged in vitro exposure to HTO and 60Co gamma
radiation on pre-implantation mouse development after mating female C57BL/C3H F1 and male ICR
mice. With the development to blastocyst as the end point, the LD50 was 4.4, 8.5 and 15.8 MBq/mL
corresponding to beta radiation dose rate of about 10, 20 and 40 mGy/h, respectively, for pronuclear,early two-cell, and late two-cell embryos, respectively. Compared to 60Co gamma radiation, the RBE
value of tritium beta radiation was in the range of 1.0–1.7.
188. The effects of tritium on the morphogenesis and development of rats and mice have been studied byseveral authors. Wang and Zhou [W3] reported modifications in the cognitive function of young rats bornto mothers that had been injected with HTO on the thirteenth day of pregnancy that resulted in 0.1 and
0.3 Gy of in utero exposure. Gao et al. [G2] showed decreases in cognitive behaviour and a significant
decrease in hippocampal pyramidal cells in the brain’s CA1 area with intraperitoneal injections of HTO
on the thirteenth day of gestation that resulted in 0.09 and 0.27 Gy of in utero exposure. Sun et al. [S25]reported a decrease in the brain weight of mice exposed on the thirteenth day of gestation to a dose of
0.4 Gy (from the injection of HTO of 964 kBq/g of body weight); both the thickness of the somatosensory
cortex and pyramidal cell density were significantly decreased at this dose. Jain and Bhatia [J2] also
observed pathological changes in the cerebellum of mice following exposure of the mothers to HTO.
An initial injection of HTO was given at 17 days post-conception followed by additional intake of HTO
of 111 and 11.1 kBq/mL. The observed damage was dose dependent.
189. Zamenhof and van Marthens [Z3] studied how five generations of rats were affected by pre- andpost-natal exposure to HTO. Female rats were given water containing HTO with an activity
concentration of 111 kBq/mL beginning in adolescence and continuing throughout pregnancy. This
exposure to tritium did not produce any signs of radiation-induced disease in the mothers. The courses
and outcomes of pregnancy were also normal, but 60% of the newborn rats exhibited haematomas,
oedemas and subdural haemorrhages. None of these effects lasted beyond 30 days. Zamenhof and van
Marthens [Z4] subsequently found that there were decreases in the weight and DNA and protein
content of the brain in later generations of rats continuously drinking water containing HTO with anactivity concentration of 111 kBq/mL (estimated average daily intake for a 60 g rat—the weight of a
30-day-old—was 2.1 MBq, corresponding to an absorbed dose rate of about 1.4 to 8 mGy/d). The
effects were generally most evident in all generations (F2–F5) except the F1 generation. In addition, all
but the F1 generation showed some recovery over time.
190. In summary, radiation-induced effects on the embryo or fetus (principally neurological or
reproductive tissue impairment) have been demonstrated in laboratory animals exposed to tritium. Sucheffects start occurring at chronic intake of activity concentrations of HTO of about 50–100 kBq/mL ofbody water and are consistent with similar effects from external photon irradiation. Embryo–fetal
effects in animal studies are observed at doses of about 0.4–0.6 Gy from chronic intake of HTO.

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291

Immunological effects of HTO

191. The Committee has conducted a number of reviews of the effects of exposure to ionizing radiationon the immune system, the most recent being given in its UNSCEAR 2006 Report, annex D [U12]. The
major conclusions from that review were the following:−“High doses of radiation produce immunosuppression mainly through destruction of cells.
Lymphocytes are very radiosensitive, and their destruction is currently used as an early
indicator of the level of an accidental acute exposure. Radiation-induced changes in immune
parameters seem to be more dependent on total dose than on dose rate. Persisting effects on
immune system have been observed after exposure to ionizing radiation.

−

At low doses and dose rates, the effects of ionizing radiation on the immune system may besuppressive or stimulatory. The long-term effect of low radiation doses on the immune
function in relation to human health needs to be further evaluated.”

192. However, very few studies specifically reference exposure to tritium. Some reports have comefrom studies of workers exposed occupationally and also from experiments carried out in the in vitro orin vivo systems.
193. Tuschl et al. [T17] investigated some immunological parameters in ten NPP workers exposed
during a four-week period to external radiation (total effective doses ranged from 1.4 to 9.8 mSv) andtritium inhalation (committed effective doses ranged from 1.2 to 2.8 mSv). Twenty-five days after the
beginning of the exposure, only the CD4+/CD8+ T cell ratios were markedly elevated compared to theratios detected in non-exposed controls and the effect was mainly due to an increase in the absolute
numbers of CD4+ T helper cells. In five exposed subjects who agreed to give blood samples fivemonths after the first sampling, the CD4+/CD8+ T cell ratios were still elevated.
194. Milacić [M14] analysed blood samples from 53 workers exposed to tritium. The workers wereseparated into three groups depending on the duration of occupational exposure: 0–5, 6–15 and16-30 years. The groups were compared with each other and with a control group. Effective doses were
not provided, but the average tritium activity concentration in urine of the exposed subjects was
1.9 kBq/L. Although total leukocyte counts did not differ from the control group, lymphocyte andeosinophil counts were higher in the workers exposed to tritium and varied with the duration of the
exposure, being the lowest in the 6–15 years group. Chromosomal aberrations were detected in 49% ofworkers exposed for 10.5 years and with a significantly higher average tritium concentration in urine(3.5 kBq/L). However, the average tritium activity concentration in the exposed subjects with noaberrations was 0.35 kBq/L. Alkaline phosphotase and myeloperoxidase activities in granulocytes weresignificantly lower in all exposed workers. The author interpreted the increase in lymphocyte counts asa stimulation of the immune system by tritium and in eosinophil counts as a compensatory reaction of
the bone marrow in response to impaired function of granulocytes. The workers had no clinicalmanifestations of immunity disorders.
195. In 1980, Kirillova and Luzanov [K16] compared prolonged exposure of CBA mice to HTO andexternal gamma radiation (from 137Cs) delivered in equal daily radiation doses (cumulative dose
4.1 Gy). They observed a decrease in the immune response. The antibody and rosette-forming capacityof spleen cells was lower in the group exposed to HTO than in the group exposed to gamma radiation.
On the basis of the total immune response observed over the entire treatment period, they concluded
that HTO was 1.27 times more efficient in decreasing the immune response than gamma radiation.
They suggested that the observed diminution of the immune response was due mainly to damage to thelymphoid tissue and also to disruption of haematopoiesis. Later on, Kirillova [K19] conducted a similar292 UNSCEAR 2016 REPORT

experiment, but used rats instead of mice. After a prolonged intake of HTO, the decrease in normal
killer cells (NK) 2 activity in rats was less and the recovery was more rapid than observed in mice forthe same total dose (about 8 Gy). The authors linked this observation to the higher metabolic rate inmice compared to rats.
196. In another study of changes in the concentration of NK cells as an end point, Kirillova [K17]treated mice to HTO at an activity concentration of 0.37 MBq/g of body mass over six months andcompared them with a non-treated group of mice. The exposure to HTO resulted in a decrease in theactivity of NK cells by 35−45% in the treated mice when compared with the reference group (p <0.05)
at three months after administration of HTO had ceased. Six months later, the NK cell function had
recovered and was even higher by a factor of 1.8 in the mice that had been treated with HTO than in thecontrol group.
197. Kirillova et al. [K18] also studied the quantitative and qualitative degree of recovery of theimmunological system in mice after prolonged exposure (throughout life; around 725 days for thecontrols and 650 days for the treated mice) to HTO with an activity concentration of 0.37 MBq/g ofbody weight giving a total dose of 8.7 Gy. The authors observed a depopulation of cells in bonemarrow, spleen and thymus, which persisted until the end of the lives of the animals. They suggested
that the disruption in immunological response was caused early in the treatment by a reduction in thenumbers of lymphocytes and by a lower activity of B cells and tritium helpers and, later after radiation
exposure, by the impairment of tritium helpers function. There was a direct relationship between the
immunodeficiency and the dose rate and the total absorbed dose of beta radiation.
198. Smirnov et al. [S14] concluded that prolonged exposure of CBA mice to daily intake of HTO of
0.19, 0.37 and 0.74 MBq/g of body weight up to doses ranging from 0.2 to 1 Gy, resulted in an
impairment of humoral immunity at various stages of the immunopoesis at all dose levels. They also
showed a direct relationship between the depopulation of colony forming units (CFU) (early precursors
of T and B lymphocytes) and dose rate. They concluded that the production of antibody forming cells
was a function of the absorbed dose and that a prolonged exposure to HTO resulted in impaired
humoral immunity at various stages of immunopoesis. They also demonstrated that the mechanisms
generating the immunological response were highly sensitive to the tritium beta radiation.
199. As demonstrated more recently by Umata et al. [U4], a single intraperitoneal injection of HTO to
C57BL/6N mice to give a total whole-body dose of 3 Gy significantly increased the number of variantsof the T-cell receptor expressed on splenocytes. The frequency of apoptotic cells of the spleen 12 hours
after HTO injection increased to 5.0%.
200. In summary, experimental studies carried out in animals indicate that prolonged exposure ofanimals to HTO associated with relatively high radiation doses (in the range of 1−8.7 Gy) resulted insome diminution in the immune response in animals which, depending on the dose, could be reversed.
In contrast, the few studies carried out in workers occupationally exposed to low doses of tritium
demonstrate that such exposure can stimulate immune functions, but also increase the frequency ofchromosome aberrations in lymphocytes. However, it is not clear whether such effects, even at highdoses, had any significant health consequences. More research on the effects of acute and protracted
low-level exposure to tritium on the immune system is warranted.
2

A cytotoxic lymphocyte.

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293

Germ cell effects of HTO(a) Effects in females
201. Because tritium distributes itself throughout the body, it can be taken up by developing oocytes
and incorporated into DNA. Tritium incorporated into human oocyte DNA could theoretically irradiate
the oocytes over 30 or more years. Because oocytes do not divide until fertilized, there is little turnoverof the DNA molecules, which implies that the biological half-time of the tritium embedded in oocyte
DNA could approach the radioactive half-life of tritium of 12.3 years. Forell et al. [F8] pointed out that
if a biological process were to gradually exchange all the components of DNA molecules—includingthat of tritium-labelled DNA—it would take 50 years to replace 2–5% of a cell’s genome (DNA). The
author estimated that in resting lymphocytes 2,000 bases per hour were turned over. Therefore, most of
the tritium incorporated into an oocyte will remain there for its whole life. This issue was also discussed
by the United Kingdom Health Protection Agency [H16] in relation to the discharges from aradiopharmaceutical plant. They concluded that tritium could be incorporated during pregnancy into the
DNA of fetal oocytes and remain there until fertilization decades later.
202. Straume and Carsten [S23], in reviewing the effects of radiation exposure on oocytes, reported
that most of the information on radiosensitivity in humans came from autopsies of women who had
been exposed to substantial doses of external radiation [L13], and from fertility histories of women who
had undergone radiotherapy or had been exposed as a consequence of the atomic bombings in Japan[B5, L14, U5]. In all cases, the data were for short-term external exposure of adult women.
203. In female mice, data suggest that premature oocytes are more radiosensitive than mature oocytes[B3]. This is in contrast to results in human females that demonstrate that from 3 days after birth, alloocytes are equally radiosensitive. In women, exposure of 2.5−6 Gy to X- or gamma radiation will leadto permanent sterility [I11, I20]. There is no temporary sterility in human females as there is in mice.
204. Dobson and Kwan [D6, D7] continuously exposed non-inbred Swiss-Webster mice to HTO and60
Co gamma rays during the early period of oocyte development from conception until 14 days after
birth. Oocyte survival following tritium exposure decreased exponentially with dose rate with no
threshold; the LD50 level was 0.074 MBq/mL of body water corresponding to 4.4 mGy/day. Exposure to60
Co gamma rays was shown to be less effective at the same dose and did not follow an exponential
relationship. As a result, the RBE value of tritium beta radiation, compared with gamma radiation,
increased with decreasing dose and dose rate from about 1.6 at gamma-ray dose of 0.4 Gy to about 3 atlower doses. In studies of both mice and rats, Satow et al. [S4, S5] studied the effectiveness of exposure
to tritium in killing immature oocytes. Statistically significant oocyte resorption was found at activity
concentrations of HTO of 0.34 MBq/g of body weight and more, corresponding to a total dose of
77 mGy and more.
205. A number of animal studies have investigated the continuous administration of HTO or tritiatedthymidine throughout pregnancy and analysis of the subsequent effect on oocytes in the offspring. Thismethod of continuous administration in rats was pioneered by Fliedner et al. [F7] and Schreml et al.
[S7] and subsequently adapted by Lambert and Phipps [L6] using mice. Using this method, Haas et al.
[H1] studied the effect of HTO infusion in utero on post-natal oocyte development in rats. During thefirst 21 days of life post-conception, a dose dependent reduction in oocyte numbers was observed. At
birth, 54 MBq infused had reduced the numbers by 50% whereas 215 MBq produced total aplasia ofoocytes. They suggested that tritiated thymidine was about ten times more effective than HTO for thiseffect. In a subsequent paper [S8], these authors concluded that this factor was about 3.7 in relation toradiation dose to the oocyte cell nucleus.

294 UNSCEAR 2016 REPORT206. Lambert and Phipps [L6] exposed pregnant SAS/4 mice to HTO in drinking water and, byconstant infusion, to tritiated thymidine throughout pregnancy. A number of parameters were studied in
the offspring including oocyte survival at 14 days of life. They concluded that for this parameter
tritiated thymidine was about two–three times as effective as HTO in causing oocyte lethality.
207. Pietrzak-Flis and Wasilewska-Gomulka [P8] also studied the effect of constant intake of HTO ortritiated food on oocyte survival in Wistar rats from birth and sampled at ages 21 and 71 days. They
found that tritiated food was more effective at reducing oocyte numbers.
208. In summary, the female reproductive system shows great discrepancies between mice andhumans. For humans, the reproductive system is radiosensitive to a dose of 2.5−6 Gy given in a single
fraction. Exposure to doses within that range will cause permanent ovarian failure due to the killing ofoocytes, and will be accompanied by features associated with menopause. It is anticipated that such adose of tritium in the ovaries would have a comparable effect.

(b) Effects in males
209. Unlike oocytes, spermatogonia are continuously produced from stem cells throughout adult life.
Like all tissues that are rapidly replaced, there are certain germ-cell stages that are highly sensitive tocell killing by ionizing radiation. Experiments in mice conducted by Oakberg in 1955 and 1959 showedthat the most sensitive cells are the type A and B spermatogonia, which can be reduced by 50% withdoses of only about 0.3 Gy of acute X-rays (reported in [S23]). The spermatid and spermatozoa stages
are much less sensitive than the spermatogonia stage (Oakberg and Clark, 1964 as reported in [S23]).
Lambert [L3] found a 27% reduction in spermatogonia of mice injected with tritiated thymidine at an
activity concentration of 0.19 MBq/g of body mass and with HTO at an activity concentration of
2.2 MBq/g of body mass (dose to the cell nucleus of 84 mGy and 49 mGy, respectively).
210. Carr and Nolan [C9] studied the reduction of testis mass in mice following single injections oftritiated thymidine (0.037–0.74 MBq/g of body mass) or HTO (0.37–1.48 MBq/g of body mass) and60
Co gamma rays (delivered to match the dose-rate vs. time curve in the 1.48 MBq/g HTO group). The
radiobiological effect was investigated at times from one hour to 24 weeks after injection.
Measurements of the testicular retention of tritium were also made at these times. There was aprogressive loss in mass, up to 30% after 4–5 weeks, followed by an irregular recovery, which wasmore delayed in the case of the animals injected with tritiated thymidine. Time-integrated fractional
testis mass loss was a linear function of injected HTO activity. The RBE value of HTO compared with60
Co gamma rays was 1.43 at gamma-radiation dose of 0.6 Gy. As only one gamma-radiation dose level
was used, RBE dependence on dose could not be assessed.
211. Balonov et al. studied reduction of testis mass in mice following single intraperitoneal injection of
HTO (0.4–12.6 MBq/g of body mass) or exposure to 137Cs gamma radiation delivered during ten days
with exponential reduction of dose rate with a half-time of 2.5 days similar to HTO excretion rate
[B10]. The range of tritium beta-radiation dose was 0.12–3.4 Gy and that of gamma-radiation dose was
0.25–3.7 Gy. Statistically significant relative testis mass reduction was observed in the dose range from0.25 Gy of both tritium beta radiation and gamma radiation. RBE values in terms of relative testis mass
reduction increased from 1.9 at mass reduction of 40% (dose of tritium beta radiation 0.8 Gy) to 2.2 atmass reduction of 10% and corresponding lower dose of about 0.2 Gy. In summary, it has been shownby studies using both external radiation and tritium administration that certain stages of spermatogonia
development are particularly sensitive to radiation. Most studies concluded that a ratio of effects(mostly lethal) compared to gamma radiation was in the range of 1.4−2.2.

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295

C. Stochastic effects of HTO in mammals
1.

Carcinogenicity

212. Many laboratory studies on animals have demonstrated that exposure to tritium, both as HTO andtritiated compounds, can induce cancer although the carcinogenic effect of exposure to tritium has notbeen studied as extensively as that of gamma radiation and X-rays.
213. Cahill et al. conducted two studies involving the administration of HTO to pregnant SpragueDawley rats to term in a range resulting in whole-body doses during gestation of 0.066–6.6 Gy [C1,C2]. In their first study, increased incidence of mammary fibroadenomas was detected in dams exposed
at 3.3 Gy and 6.6 Gy, but not at lower doses. In the second study, offspring surviving beyond 30 days
were observed throughout their life and neoplasia recorded. Intrauterine exposure to doses of up to
0.66 Gy had no significant effects on overall cancer incidence rate or onset of mammary
fibroadenomas. In addition, females exposed in utero to 3.3 or 6.6 Gy had lower incidence rates ofmammary fibroadenomas and at 6.6 Gy females had a lower incidence of overall neoplasia compared tothe control unexposed rats. These females, however, were sterile and had reduced mean life spans.
214. Seyama et al. [S13] reported on a series of studies involving acute intraperitoneal injections ofrelatively high levels of HTO to 7–8-week-old female (C57BL/6N × C3H/He) F1 mice at the activities
resulting in internal whole-body doses from 2.0 to 10.5 Gy. The animals were observed for up to
750 days and cumulative neoplasia was compared to the effects of chronic irradiation by either gamma
rays or fission neutrons. The effect (incidence of cancer) seems to have nearly saturated at the lowest
dose level; that is, the total incidence of tumours was similar at 500 days and later in all exposed groups(about 80%) in contrast to the control group (less than 5%). The authors also studied induction ofthymic lymphoma in mice that received 7.9 or 10.5 Gy from a single intraperitoneal injection withthose that received equal doses in four subsequent injections with weekly intervals. In the latter case,the latent period was much shorter and lifetime lymphoma incidence was significantly higher than after
a single injection.
215. Yamamoto et al. [Y3] reported a study involving continuous oral administration of five levels of
HTO to female (C57BL/6N × C3H/He) F1 mice from 10 weeks of age resulting in a dose rate to soft
tissues of 0.01–0.24 Gy/day. Lifetime tumour incidence approached the maximum level (83%) alreadyat the lowest dose rate (0.01 Gy/day) while spontaneous incidence was 54%. Exposure to larger dose
rates accelerated development of most studied tumours, which resulted in substantial life shortening.
The main cause of death of mice exposed to higher dose rates (0.096−0.24 Gy/day) was thymic
lymphoma. At lower dose rates, non-thymic lymphomas and solid tumours were also observed. Bothshortening of the latent period and life shortening due to development of tumours significantly
increased with dose rate of tritium beta radiation and the increases were significantly larger than thoseafter irradiation of mice with X-rays or gamma radiation from a 60Co source. In a later similar study[Y4], the same authors exposed mice to three lower levels of HTO in drinking water from ten weeks ofage resulting in a dose rate to soft tissues of 0.0002–0.0036 Gy/day. The life span was discernibly
shortened in the group with a larger dose rate (0.0036 Gy/day) due to shortening of latent periods fortumour development. In the groups of mice with lower dose rates, both this effect and incidence ofthymic lymphoma were missing.
216. A study by Johnson et al. [J5] estimated the lifetime incidence of myeloid leukaemia in sevengroups of about 750 CBA/H mice each; radiation exposure was approximately 0, 1, 2 and 3 Gy both for
HTO and for X-rays. The lifetime incidence of leukaemia in these mice increased from 0.13% in the296 UNSCEAR 2016 REPORT

control group to 6–8% in groups exposed to higher radiation doses. The results were fitted to various
equations relating leukaemia incidence to radiation dose, using both the raw data and data corrected for
cumulative mouse-days at risk. The calculated RBE values for tritium beta rays compared to X-rays
ranged from 1.0±0.5 to 1.3±0.3. A best estimate of the RBE value for this experiment was about 1.2±0.3.
217. Gragtmans et al. [G9] estimated RBE for tritium radiation in reference to 200 kVp X-rays, using
acceleration of breast tumour appearance in the female Sprague-Dawley rat as the end point. Chronic
X-ray doses of 0.3–2.0 Gy were delivered over ten days. Intraperitoneal injections of HTO ranging inconcentration from 45 to 370 MBq/100 g of body weight were administered, followed by four
additional injections at two-day intervals and half of the initial concentration. RBE estimations werebased on various criteria including the tumour incidence per Gy at 450 days post-irradiation and thetime required to induce tumours in 50% of the animals at risk. The results suggest that tritium beta rays
are about 1.1–1.3 times more effective than chronic 200 kVp X-rays for acceleration of the appearanceof rat mammary tumours. However, the uncertainties involved in these calculations are such that theeffects of tritium beta rays could not be reliably distinguished from those of chronic 200 kVp X-rays.
218. Revina et al. [R1] described a study in rats which were administered HTO intragastrically fivetimes a week for six months. The effects (leukaemia and other cancer) on these animals were comparedwith those on a group chronically exposed to gamma radiation delivered in daily doses comparable tothe tritium dose rate. The main problems with this study are the high doses and the fact that there wasonly one dose point in the tritium and gamma exposed groups. An estimate of RBE of about 2.5 wasmade in this study.
219. Intraperitoneal injection of HTO to male N5 mice several times over 30 days (so that the exposure
period comprised all stages of spermatogenesis) to give a total dose of 1.5 Gy resulted in a statisticallysignificant increase of leukaemia incidence among their young (less than 210-day-old, but not yet
one-year-old) offspring [D1]. It appeared that the overall leukaemia incidence in the offspring of the
HTO-exposed fathers was significantly dependent on the maturation stage of the sperm-forming cells
during the HTO exposure. However, in a paper by Balonov et al. [B12] which summarizes Russianstudies on tritium carcinogenicity. Mice and rats were given HTO in drinking water at a dose range of
0.24−25.3 Gy. Although most malignancies were increased, the absence of a positive dose–response
and occasionally a negative response makes interpretation of these data difficult.
220. Yin et al. [Y5] used 12-day-old male and female pups of C3H/HeN mice that were given a single
intraperitoneal injection of HTO at the activities of 0.23, 0.92 and 3.70 MBq/mouse and then observedfor 14 months for the development of tumours. In the males, a significantly increased incidence of liver
neoplasms was detected whereas in the females, only an insignificantly elevated incidence of ovariancancer was observed in mice exposed to the highest concentration of HTO.
221. Studies in which administration of tritiated material was compared with the effects of similarly
protracted X- or gamma-radiation exposure are the most reliable sources of RBE data. This is importantin the absence of such data for human exposure. Four of the more comprehensive studies [G9, J5, R1,
S13] were critically examined in the AGIR report [H16] and the conclusion was that, due to
deficiencies in the experimental design and statistical analysis, the findings from these studies shouldbe treated with caution. However, taken together, they indicate an RBE value with a central estimate in
the range 0.8–2.5 with an upper 97.5 percentile value of no more than about 3.
222. Straume [S22] undertook a literature review of the risks, including cancer induction, fromexposure to tritium. Because information was not available for humans, cancer-risk estimates fortritium were derived from experimental animal (mostly mouse) studies. Straume calculated a skewed
risk distribution (with a fiftieth percentile risk per unit dose of 81 × 10−6 mGy−1 with a 90% confidenceANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

297

range of 38–185 × 10−6 mGy−1) using Monte Carlo methods and distributions of dose-rate effectivenessand multiplying by best estimate RBEs for tritium (based a central value of the RBE of 2–3 ranging up
to 4.5). This was comparable to radiation risk estimates in the Committee’s UNSCEAR 1988 Report[U6], in the ICRP Publication 60 [I13] and in the BEIR V Report of the Committee on the Biological
Effects of Ionizing Radiation [N6].

2.

Heritable effects

223. In its UNSCEAR 2001 Report [U10], the Committee noted that no radiation-induced hereditarydiseases had (to that date) been demonstrated in humans. Nonetheless, since such effects are seen inplants and animals, the Committee provides an approach for estimating such risks [U10].
224. Russell et al. [R17] studied the incidence of seven specific locus mutations in (101 × C3H) F1
wild-type male mice using about 40,000 offspring of males exposed to HTO with weighted mean dosesof 4.3 Gy to post-spermatogonial germ cells and 6.2 Gy to spermatogonia. The observed mutation
spectrum was similar to those following previous exposure to external X- or gamma radiation. The
radiosensitivity of post-meiotic cells was similar to that observed for acute exposure to X-rays. For
spermatogonia, comparison was made with earlier experiments with low-dose rate gamma radiation
showing that the mutation rate was twice as high in the case of tritium exposure.
225. Pomerantseva et al. [P10] studied reciprocal translocations (RTs) in mouse stem spermatogonia
induced by HTO and 137Cs gamma radiation. HTO was administered to males by a single
intraperitoneal injection and excreted with a half-time of about 2.5 days. In order to adjust exposure
conditions, the dose rate of gamma radiation was reduced exponentially with the same half-time. Mean
doses of testis cells were 0.5, 1.0, 1.9 and 3.4 Gy of tritium beta radiation and 1.0, 1.9 and 3.7 Gy of
gamma radiation [B9]. In the post-sterile period, three–five months after exposure commenced, ten–twelve males from each experimental group and four–five males from control groups were sacrificed.
From each male, 100–200 spermatocytes were analysed at the stage of diakinesis-metaphase of the firstmeiotic division, and the number of multivalents in the form of rings and chains recorded. The groupsof animals exposed to beta and gamma radiation at a dose of 2 Gy were also studied over two–eightmonths and reciprocal translocation frequency did not significantly change during this period. Thisobservation indicates that elimination of cells with translocation from populations of spermatogonia
exposed continuously at low dose rates is insignificant. The increase of reciprocal translocation
frequency with increasing dose was observed over the entire dose range up to 3–4 Gy. The RBE of
tritium in this study was estimated to be 1.8; dose dependence of RBE was not observed.
226. The ICRP Publication 103 [I23] has stated that “there continues to be no direct evidence that
exposure of parents to radiation leads to excess heritable disease in offspring”. Nevertheless, on thebasis of the results of animal experiments and, citing the Committee’s UNSCEAR 2001 Report [U10],ICRP estimated, for protection purposes, a nominal genetic risk of about 0.2% Gy−1 for up to thesecond generation (grandchildren). For low-LET radiation, the ICRP value for the probability of severe
heritable effects is 0.5% per Gy for the reproductive population, estimated on the basis of mouse data.
227. Straume and Carsten [S23] noted that the heritable effects observed for other low-LET radiation
were also present following exposure to HTO. By grouping the RBE studies with genetic end points(such as chromosome aberrations and mutations in mice), they determined that the RBE values ranged
from 1 to 3, with the higher values associated with low doses and low-dose rates, largely owing to thecurvilinear response for the reference radiation.

298 UNSCEAR 2016 REPORT

3.

Germ cell effects

228. This subsection presents effects of tritium exposure of mammalian germ cells in one or several
generations that result in progeny death. These effects are interpreted as stochastic, presumably causedby damage to genetic material of a single cell, but not transferred to the next generations because of thelethal nature of the radiation-induced mutation.
229. Carsten and Commerford studied dominant lethal mutations (DLMs) in Hale-Stoner-Brookhavenstrain mice resulting from chronic HTO ingestion [C10]. In a two-generation study, mice weremaintained on drinking water with 0.11 MBq/mL HTO. Radiation doses to germ cells of the second
generation animals tested for DLMs were 0.28 Gy in females and 0.38 Gy in males. Mice were matedat the age of eight weeks, pairing exposed males and females and exposed and non-exposed.
Statistically significant increases in DLM frequency were detected in all the exposed groups comparedwith the control one. There was no effect of tritium exposure on breeding efficiency.
230. Mewissen et al. [M12, M13] studied cumulative genetic effects following exposure of male
C57BL/6M mice to tritium for six and ten generations. At each generation, weaned male breeders aged
35 days either received a single injection of tritiated thymidine (0.037 MBq/g of body weight) or were
exposed for five weeks to HTO (0.37 MBq/mL). The dose to male sperm over a 35-day period of
exposure in each generation was estimated at 0.037 Gy from HTO. At 10 weeks of age, all breeders
were sibling-mated. At the fifth generation, the average litter size was 6.56 and 6.72 vs. 6.92,
respectively, in the sibling line receiving tritiated thymidine, exposed to HTO and in the control. The
observed variations are significant at the 0.01% level by chi square test. Also, the average weight ofmice at weaning consistently decreased through successive generations in the sibling line exposed totritiated thymidine and to a lesser extent in those exposed to HTO, whereas individual weight remained
fairly constant in control mice. F1 and F2 offspring from the ninth generation were studied for litter size
and infant mortality. The litter size had decreased and infant mortality increased in experimentalgroups. DLM frequency (pre-implantation death) had increased in both experimental groups.
231. Balonov and Kudritskaia studied the frequency of DLMs in germ cells of male randomly bredmice induced by HTO and 137Cs gamma radiation [B9]. HTO was administered to males by a single
intraperitoneal injection and excreted with a half-time of about 2.5 days. In order to adjust exposure
conditions, the dose rate of gamma radiation was reduced exponentially with the same half-time. DLM
frequency was estimated from the results of four weekly matings of each male with 2–4 intact females
beginning from the tenth day after the exposure launch. At 17–18 days after the beginning of mating,the females were dissected and the number of yellow bodies in the ovaries, places of implantation anddead embryos were counted. It was concluded that HTO beta radiation was more effective than 137Cs
gamma radiation (figure XII). RBE values tended to increase with dose and effect reduction: from 1.6
at DHTO = 2 Gy to 2.2 at DHTO = 0.5 Gy. Linear extrapolation of RBE dependence on dose to zero gives
RBEmax = 2.6 [B9, B12].
232. In summary, the studies presented here demonstrated that internal exposure of mammal germ cells
to HTO can induce DLMs of the progeny at a wide range of radiation doses. RBE estimation showed
tritium beta radiation to be more effective than gamma rays for this biological end point by factors of
1.6 to 2.6 [B9, B12].

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299

Figure XII. Frequency of induced post-implantation death of randomly bred mouse embryos depending
on dose of 137Cs gamma radiation (lower curve) and of HTO beta radiation (upper curve) [B9, B12]

D. Effects of tritiated biochemical substrates
233. The extent of cellular injury caused by a tritiated biochemical substrate depends largely on where
it is incorporated into a cell and the duration of exposure. Tritiated DNA precursors, such as tritiatedthymidine, are theoretically more efficient in causing cellular injury because they form part of the basic
building block of a DNA strand. This effect was already reported in 1958 by Painter et al. [P1]. On theother hand, compounds containing tritium that are not close to the DNA in the cell, such as fats or someamino acids included in a non-nuclear protein, should pose a lesser risk.
234. Whereas radiobiological studies with HTO aimed mostly at specification of RBE values for either
deterministic or stochastic effects in mammals, similar studies with tritiated biochemical substrates
aimed to reveal their effects compared with those from HTO. In practical terms, these studies are used
for specification of dosimetric models for OBT and tritiated biochemical substrates. It is understoodthat direct studies of OBT-related biological effects are hardly feasible either in the environment or inexperiments on mammals because of low tritium concentration in OBT. Therefore, some tritiatedbiochemical substrates are used as an experimental surrogate for OBT in food, which is potentially afactor of public exposure.
235. Studies on the effects of nucleotropic forms of tritium such as tritiated DNA-precursors are of
special interest because they promote better understanding of radiobiological mechanisms of internal

300 UNSCEAR 2016 REPORT

exposure. In practical terms, their results will be used as scientific bases for future models for theprotection of workers dealing with those forms of tritium. However, the number of workers potentiallythus exposed is rather limited. Thus, exposure to and incorporation of tritiated thymidine is essentiallylargely of scientific interest.

1.

Studies in vivo

236. In experimental studies on mammals, studies have focused on biological effects resulting fromtritiated thymidine, the first synthesized tritium-labelled nucleoside, because of its pronounced
intranuclear localization in contrast to uniformly distributed HTO. Effects of tritiated amino acids withvaried intracellular distribution and of uniformly distributed 3H-glucose have been studied to a lesser
extent. Deterministic radiation-induced effects were the focus of these studies, except for one series ofexperiments (table 16) studying stochastic effects in germ cells of male mice, i.e. induction of heritable
reciprocal translocations in spermatogonia and non-heritable DLMs.
Table 16. Time-integrated frequency of excess post-implantation embryo death (DLM, % weekdays),frequency of reciprocal translocation (RT) in spermatogonia and relative testis mass reduction (RTMR, %)in one month after injection of HTO and tritiated biochemical substrates in male mice [B9, B10, B12]
TritiumcompoundIntake (MBq/gof body weight)

Testis cell dose
(Gy)

DLM
(%·week)

RT frequency
(%)

RTMR
(%)

HTO

3.3

1.0

38

0.44±0.20

47±2Glucose

3.3

0.9

32

0.46±0.16

46±4

Glycine

3.1

1.3

15

0.36±0.11

44±3

D,L-Lysine

3.1

1.5

63

0.38±0.15

51±4

L-Lysine

1.9

2.2

130

0.87±0.22

64±3

0.04

—

10

0.21±0.07

—

0.4

—

35

0.56±0.20

29±4

1.1

—

a

35

0.44±0.17

—

0.4

—

20

0.12±0.06

14±4

1.1

—

56

0.26±0.11

35±3

—

—

0

0.025±0.014

0Thymidine

Deoxycytidine

Controla

Low fertility.

237. Lambert [L3] found that the number of resting primary spermatocytes per tubule in rat
spermatogonia was halved in 72 hours by exposure to HTO and tritiated thymidine concentrations of
2.2 and 0.55 MBq/g of body mass, respectively, given as a single injection. He pointed out that thesevalues should be viewed with caution owing to uncertainties in several factors, such as the time of deathof the spermatogonia and, therefore, the dose from tritium that induced it.
238. Baker and McLaren studied effects of tritiated thymidine on the developing oocytes of randomly
bred Q strain mice [B4]. Following seven intraperitoneal injections of pregnant mice with 3H-thymidine(0.15, 1.5 and 15 MBq per injection), tritium label was detected in the ovaries of their progeny. The
total number of oocytes in the ovaries was reduced in all the exposed groups proportionally to injected
tritium activity. The primordial oocytes were more affected than multilayered follicles. Other functions

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301(e.g. body weight, fertility, ovarian weight) were affected only in the group with the highest injected
activity. It was concluded that mouse oocytes are highly sensitive to beta radiation from incorporated
3
H-thymidine during embryonic life.
239. In the experiments of Carr and Nolan [C9] on reduction of testis mass of the mouse following
single injections of tritiated thymidine (0.037–0.74 MBq/g of body mass) or HTO (or exposed to 60Co
gamma rays), a significant effect on the testis mass was seen after the injection of tritiated thymidine at
0.037 MBq/g of body mass, which delivered an estimated average absorbed dose to the testis of about
0.035 Gy during 16 weeks. The authors assessed that tritium from tritiated thymidine “fixed” in thetestis was about twice as effective as the more labile and uniformly distributed tritium from HTO andthat, in terms of injected amount tritiated thymidine, it is unlikely to be more than five times as
effective as HTO even at very low injected amounts.
240. Balonov et al. [B12], in a series of experiments on randomly bred male mice, studied both
reduction of testis mass and frequency of DLMs in germ cells induced by 3H-glucose, two amino acids(3H-glycine and 3H-lysine) and two nucleosides (3H-thymidine and 3H-deoxycytidine) following
intraperitoneal injections. Tritiated glucose and amino acids were administered in single injections andnucleosides in six portions during three days. In the post-sterile period, three–five months after
exposure commenced, ten–twelve males from each experimental group and four–five males fromcontrol groups were sacrificed and reciprocal translocations in mouse stem spermatogonia counted. The
effects were assessed in comparison with HTO [B9, B10]. For injected tritiated nucleosides, theconcept of tissue or cell dose was not applied. The main results of experiments are presented intable 16. As the time dependence of DLMs was different following administration of HTO and various
tritiated substrates, especially the nucleotropic forms, the presented parameter is time-integrated excess
post-implantation embryo death.
241. The administration of all the tested tritiated substrates resulted in the production of three
radiobiological effects that are qualitatively similar to those observed for HTO. As far as the timeintegrated DLM frequency is concerned, the order of effectiveness of the various chemical forms, perunit of injected activity per gram of body weight, is as follows: 3H-glycine – HTO and 3H-glucose –D,L-lysine-3H – 3H-deoxycytidine – L-lysine-3H – 3H-thymidine. DLM frequencies induced by labelledlysine, thymidine, and deoxycytidine are five–eight times higher than that from an equal HTO activity.
Per unit of testis dose, 3H-glycine is three times less effective than HTO, 3H-glucose and 3H-D,L-lysine
do not differ from HTO, and 3H-L-lysine is more effective by a factor of 1.5. The difference in genetic
efficiency will not only reflect different dynamics of retention of bound tritium in the testes but also its
different location with regard to the cell nucleus. The nuclear location of 3H-nucleosides and 3H-lysine
is well established, and the suggestion of a predominantly extranuclear location of 3H-glycine in germ
cells is consistent with observations. In contrast to DLM, reciprocal translocation frequencies in stem
spermatogonia reflect late mutagenic effects of tritium. The deterministic effect of testis mass reduction
at one month after tritium injection generally agrees with assessed testis doses [B9, B10, B12].
242. In summary, experiments on rodents with administration of biochemical substrates labelled withtritium confirmed the theoretical consideration that some of them are more efficient with regard to
induction of both deterministic (cell death) and stochastic (mutation) effects compared withadministration of equal activities of HTO. Per unit of intake, 3H-thymidine is 5–10 times more efficientthan HTO. The experiments also demonstrated elevated efficiency of other nucleotropic forms oftritium, L-lysine-3H and 3H-desoxicytidine. For 3H-labelled nucleosides, the application of concepts oftissue and cell dose is hampered by lack of data on their distribution in mammalian cells.

302 UNSCEAR 2016 REPORT

2.

Studies in vitro

243. In 1973, Snow [S16] showed that tritiated thymidine at activity concentrations between 0.37 and3.7 kBq/mL (0.01 and 0.1 µCi/mL) significantly reduced the number of cells in mouse blastocysts
cultivated in vitro after exposure starting during the two-cell stage. Activity concentrations exceeding
3.7 kBq/mL were lethal to the two-cell embryos. Several other authors have published consistentfindings [H11, K10, M23, O5, S21].
244. By incubating embryos in solutions of tritiated thymidine, Streffer et al. [S24] found that an
activity concentration of 18.5 kBq/mL almost completely inhibited the development of mouse
blastocytes and was about 1,000 times more effective than similar activity concentrations of HTO.
Similarly, they found that activity concentrations of 1.85 MBq/mL of HTO and 1.85 kBq/mL oftritiated thymidine reduced the yield of fully developed blastocytes to 50–60%.
245. The extreme ratios of “toxicity” reported refer to concentrations of tritium in the culture media
rather than dose to cells. In fact, Streffer et al. [S24], when culturing blastocysts, estimated that tritiatedthymidine at a concentration of 1.85 kBq/mL resulted in a dose rate to DNA of 70 mGy/h and its effectwas similar to that of 1.85 MBq/mL of HTO, which gave a (uniform) dose rate of 60 mGy/h. The
tritiated thymidine in these in vitro cultures was available for the duration of the experiment, thus
resulting in the labelling of all DNA synthesis in the developing embryo. This situation is different
from the situation that might arise in vivo. This was referred to by Furuno-Fukushi [F14].
246. Particularly interesting are those papers that compared the effects of exposure to different tritiatedcompounds. Clerici et al. [C22] compared the toxicity of tritiated thymidine with that of the toxicity offour tritiated amino acids; arginine, lysine, histidine and aspartic acid with regard to growth anddevelopment of two-cell mouse embryos exposed in vitro. Surprisingly, arginine was the most lethal of
all the tritiated compounds, requiring 1.1 kBq/mL of medium to kill 50% of the embryos. In
comparison, tritiated thymidine had an LD50 of about 3.0 kBq/mL. The LD50 of the other amino acids
was 2.2 kBq/mL for lysine, 4.8 kBq/mL for histidine and 14.8 kBq/mL for aspartic acid. The LD50 for
tritiated tryptophan (1 kBq/mL) was almost the same as that for tritiated arginine [K12]. The authors
commented that this was surprising because tryptophan was not excessively incorporated into histones.
However, two non-histone chromosomal proteins with high amounts of tryptophan were identified.
247. Similarly, in vitro experiments performed by Müller et al. [M24] on preimplantation mouseembryos have shown that, given the heterogeneous distribution and specific incorporation into DNA,
tritiated thymidine is 1,000–5,000 times more effective than HTO in inducing harmful effects at thesame level of applied activity concentration. Müller et al. also found that tritiated arginine, a histone
precursor, was more damaging than tritiated thymidine for a number of in vitro end points in the mouseembryo, including blastocyst formation, hatching of blastocysts, trophoblast outgrowth, inner cell mass
formation, number of cells per embryo and micronucleus formation. The mice embryos were incubatedin vitro in solutions with activity concentrations of 0.37 kBq/mL and 0.93 kBq/mL of both tritiatedarginine and tritiated thymidine. The authors postulated that the greater radiotoxicity of the tritiatedarginine was due to faster uptake and possibly because histone synthesis is not restricted to the S phaseof the cell cycle while thymidine would be incorporated only during DNA synthesis. The latterassumption was confirmed some years later in cell-cycle specific experiments [M25]. These studies
emphasize the importance of understanding the intracellular distribution of tritium-labelledbiochemicals, with effects relating more directly to nuclear dose than to averaged cell or tissue dose or
to levels of administered activity.
248. Furuno-Fukushi et al. [F14] treated lymphocytic leukaemia cells of mice for 50 hours withvarious tritiated compounds: thymidine, lysine, arginine, leucine, and aspartic acid. Cell doses for HTO

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303

and tritiated amino acids or nucleus doses for 3H-thymidine ranged from approximately 0.1 to 8 Gy (asinterpreted from graphical data). Cell survival decreased exponentially with increased substrate activity
concentration in culture medium for all compounds with the effects being greatest for thymidine,
followed by arginine, lysine, leucine and aspartic acid whereas cell-mutation frequencies increased
linearly. The concentrations for detectable cell killing and mutagenesis were about 37 kBq/mL for
tritiated thymidine, 37–370 kBq/mL for tritiated amino acids and 18.5–185 MBq/mL for HTO. When
activities in cells were measured and converted to dose in cells for tritiated amino acids and dose innuclei for 3H-thymidine, the response both in terms of cell survival and mutation frequency per unit
dose was estimated equal for all amino acids; however, elevated by a factor of 2–3 for 3H-thymidine.
249. Wang et al. [W4] examined the effects of tritiated biochemical substrates on cultured embryonic
mid-brain cells of mice using the following tritiated compounds: thymidine, uridine, arginine andglutamic acid. The cells were exposed to different concentrations of these compounds over a 20-hour
period. Assays of cell proliferation and differentiation and DNA and protein content were conducted.
Contrary to the studies by Müller et al. [M24] and Clerici et al. [C22], Wang and Zhou [W3, W4] foundthat both tritiated thymidine and tritiated uridine were more radiotoxic than tritiated arginine andtritiated glutamic acid. This was probably due to the different biological end points studied. Table 17
provides a summary of the effects of exposure to the tritiated compounds measured as the tritium
activity concentration (ID50) in the culture medium and corresponding dose for cells or nuclei (the latterfor 3H-TdR) necessary to inhibit the cellular processes (proliferation and differentiation) by 50%.
Tritiated thymidine behaved very differently from the other three tritiated compounds, with a muchsteeper dose–response curve. This is evident in table 17 where the ID50 for tritiated thymidine is much
lower. However, corresponding absorbed doses for cells or nuclei do not vary much whether thetritium-labelled substance is distributed uniformly or concentrated in the nucleus as is 3H-thymidine.
Table 17. Inhibitory effect of beta radiation from tritiated biochemical substrates on cellular
proliferation and differentiation [W4]
Proliferation
OBT

Differentiation

ID50(kBq/mL)

Absorbed dose
(Gy)

ID50
(kBq/mL)

Absorbed dose(Gy)

3

H-Thymidine

29

0.58

21

0.42

3

H-Arginine

193

0.85

163

0.66

3

H-Uridine

193

0.60

141

0.43

3

H-Glutamic acid

525

0.95

438

0.77

250. In summary, incubating mammalian cells and embryos in media containing biochemical
substrates labelled with tritium may result in various biological effects, such as death of cells andembryos, cell mutations and inhibited proliferation and differentiation. The radiobiological efficiencyof these outcomes, assessed per unit labelled substrate concentration in culture media, varies by a factor
of up to 1,000 compared with HTO and by a factor of up to some tens between substrates. These
differences reflect the active involvement of labelled biochemical substrates in biochemical processes
that lead to incorporation of tritium into cell organelles and subsequent exposure of cells and nuclei(such as in the case of labelled DNA precursors) with tritium beta radiation. However, theradiobiological efficiency of the different substrates is comparable when assessed per unit of cell or
nuclei radiation dose.

304 UNSCEAR 2016 REPORT

3.

Biophysical models

251. In order to study the impact of tritiated biochemical substrates or OBT, Chen [C14] performed
microdosimetric simulations to compare differences in energy deposition between uniform distributionof tritium within a cell (as expected with HTO) and a non-uniform distribution based on the assumptionthat all OBT was bound uniformly within biologically critical sites of dimensions from 10 nm to 2 µm.
The dose mean lineal energies within these critical targets were calculated to be a factor of 1.7 higher
for OBT bound to the critical site compared to HTO over a wide range of target dimensions. This effectresults from a localized increase in dose to the critical target due to a non-uniform distribution ofenergy deposition within the cell. However, the extent of any increase would depend on the extent towhich OBT preferentially localizes within critical targets.
252. Alloni et al. [A1] simulated radiobiological effects of tritium concentration, depending on its
chemical form, either in the cytoplasm or in the nucleus of the target cell. The biophysical trackstructure code PARTRAC was used to calculate nuclear doses, DNA damage yields and fragmentation
patterns for different localization of tritium in human interphase fibroblasts. For tritium distributed
selectively in the cytoplasm but excluded from the cell nucleus, the dose in the nucleus is 15% of theaverage dose in the cell. In the low- and medium-dose regions investigated in the paper, numbers of
double-strand breaks (DSBs) are proportional to the nuclear dose, with about 50 DSB/Gy. These results
illustrate the potential for over- or underestimating the risk associated with tritium intake when its
distribution at subcellular levels is not appropriately considered.
253. In summary, while many studies have examined how tritiated biochemical substrates are
partitioned within the body and within the cell, studies specifically looking at health effects due toexposure to those substances are limited. Those that are available indicate that most organic compounds
have about the same effectiveness as HTO, since they are distributed throughout cells and do not lead topreferential irradiation of the nucleus. Incorporation of some tritiated amino acids and tritiatednucleosides (e.g. thymidine), however, can lead to the accumulation of tritium in the nucleus withlonger retention times and a proportionately larger dose per unit intake.

VII. RELATIVE BIOLOGICAL EFFECTIVENESS
254. The RBE is the ratio of the absorbed dose of a reference radiation needed to cause a specificbiological response divided by the absorbed dose of the radiation of interest that causes the same
response. RBE values are experimentally observed values and differ for particular radiation types
according to the biological system and end point under consideration, dose, dose rate, and the reference
radiation. RBE values are the basis for but have to be distinguished from the concept of radiation
weighting factors (wR) used by ICRP in the calculation of equivalent and effective doses, in whichsimplifications are made that are considered appropriate by ICRP for protection purposes. In
radiological protection, the RBE for stochastic effects at low doses (RBEM) is of particular interest.
255. RBE data and biophysical considerations (see below) indicate that lower energy electrons (such as
those released by tritium) or photons are biologically more effective than higher energy gamma rays fora range of deterministic and stochastic end points. Although ICRP [I13, I23] recognized that there wasevidence for a significant variation in RBE values for low-LET radiation (e.g. increasing RBE withdecreasing photon energy), it was argued that a more detailed distinction was not warranted for the
purposes of radiological protection. Thus, a value for wR of 1 was chosen for practical reasons to apply
to all electrons and photons, including beta particles from tritium [C30].

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305

A. Track structure considerations
256. Tritium decay results in the production of a very low-energy beta particle (average energy
5.7 keV) of short range (average track length in water 0.56 µm) and, as a result, the average ionization
density (and LET) produced by the emitted beta particle is significantly higher than that produced byhigher energy electrons or photons, such as 60Co gamma rays (see table 18). Lower energy photons orelectrons similar to those produced by tritium decay also show a significant shift in microdosimetricenergy deposition patterns towards higher lineal energy (y) compared to higher energy photon orelectron fields. The spectra of energy deposition in low-pressure proportional counters over a range ofsimulated tissue site sizes for tritium, 250 kVp X-rays and 60Co gamma rays were measured by Ellett
and Braby [E2]. The results were then interpreted using the earlier site model of the Kellerer-Rossi
theory of dual radiation action (DRA) (e.g. [K7]) to estimate the RBE value for limiting low doses. The
DRA model simply assumes that the biological effect is proportional to the square of the energy
deposited in some small volume, often taken to be about 1 µm in diameter. They reported theoretical
RBE values for tritium of 3.75 compared to 60Co gamma rays and 1.5 compared to 250 kVp X-rays(half-value layer 1.8 mm Cu), assuming a critical site size of 1 µm.
Table 18. Track average LET, L∆ , in water for various radiations based on a cut-off energy, ∆, of
100 eV [I24]
RadiationL∆ (keV/µm)

Co gamma rays

0.22

200 kV X-rays

1.7

H beta rays

4.7

50 kV X-rays

6.3

60

3

257. On the nanometre scale also, analysis of the energy deposition patterns of tritium beta particles
has shown tritium to be more effective in producing larger sized clusters of ionization which can be
enfolded within a 2.3 nm diameter sphere compared with photons with energies above 100 keV [M17].
This represents ionization events on the dimensional scale of DNA.
258. A joint task group of ICRP and the International Commission of Radiation Units andMeasurements suggested a relationship between what the group called a quality factor, Q(y), and lineal
energy, y, defined as the energy imparted in a 1 µm diameter spherical tissue volume divided by its
mean chord length [I25]. The relationship was based in part on general observations and theoreticalconsiderations, with special consideration given to the experimental data on chromosome aberrations inhuman lymphocytes. The value of Q(y) obtained for tritium beta particles was approximately 2compared to orthovoltage X-rays. This is supported by theoretical calculations performed by Bigildeev
et al. [B19] on the basis of similar microdosimetric quantities on the micrometre scale.
259. Morstin et al. [M20] calculated the lineal energy spectra of tritium beta particles for spherical sites
with diameters from 1 nm to 10 μm and showed that the mean values varied by more than an order ofmagnitude over this range. Then, when they applied the assumptions of the site model of DRA theory,
they found that the predicted RBE of low doses of tritium beta particles relative to 250 kVp X-rays rosefrom a value of <1.1, for assumed 10 nm sensitive sites, to a peak of ~1.5 for 1 μm and then decreased
to ~0.6 for 10 μm. The corresponding predicted RBE value of tritium beta radiation relative 60Co
gamma rays was ~1.5 for 10 nm sites, ~2.9 for 1 μm and ~1.6 for 10 μm. Morstin et al. [M20] pointed
out limitations of the DRA approach and they also considered the possibility of two different pathways

306 UNSCEAR 2016 REPORTof radiation damage related to two different target sizes. They produced bidimensional correlated
distributions of lineal energy for spherical sites of 10 nm and 20 nm diameter (to represent DNA
double-strand break formation) within a gross sensitive volume of 1 μm diameter. By then “assuming
arbitrarily (by somewhat questionable analogy to the DRA theory)” a squared dependence of RBE on
the product of the lineal energies for the large and small sites, they obtained estimated theoretical RBE
values for tritium compared to 250 kVp X-rays of 1.6 for the 10 nm sites, and 1.8 for the 20 nm sites,
within the 1 μm gross sensitive volumes.
260. In a recent theoretical study, Chen [C16] performed microdosimetric simulations to compare dosemean lineal energies for HTO and OBT with that for 60Co gamma rays in the same size range from10 nm to 2 µm of spherical radiosensitive sites. Compared with 60Co gamma rays, the estimated RBE
value varied from 1.3 to 3.5 for HTO and for 2.3 to 5.6 for OBT.
261. Most of the above theoretical calculations of RBE values are based on the assumptions of uniform
interaction between pairs of elementary biological “sub-lesions” within sensitive sites of approximately
1 µm. However, a number of experimental investigations have indicated that the biologicaleffectiveness of radiation at low doses is determined predominantly by patterns of energy deposition
over much smaller distances down to nanometre dimensions and, therefore, micrometre sized simulated
volumes will not typically provide an adequate description of these patterns [G3, G5, G12, K8].
262. The effectiveness of low-energy electrons, similar to those produced by tritium, can be studied
using ultrasoft X-rays (0.1–5 keV) that interact in the cell to produce low-energy electrons. Data from a
range of laboratories around the world, with few exceptions, show ultrasoft X-rays to have increased
effectiveness for a wide range of biological end points compared to equal doses of conventional X-rays
or gamma rays [G6, G7, H10], with RBE values typically increasing with decreasing ultrasoft X-ray
energy down to CK X-rays (0.28 keV; producing a single photoelectron with a range less than 7 nm).
RBEs greater than unity were also found for TiK and CuK X-rays with energies (4.5 keV and 8.0 keV,
respectively) similar to the average energy of the emitted beta particle from tritium.
263. Hill [H11] reviewed several studies that looked at in vitro end points, such as dicentric aberrations
in human chromosomes, micronuclei induction, and mutations over a range of photon energies fromultrasoft X-rays to 60Co gamma rays. He observed a pronounced trend of an increase in RBE values
with decreasing photon energy for several biological end points, particularly for the induction ofdicentrics in human lymphocytes. He noted that, because of differences in cell types and biological end
points, the extent to which lower photon energy caused RBE values to rise was still uncertain. Recent
data by Frankenberg et al. [F11] report an RBEM (maximum RBE value for very low doses) of about 4
for soft (mammography) X-rays compared to 200 kVp X-rays.
264. The percentage of absorbed dose deposited by low-energy electrons (0.1–5 keV) in tissue is ~33%
for 60Co gamma rays, ~49% for 220 kV X-rays and rising to ~78% for tritium beta particles [N5]. These
are similar to the low-energy electrons produced by ultrasoft X-rays. It has been inferred that these lowenergy secondary electron track ends produced by low-LET radiation are the predominant cause of
DSB induction, cell inactivation and other cellular effects, with isolated sparse ionizations andexcitations apparently having little biological effect [B24, G6]. The contribution to absorbed dose of
these low-energy electrons is large for tritium beta particle compared to orthovoltage X-rays or gamma
rays. Therefore, the ultrasoft X-ray data would predict an increase in biological effectiveness as a resultof increased clustering of ionization events on the nanometre (DNA) scale leading to an increase in thenumber of DSB per unit absorbed dose, along with a slight increase in complexity of the breaks due toadditional associated damage within a few base pairs.

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307

265. Using Monte Carlo simulation, Moiseenko et al. [M18] modelled DSBs and single-strand breaks(SSBs) in cells exposed to tritium beta particles and low-energy photons. They found that a directenergy deposition of 10 eV could result in an SSB. They further studied base damage associated with
DSB and were able to differentiate between simple DSBs and complex DSBs. They later developed aMonte Carlo model to calculate yields of DSBs in DNA after irradiation with 137Cs gamma radiation,orthovoltage X-rays (typically 150–300 kVp) and tritium beta particles [M19]. The RBE values for
DSB production for tritium beta radiation (with 137Cs gamma radiation as the reference) was 1.2 for the
total DSB yield and 1.3 for complex DSBs. They explained that low-energy X-rays and tritium beta
particles tended to deposit energy in a more clustered fashion than 137Cs gamma rays. They concluded
that tritium beta particles were more efficient in producing DSBs in DNA compared to 137Cs gamma
rays and that their relative effectiveness was even greater for the production of complex DSBs.
266. In summary, track structure considerations suggest that the low-energy beta particles produced bytritium decay are more biologically effective than hard X-rays and gamma rays per unit absorbed dose,at least in producing DSBs in DNA. This is a result of the average ionization density along the track of
the tritium beta particle being significantly higher than produced by much higher energy photons.
Theoretical calculations based on microdosimetric considerations suggest an RBE value ofapproximately 2, relative to 60Co gamma rays.

B. RBE literature reviews and experimental studies
267. In studies aimed at deriving values of RBE for tritium beta radiation, there is the potential for thevalues obtained to depend on the total dose and rate at which dose from both tritium and the reference
radiation are delivered. In the case of 3H, the dose is expected to be protracted in time since the dose
rate is dictated by the rate of tritium loss from the body and, to lesser extent, by its radioactive
disintegration. Experiments using X- or gamma rays, on the other hand, often deliver dose in a single
acute exposure because this is more convenient in practice. It is generally accepted that the same dosedelivered in a protracted manner can have a lower effect than an acute dose would, due to the greater
opportunity for DNA repair in the protracted case [I13, N3, N7, U7, U8].
268. Most of the lower values for the RBE of tritium reported in the literature are from studies thatused higher doses and dose rates of the reference radiation. This trend can be explained by the highdose rate of the reference radiation reducing the apparent relative effectiveness of the tritium doses. A
review of tritium RBE studies was carried out by Ujeno [U2] illustrated this phenomenon. Thosestudies, which included external reference radiation, showed a tendency for an inverse relationshipbetween dose rate and RBE value. The author concluded that use of a RBE value of 1 would bereasonable for assessing the dose from a very large intake of tritium but that a figure larger than 1
would be more appropriate for environmental and occupational exposure situations. Therefore, thedifferences in RBE values can be viewed assuming that the response at high acute doses is lessdependent upon or is independent of photon/electron energy.
269. In numerous studies of tritium radiobiological effects conducted since the early 1960s, the focus
has been on the RBE values for the tritium beta radiation for assessing stochastic health risks (primarily
cancer induction and heritable effects) of lower doses in mammals to be used in radiation protection ofhumans and environment. Those studies were carefully reviewed by several groups of authors over the
past two decades [H16, K20, L12, O1, P2, S23].

308 UNSCEAR 2016 REPORT270. Straume and Carsten [S23] provided a comprehensive review of the literature on the
carcinogenic, heritable, developmental and reproductive effects associated with tritium exposure.
They identified 33 published studies on the RBE of the tritium beta radiation using HTO: 12 studies
used X-rays (200–500 kVp) as the reference radiation; 21 studies used gamma rays from 137Cs or
60
Co as the reference radiation. Combining these studies, they calculated an arithmetic mean of 1.8
and 2.3 for the range of RBE values when X-rays and gamma rays, respectively, were used as thereference radiation.
271. The AGIR report [H16], discussed RBE values of tritium beta radiation and also the relationshipbetween RBE value and the dose and dose-rate effectiveness factor (DDREF), the use of different
reference radiation, and some experimental studies. They noted that the RBE values for tritium beta
radiation using HTO from a wide variety of cellular and genetic studies were generally found to be in
the range of 1–2 when X-rays were the reference, and 2–3 when gamma rays were the reference.
272. Little and Lambert [L12] subsequently conducted a comprehensive analysis of several peerreviewed studies with the intent of determining maximum low-dose RBE values, denoted RBEM. The
biological end points of these tritium studies included carcinogenesis, chromosomal aberration, and cell
death. The studies were divided into in vivo and in vitro experiments, and subdivided further according towhether the reference exposure was prolonged or acute. To be classified as a prolonged exposure, theexternal irradiation dose rate had to be comparable to that for exposure to tritium. The authors concluded
that the overall aggregated results implied RBE values with a central estimate in the range of 1.2–2.5 anda 97.5 upper percentile of no more than 3.0. The six studies with prolonged gamma radiation as reference
exposure and end points apart from cell survival yielded aggregate RBE values of 2.19 (95% CI: 2.04,2.33) versus 1.17 (95% CI: 0.96, 1.39) when chronic X-rays were used as reference radiation.
273. A review of RBE values by Kocher et al. [K20] was conducted to support assessments of cancer
risk from known exposure to ionizing radiation and estimation of the probability of causation. A
defining characteristic of analysis of data on RBEs by these authors was that estimates of radiation
effectiveness factors were expressed as subjective probability distributions to represent uncertainty
arising from uncertainties in the underlying estimates of RBE values.

C. Factors affecting RBE values
274. The reference radiation types used in RBE studies were 150–250 kVp X-rays, and gamma
radiation from 60Co (1,173 and 1,332 keV) or 137Cs (662 keV). However, studies such as those by Bondet al. [B23], Sasaki [S3] and Schmid [S6] reported in the ICRP Publication 92 [I20], demonstrated thatorthovoltage X-rays are typically about twice as effective at low doses as are high energy gamma rays
and that this difference is consistent with biophysical calculations [E2, K9].
275. Straume and Carsten [S23] discussed this issue, indicating that RBE can vary significantly withchanges in dose rate and radiation quality. Furthermore, as noted by AGIR [H16], the dose–response
curve for acute doses of low-LET reference radiation is often curvilinear in relation to the response toradiation with higher LET. This curvilinearity of response to reference radiation in the low-dose range
means that the RBE between the two types of radiation is maximum at lower doses. The maximal RBE
value is referred to as the RBEM.
276. Cell mutations and chromosome aberrations represent chromosome damage that could lead tocancer development. However, as Little and Lambert [L12] noted, chromosome damage may be only
one of many steps in carcinogenesis. Hill [H11] also pointed out that many studies used the induction of

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

309

dicentric aberrations as an end point because it was a reliable and repeatable method for comparing
biological response, recognizing that this was a short-term effect not directly relevant to quantitative
assessment of long-term stochastic effects.
277. The yield of DSBs is considered to be strongly dependent on biological systems and cellularenvironments and has been used for RBE determinations. After reviewing data on DSB yields for lowenergy electrons and high-energy protons of comparable microdosimetric characteristics to those oftritium in the dimensions relevant to DSBs, Chen [C15] estimated that average yields of 2.7, 0.93, 2.4
and 1.6 × 10−11 DSBs/Gy·Da were reasonable estimates of DSB yields for tritium in plasmid DNAs,
yeast cells, Chinese hamster V79 cells and human fibroblasts, respectively. If a biological system is not
specified, the DSB yield from tritium exposure can be estimated as a simple average overexperimentally determined yields as 2.3±0.7 × 10−11 DSBs/Gy·Da in various biological systems.

D. Summary of RBE value determinations
278. In summary, about 50 different experimental estimates of the RBE values for the tritium beta
radiation in animals or animal cells have been reported that ranged from 1.0 to 5.0 (centred around2-2.5) and 0.4-8.0 (centred around 1.5–2) with gamma rays and orthovoltage X-rays as reference
radiation, respectively. There is tendency for RBE values to increase with decreasing doses. RBE
values derived from stochastic effect studies are generally higher (centred around 2.5–3 compared withprolonged gamma radiation) than those obtained from studies of deterministic effects (cell killing invivo and in vitro). Considerable variation from one experiment to another exists. Only threeexperimental studies directly addressed carcinogenic effects in mammals. In addition to experimental
uncertainty and the choice of a reference radiation, a number of other factors contribute to thisvariability, including differences in radiosensitivity of tissues, organs and organisms; differences in the
biological end points; variation in dose and dose rate; and choice of in vitro or in vivo test systems.
279. Tables 19 to 22 provide summaries of the studies of the RBE of tritium beta radiation in
mammals, the studies being grouped by the experimental condition (in vivo studies in tables 19 and 20,
and in vitro studies, including murine and human cells, in tables 21 and 22) and according to the type ofradiation used as reference (X-rays in tables 19 and 21, and gamma rays in tables 20 and 22). In all thelisted studies, tritium was used in HTO form. Tables 19 and 22 present studies in chronological orderwith both stochastic and deterministic end points; these are separated in summary table 23. When two
or more end points or various exposure conditions were used in the same study, they are presented intables as separate studies.
280. Using the data from some of these studies, Little and Lambert [L12] recalculated RBE values forthe tritium beta radiation and their results are also provided in the tables 19 to 22 and used in summary
table 23.

Study reference

Exposure conditions

Reference
radiation

Dose range (Gy)RBE values for the tritium beta radiation(95% CI where indicated)

Lambert [L3]Spermatogonial survival in mice

Single injection of HTO andprolonged X-rays

200 kVp X-rays

0.05 tritium beta
0.11 X-rays

Gragtmans et al. [G9]Cumulative incidence of
mammary tumours in S-D rats

Five HTO injections every two days
and acute and prolonged X-rays

200 kVp X-rays

0.46–3.85 tritium beta
0.57–1.78 X-rays (acute)
0.29–2.00 X-rays (prolonged)

0.68
1.1–1.3

Cumulative percentage ofmammary tumours in S-D rats

Five HTO injections every two days
and acute and prolonged X-rays

200 kVp X-rays

0.46–3.85 tritium beta
0.57–1.78 X-rays (acute)
0.29–2.00 X-rays (prolonged)

0.83
n.a.

Chopra and Heddle [C21]

Chromosome aberrations inperipheral blood in female mice

Single injection of HTO andprolonged X-rays

250 kVp X-rays

2.0–6.0 tritium beta
1.5–6 X-rays

1.14 (0.8–1.5)

Chopra and Heddle [C21]

Chromosome aberrations inspermatogonia in mice

Single injection of HTO andprolonged X-rays

250 kVp X-rays

1.5–4.5 tritium beta
1.5–4.5 X-rays

1.21 (0.8–1.9)

Johnson et al. [J5]Myeloid leukaemia in mice

Single injection of HTO andprolonged X-rays

200/150 kVp
X-rays

0.85–3.04 tritium beta
1.06–2.64 X-rays

1.24 (0.63–1.85)

Kozlowski et al. [K24]

Chromosome aberrations inbone marrow in miceContinuous intake of HTO andacute X-rays

250 kVp X-rays

0.6 tritium beta
0.5 X-rays

1–2
0.43 (0.20–0.81)aGragtmans et al. [G9]aBiological end point

Recalculated value by Little and Lambert [L12].

2.3

310 UNSCEAR 2016 REPORT

Table 19. In vivo studies using X-rays as reference radiation

Table 20. In vivo studies using gamma rays (prolonged irradiation) as reference radiation
Study reference

Biological end point

Exposure conditions

Reference
radiation

Dose range(Gy)

RBE values for the tritium beta radiation(95% CI where indicated)

Furchner [F13]Mortality in mice

Single injection of HTO andprolonged gamma

Cobalt-60 gamma(tritium simulator)

5.3–16.5 tritium beta
12.3–16.5 gamma

1.7

Dobson and Kwan [D6, D7]

Oocyte survival in miceContinuous intake of HTO
and prolonged gamma

Cobalt-60 gamma(tritium simulator)

0.07–0.88 tritium beta
0.22–1.25 gamma

2.8

Carr and Nolan [C9]Testes weight loss in mice

Single injection of HTO andprolonged gamma

Cobalt-60 gamma(tritium simulator)

0.14–0.58 tritium beta
0.58 gamma

1.43 (1.06–1.80)

Russell et al. [R17]Seven specific locus mutations in F1
mice from spermatogonia exposure

Single injection of HTO andprolonged gamma

Caesium-137 gamma
0.4 Gy/h

6–9 tritium beta

Balonov et al. [B10]

Testes weight loss in mice

Single injection of HTO andprolonged gamma

Caesium-137 gamma(tritium simulator)

0.12–3.4
0.25–3.7

1.8–2.2

Balonov et al. [B9, B12]

Dominant lethal mutations in malemice

Single injection of HTO andprolonged gamma

Caesium-137 gamma(tritium simulator)

0.5–3.4 tritium beta
1.0–3.7 gamma

1.6–2.2
RBEM = 2.6

Pomerantseva et al. [P10]and Balonov et al. [B12]

Reciprocal translocations in mice
spermatogonia

Single injection of HTO andprolonged gamma

Caesium-137 gamma(tritium simulator)

0.5–3.4 tritium beta
1.0–3.7 gamma

1.8

Zhou et al. [Z5]Dominant lethal mutations in female
mice

Single injection of HTO andprolonged gamma

Cobalt-60 gamma(tritium simulator)

0.04–0.91 tritium beta
0.53–2.7 gamma

2.5
2.94 (2.00–4.28)aIjiri [I26]

Apoptosis of small intestinal cells in
mice

Single injection of HTO andprolonged gamma

Caesium-137 gamma(tritium simulator)

0.0–0.29 tritium beta
0.0–2.9 gamma

2.1 (1.7–2.5)
1.6 (1.2–2.0)aIjiri [I26]

Apoptosis of descending colon cells
in mice

Single injection of HTO andprolonged gamma

Caesium-137 gamma(tritium simulator)

0.0–0.2 tritium beta
0.0–0.4 gamma

1.8 (1.4–2.2)
1.4 (1.2–1.6)aSatow et al. [S5]Oocyte killing in mice

Single injection of HTO andprolonged gamma

Caesium-137 gamma,(tritium simulator)

0.04–0.25 tritium beta
0.04–0.25 gamma

1.1–3.5

Satow et al. [S5]Teratogenic effects on rat embryos

Single injection of HTO andprolonged gamma

Caesium-137 gamma(tritium simulator)

2.0–6.0 tritium beta
1.75–6.8 gamma

2.6
1.01 (0.57–1.78)aZhou et al. [Z6]Dominant lethal mutations-oocytes

Single injection of HTO andprolonged gamma

Cobalt-60 gamma(tritium simulator)

0.2–0.6 tritium beta
0.7–2.7 gamma

2.8–3.4

Zhou et al. [Z6]Dominant lethal mutationsspermatocytes

Single injection of HTO andprolonged gamma

Cobalt-60 gamma(tritium simulator)

0.2–0.6 tritium beta
0.7–2.1 gamma

1.6–3.9

Zhou et al. [Z6]Dominant skeletal mutationsspermatogonia

Single injection of HTO andprolonged gamma

Cobalt-60 gamma(tritium simulator)

0.2–0.6 tritium beta
0.7–2.9 gamma

3.5–3.9

2.2

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM
311

Biological end point

Exposure conditions

Reference
radiation

Dose range(Gy)

RBE values for the tritium beta radiation(95% CI where indicated)

Zhou et al. [Z6]Oocyte survival

Single injection of HTO andprolonged gamma

Cobalt-60 gamma(tritium simulator)

0.2–0.6 tritium beta
0.7–2.9 gamma

1.4–2.0

Zhou et al. [Z6]Spermatogonial survival

Single injection of HTO andprolonged gamma

Cobalt-60 gamma(tritium simulator)

0.2–0.6 tritium beta
0.7–2.9 gamma

2.1–2.8

Zhou et al. [Z6]Chromosome aberrationsspermatogonia

Continuous intake of HTO
and prolonged gamma

Cobalt-60 gamma(tritium simulator)

0.2–0.6 tritium beta
0.7–2.9 gamma

2.9–3.8

Zhou et al. [Z6]Primary oocyte survival

Continuous intake of HTO
and prolonged gamma

Cobalt-60 gamma(tritium simulator)

0.2–0.6 tritium beta
0.7–2.9 gamma

1.5

Zhou et al. [Z6]Spermatogonial survival

Continuous intake of HTO
and prolonged gamma

Cobalt-60 gamma(tritium simulator)

0.2–0.6 tritium beta
0.7–2.9 gamma

2.3−2.5

Seyama et al. [S13]Cancer in mice

Single injection of HTO andprolonged gamma

Caesium-137 gamma(tritium simulator)

2.0–10.5 tritium beta
and gamma

2.5baRecalculated value by Little and Lambert [L12].

b

Calculated at 500 days.

312 UNSCEAR 2016 REPORT

Study referenceTable 21. In vitro studies using X-rays as reference radiation
Study reference

Biological end point

Exposure conditions

Reference radiation

Dose range(Gy)

RBE values for the tritium beta radiation(95% CI where indicated)

Chromosome aberrations inhuman lymphocytes

HTO (2 h or 53 h)
and acute X-rays

180 kVp X-rays

0.28–2.45 tritium beta
0.5–3.0 X-rays

1.17 (1.13–1.21)
1.91 (0.64, 3.18)b

Prosser et al. [P14]

Chromosome aberrations inhuman lymphocytes

HTO (30 min or 24 h)
and acute X-ray

250 kVp X-rays

0.2–4.0 tritium beta
0.1–4.1 X-rays

RBEM=1.13 (0.95–1.31)

Vulpis [V3]

Chromosome aberrations inhuman lymphocytes

HTO (20 min to 2.5 h)and acute X-rays

250 kVp X-rays

0.25–7.0 tritium beta
0.05–9.0 X-rays

2.6 at 0.25 Gy
1.10 at 7 Gy
8.0 (0.2–15.8)aLittle [L11]Transformation in mouse cells

HTO (5–168 h)
and acute X-rays

220 kVp X-rays

0.25–5.0 tritium beta
0.5–4.0 X-rays<1–2aKamiguchi et al. [K3, K4]

Chromosome-type aberrations
in human sperm

HTO (~80’) and acute X-rays

220 kVp X-rays

0.14–2.06 tritium beta
0.25–3.74 tritium beta
0.23–1.82 X-rays

1.08 max dose
1.96 min dose
1.39 (1.26–1.54)aKamiguchi et al. [K4]Chromatid-type aberrations inhuman sperm

HTO (~80’) and acute X-rays

220 kVp X-rays

0.14–2.06 tritium beta
0.25–3.74 tritium beta
0.23–1.82 X-rays

1.65 max dose
3.0 min dose
2.17 (1.73–2.73)aKamiguchi et al. [K4]Chromosome breakage
aberrations in human sperm

HTO (~80’) and acute X-rays

220 kVp X-rays

0.14–2.06 tritium beta
0.25–3.74 tritium beta
0.23–1.82 X-ray

1.14 max dose
2.07 min dose
1.47 (1.33–1.62)aKamiguchi et al. [K4]Chromosome-exchangeaberrations in human sperm

HTO (~80’) and acute X-rays

220 kVp X-rays

0.14–2.06 tritium beta
0.25–3.74 tritium beta
0.23–1.82 X-rays

1.54 min dose
2.81 min dose
1.96 (1.49–2.62)aa

Recalculated by Little and Lambert [L12].

b

Recalculated by Prosser et al. [P14].

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

Bocian et al. [B20]313

Study referenceaBiological end point

Exposure conditions

Reference radiation

Dose range(Gy)

RBE values for the tritium beta radiation(95% CI where indicated)

Ueno et al. [U1]Cell survival in mouse cells

HTO (20 h) andprolonged gamma

Cobalt-60 gamma

1.3–8.0 tritium beta
2.0–9.0 gamma

1.5
1.3–1.6a

Ueno et al. [U1]Micronuclei in mouse cells

HTO (20 h) andprolonged gamma

Cobalt-60 gamma

1.3–8.0 tritium beta
2.0–9.0 gamma

2.0
1.8–2.3a

Ueno et al. [U1]

Mutation induction in mouse cells

HTO (20 h) andprolonged gamma

Cobalt-60 gamma

1.5–4.7 tritium beta
1.5–4.7 gamma

1.8

Yamada et al. [Y1]

Mouse pronuclear embryo cell survival

Prolonged HTO andprolonged gamma

Cobalt-60 gamma

0.009–0.07 Gy/h tritium beta
0.02–0.12 Gy/h gammaa

1.09(0.50–1.68)

Yamada et al. [Y1]

Mouse early 2-cell embryo survival

Prolonged HTO andprolonged gamma

Cobalt-60 gamma

0.009–0.10 Gy/h tritium beta
0.02–0.12 Gy/h gammaa

1.70(1.21–2.20)

Yamada et al. [Y1]

Mouse late 2-cell embryo survival

Prolonged HTO andprolonged gamma

Cobalt-60 gamma

0.009–0.19 Gy/h tritium beta
0.02–0.30 Gy/h gammaa

1.25(0.88–1.62)

Matsuda et al. [M3]

Chromosome aberrations in mousezygotes

HTO (2 h) andprolonged gamma

Cobalt-60 gamma

0.09–0.34 tritium beta
0.05–0.30 gamma

2.0
1.62 (1.30–2.07)aTanaka et al. [T7]

Dicentric chromosome aberrations inhuman lymphocytes

Prolonged HTO andprolonged gamma

Cobalt-60 andCaesium-137 gamma

0.14–2.10 tritium beta
0.05–4.0 gamma

2.1–2.3
2.39 (2.20–2.59)aTanaka et al. [T7]

Chromosome aberrations in human
lymphocytes: centric rings

Prolonged HTO andprolonged gamma

Cobalt-60 andCaesium-137 gamma

0.14–2.10 tritium beta
0.05–4.0 gamma

n.a.
3.14 (2.56–3.86)aTanaka et al. [T7]

Chromosome aberrations in human
lymphocytes: dicentrics and centric rings

Prolonged HTO andprolonged gamma

Cobalt-60 andCaesium-137 gamma

0.14–2.10 tritium beta
0.05–4.0 gamma

2.2–2.7
2.52 (2.33–2.72)aTanaka et al. [T7]

Chromosome aberrations total in humanbone marrow cells

Prolonged HTO andprolonged gamma

Cobalt-60 andCaesium-137 gamma

0.13–1.11 tritium beta
0.25–2.0 gamma

1.13
1.30 (0.96–1.76)aTanaka et al. [T7]

Chromosome aberrations in humanbone marrow cells (chromatids)

Prolonged HTO andprolonged gamma

Cobalt-60 andCaesium-137 gamma

0.13–1.11 tritium beta
0.25–2.0 gamma

3.1
4.96 (3.73–6.59)a

Recalculated by Little and Lambert [L12].

314 UNSCEAR 2016 REPORT

Table 22. In vitro studies using gamma rays (prolonged irradiation) as reference radiation

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

315

281. The RBE values derived from 48 experiments, on mammals in vivo (28) and mammalian cells in
vitro (20), is presented in table 23. The RBE values are presented in three study groups: for all studies,for studies of stochastic effects (carcinogenic, genetic, cytogenetic, cell transformations), and for
carcinogenic effect in mammals separately; the latter two compared with effects of prolonged exposure
to photon radiation (gamma and X-rays). The latter two RBE sets of values are more relevant to effectsof low radiation doses on humans. In line with the work of Little and Lambert [L12], the data presented
in table 23 are combined mostly from maximum low-dose RBE values (RBEM if available).
Table 23. Summary of tritium radiation RBE values from experimental studies using different end
points and different reference radiation

All studies
Studies

In vivo

In vitroaReference
radiation

Studies of stochastic effectsa
with prolonged reference
exposureStudies of carcinogenic effectin mammals with prolonged
reference exposure

Numberofstudies

RBE value(mean/medianand range)

Numberofstudies

RBE value(mean/medianand range)

Number ofstudies

RBE value(mean/medianand range)

Prolonged
gamma

21

2.5 / 2.5
(1.0−3.9)9

3.0 / 3.0
(1.8–3.9)

1

2.5

X-rays7

1.1 / 1.2
(0.4−2.3)

4

1.2 / 1.2
(1.1–1.3)

2

1.3 / 1.3
(1.2–1.3)

Prolonged
gamma

12

2.1 / 1.8
(1.1−5.0)

8

2.5 / 2.4
(1.3–5.0)

—

—

Acute
X-rays8

2.4 / 1.7
(1.1−8.0)

—

—

—

—

Including carcinogenic effects.

282. In broad terms, the RBE values from all studies ranged from 1.0 to 5.0 (centred around 2–2.5) and
0.4–8.0 (centred around 1.5–2) with gamma rays and orthovoltage X-rays as reference radiation,
respectively. Studies show a general tendency of RBE values to increase with lower doses. There issome tendency for RBE values derived from studies of stochastic effects (centred around 2.5–3 whencompared with prolonged gamma radiation) to be generally higher than those obtained from studies ofdeterministic effects such as cell killing in vivo and in vitro (not presented separately in table 23).
283. Only three of the studies identified have used cancer incidence as an end point [G9, J5, S13].
These studies include accelerated cancer incidence, i.e. cancer occurring at an early age, rather than anincrease in the overall incidence of disease, as the end point. The small number of studies andambiguous end points limit the opportunity to come to a clear conclusion regarding RBE values or theirrange for the carcinogenic effect of tritium in mammals.

316 UNSCEAR 2016 REPORT

VIII. EPIDEMIOLOGICAL STUDIES
284. Epidemiological studies of groups of people exposed (or potentially exposed) to tritium fall into
two broad categories: those exposed at work and those exposed in the environment. Workplace
exposure generally provides a better opportunity for assessing the tritium-specific risks to healthfollowing doses received by particular tissues or organs from internally deposited tritium becausemonitoring for occupational exposure to tritium will usually have been conducted at the facility where
exposure occurred (or potentially occurred) and doses may be estimated from these monitoring results.
Further, data from monitoring for other sources of occupational exposure to ionizing radiation are likely
to be available if assessments for exposure to tritium have been performed at a facility. Such dosemonitoring data are required when tritium-specific risk is estimated to distinguish it from others
radiation risk sources. Epidemiological studies of tritium workers have generally been unsatisfactory
regarding the use of tritium-specific doses monitoring data. Moreover, studies of occupational exposure
involve mainly adult men and do not include children who may be more sensitive to tritium-induced
adverse health effects.
285. Members of the public are exposed not only to natural sources of tritium, but also toanthropogenic sources such as tritium produced in nuclear weapon explosions, particularly the fallout
from the atmospheric thermonuclear weapon tests of the early 1960s. Public exposure to tritium also
occurs as a result of releases from nuclear power and nuclear weapon facilities, or from luminizing,radiochemical and other plants, and from devices containing tritium, such as wristwatches with tritiumbased luminous paint or emergency exit signs. Studies of environmental exposure to tritium have theadvantage that they usually involve exposed (or potentially exposed) individuals other than just adults
who are fit enough to be at work. However, a substantial drawback of such environmental studies isthat bioassay monitoring for exposure to tritium is unlikely to have been conducted, which greatly
reduces the reliability of environmental studies to assess risks specific to tritium exposure. However,there are some instances where such monitoring has been reported for environmental studies.
286. Any analysis of risk in terms of tritium exposure will need to take account of other sources ofradiation exposure, such as penetrating radiation from external sources and from intake of otherradionuclides, to appropriately distinguish tritium-specific risks from those arising from other sourcesof radiation exposure. It is important to ensure, if there is a positive correlation between tritium doseand other radiation doses, that any risk from these other radiation doses are taken into consideration.

A. Studies of occupational exposure
287. Given the variety of sources of exposure to tritium, epidemiological studies of the risks to healthfrom tritium exposure would seem to be attractive, especially in an occupational setting since thoseworkers potentially exposed to tritium are likely to have been monitored for such potential exposure
through the analysis of urine samples. The results of this monitoring should have been recorded and, if
these records still exist, the data could be made available for scientific use through the production oftritium-specific doses to organs/tissues. Unfortunately, although a number of epidemiological studies oftritium workers have been conducted in various countries, few of these studies have made direct use oftritium monitoring data or have used tritium-specific doses derived from urinalysis data. Studies of
exposure to tritium in the workplace are considered below in four broad groupings of studies.

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288. All workers at installations with nuclear reactors or reprocessing plants will have been exposed totritium to some extent because tritium is produced (at a low frequency) in ternary nuclear fission.
However, studies of workers exposed to tritium have concentrated on those workers who are likely tohave received non-trivial doses from tritium because of certain features of operations at the sites, suchas the presence of heavy-water-moderated reactors or tritium production or processing facilities. The
epidemiological studies considered in this section focus on studies of workers at such nuclear sites
rather than workers exposed to very low levels of tritium because of work at other sites, including lightwater-moderated or gas cooled, graphite moderated reactors.

1.

Studies of workers at installations where tritium is present

289. The weakest of the epidemiological studies of occupational exposure to tritium are those studies
that consider workers at sites where exposure to tritium occurs, but make no distinction between
workers exposed (or potentially exposed) to tritium and other workers at the site not so exposed. Boththe workers monitored for potential exposure to tritium and at least some of the other workers at a site
are likely to have been exposed to other sources of radiation.
290. For example, Cragle et al. [C31] studied mortality among almost 10,000 white male workers
employed at the Savannah River Site, United States during 1952–1980, and they noted that around
5,000 workers would have been exposed to tritium with 800 of these having received a dose of at least
0.5 mSv from tritium. The workers were found to have standardized mortality ratios (SMRs) that weregenerally less than 1.0 when compared to the population of the United States (which is probably areflection of the “healthy worker effect”), but no distinction was made in the analysis between workers
monitored for exposure to tritium and other workers. So, although the results of this study are broadly
reassuring as far as workers who have been employed at the Savannah River Site, United States areconcerned, they are of limited informative value when assessing the risk arising from exposure totritium, except that the risk of tritium exposure cannot have been grossly underestimated or this would
be apparent in the overall results of the study.
291. In another study, McGeoghegan and Binks [M5] examined mortality and cancer incidence among
workers at the Capenhurst site in the United Kingdom, a nuclear installation that has handled tritium inrelation to nuclear weapon production. This study did not specifically identify those workers monitored
for potential exposure to tritium. McGeoghegan and Binks [M5] reported that radiation workers at
Capenhurst had a significantly low SMR for all causes of death and a significantly low standardized
registration ratio (SRR) for all incident cancer, but workers at this site were exposed to externalradiation and to radionuclides other than tritium and it is not possible in this study to disentangle risk
posed by tritium from that posed by other sources of radiation. The findings of studies of workforces atestablishments where tritium is present to some extent in non-trivial quantities, but which do not
distinguish tritium workers from other radiation workers, are summarized in appendix A, table A1.

2.

Studies of workers monitored for potential exposure to tritium

292. The next set of studies embraces those workers who have been monitored for potential exposure totritium, but for whom tritium-specific doses are not available or, if available, have not been used. Thesestudies typically identify those workers at an installation who have been monitored for potential exposure
to tritium and then calculate SMRs for that particular group. For example, in the study of workers of theUnited Kingdom Atomic Energy Authority (UKAEA), Beral et al. [B17] identified the subset of workers
who had been monitored for potential exposure to tritium and calculated separate SMRs.

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293. In some studies, the SMRs for the tritium workers are compared with the SMRs for other workers
at the installation to generate rate ratios (RRs), which have the benefit of addressing (at least, to someextent) the healthy worker effect that may be present if the analysis is limited to the calculation of
SMRs alone and the reference population is the general population of a country or region. For example,
in a follow-up study of UKAEA workers, Fraser et al. [F12] not only calculated SMRs for tritium
workers but also compared these SMRs with equivalent SMRs for radiation workers at the UKAEA
who had not been monitored for potential exposure to tritium, to generate RRs. However, the absenceof tritium-specific doses in these studies means that quantitative tritium-specific risk estimates cannot
be generated, although the calculation of SMRs and RRs for tritium workers does permit theidentification of possible large effects arising from exposure to tritium, as happened when Beral et al.
and Fraser et al. [B17, F12] found significantly raised prostate cancer SMRs and RRs for UKAEA
workers who had been monitored for potential exposure to tritium. The findings of this group of studies
of workers monitored for potential exposure to tritium are presented in appendix A, table A2.

3.

Studies of workers using occupational dose estimates

294. Some studies of workplace exposure to tritium have used occupational dose estimates, but havenot directly used estimates of tritium-specific doses, if available. Typically, these studies identify
workers at an installation who have been monitored for potential exposure to tritium, calculate SMRs
and possibly also RRs, and then conduct a dose–response analysis in terms of recorded doses ofpenetrating radiation from external sources rather than tritium-specific doses. For example, in a nested
case-control study of prostate cancer risk among UKAEA workers, Rooney et al. [R14] found that forworkers who had been monitored for potential exposure to tritium (or to one of four other radionuclides
frequently found in the same workplace environment as tritium), the relative risk of prostate cancer
significantly increased with the recorded dose of external radiation, whereas it did not for otherworkers. For those workers either monitored for potential exposure to tritium or not so monitored but
assessed to have the potential for exposure to tritium, the relative risk significantly increased with theassessed level of potential exposure. However, although the significantly increased relative risk was
confined to those workers monitored for potential exposure to tritium (rather than those assessed to bepotentially exposed, but not monitored) no use was made of the tritium monitoring data to derive
tritium-specific doses for analysis [R14]. The absence of tritium-specific doses in this study
substantially limits the interpretation of the associations found in terms of tritium-specific risk.
295. Sometimes tritium doses have been derived from monitoring data but are then included withexternal doses since it is usually argued that tritium produces whole-body doses that are essentiallyequivalent to (and generally smaller than) doses received from external sources of penetrating gamma
rays. For example, in the study of workers of the United Kingdom Atomic Weapons Establishment(AWE), Beral et al. [B18] added the recorded whole-body dose from tritium based upon monitoringdata to the recorded whole-body dose from external sources. They found a significantly increasing
trend of RR for prostate cancer mortality with increasing whole-body dose (driven by one death with acumulative external dose >100 mSv), but did not conduct an analysis in which the tritium-specific dose
was separated out from the external dose. Any inference concerning a tritium-specific risk obtained
from such studies of a positive dose–response for tritium should rely on an assumption of a positive
correlation between external doses and tritium doses.
296. Zablotska et al. [Z1] studied mortality in Canadian nuclear industry workers, and added recorded
tritium doses derived from urinalysis results to recorded doses from external sources. Again, this did
not permit any tritium-specific risk to be identified from the published results because any findings ofanalyses based on tritium-specific doses were not presented. However, of interest are the results for theANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

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ERR/Sv for leukaemia (excluding CLL) and for all solid cancer, with external doses combined withtritium doses, in comparison with the ERR/Sv estimates using external doses not combined with tritium
doses (i.e. for external doses alone): for the combined doses, the ERR/Sv estimates are 18.9 (95% CI:<−2.08, 138) and 2.80 (95% CI: −0.038, 7.13), respectively, and for external dose alone, the ERR/Sv
estimates are 16.3 (CI not given) and 2.67 (CI not given), respectively. Therefore, the ERR/Sv
estimates show a small increase when the tritium doses are included with external doses but,unfortunately, it is not possible from the published results to derive ERR/Sv estimates for tritium doses
alone. The findings of studies that have used records of occupational doses of radiation, but have notused tritium-specific doses for the analysis, are presented in appendix A, table A3.

4.

Studies of workers using tritium-specific dose estimates

297. Few studies of workers have tritium-specific dose estimates derived from occupational exposure
records and use these doses in a tritium-specific risk analysis. If tritium-specific doses are used in ananalysis that appropriately adjusts for any effect of doses received from external sources of radiation(and for any doses received from any other internally deposited radionuclides), these tritium-specificdoses should enable estimates to be made of tritium-specific risks, although due account must be takenof the precision of these estimates since the number of workers included in such an analysis may besmall, leading to limited statistical power.
298. Zablotska et al. [Z2] conducted a study of Canadian nuclear industry workers and used tritiumspecific doses in addition to external doses. However, in most analyses, tritium doses were combinedwith external doses, so the tritium-specific risk was assessed for only one analysis: the ERR/Sv for all
solid cancer was reported as −4.71 (95% CI: <−5.92, 8.58). The wide confidence interval for thisestimate is indicative of the limited power of studies of tritium exposure in just one country.
299. Hamra et al. [H4] studied the tritium-specific risk of leukaemia among workers at the SRS andsupplemented tritium-specific doses derived directly from occupational monitoring records with annual
tritium doses reconstructed from external dose records using tritium monitoring results combined with a
job exposure matrix. The authors, in a Bayesian analysis, reported that for leukaemia excluding CLL,
the ERR/10 mGy estimate was −0.281 (90% credibility interval: −1.136, 0.548) while if the constraint
was imposed that the ERR/10 mGy for the dose from tritium beta particles was greater than that for thedose from penetrating gamma rays, then the ERR/10 mGy estimate became 0.334 (90% credibility
interval: 0.049, 0.817). The sensitivity of the results to this constraint and the width of the credibility
intervals are notable.
300. Studies of adverse health effects in the offspring of workers exposed to tritium in thepreconceptional period, which have made use of tritium monitoring results to calculate tritium-specificdoses, have been conducted in Canada [G10, M9] and the United Kingdom [H17, H18]. No association
between adverse health effects in offspring and the preconceptional dose derived from tritium
monitoring data were found. However, the UK-study illustrated a difficulty in the interpretation offindings based upon the assessed potential for historical exposure to tritium rather than the use ofmonitoring of dose records. In this regard, a highly significant association between the risk ofleukaemia and non-Hodgkin’s lymphoma in the offspring of Sellafield workers was assessed topotential for exposure to tritium of fathers in the preconceptional period and was not confirmed bytritium dose monitoring data. This calls into question the reliability of assessed potential for exposure totritium during historical operations when monitoring of those workers most likely to have received
doses resulted in different conclusions. The findings of these studies that have explicitly used tritiumspecific doses are summarized in appendix A, table A4.

320 UNSCEAR 2016 REPORT

B. Studies of environmental exposure301. The situation is much more uncertain in studies of members of the public potentially exposed totritium because direct measurements of tritium body burdens in non-occupationally exposed people arerare and not undertaken as part of an epidemiological study. Therefore, any individual assessments oflikely environmental exposure to tritium for use in epidemiological studies have to rely upon modelling
results, such as estimates of intake and consequent doses to people based upon measurements of tritiumin environmental media; but the derived tritium-specific doses are generally very small, leading to
extremely low statistical power to detect any tritium-specific effects. Some studies rely only on
measures of proximity to a source of tritium, such as a nuclear facility, and this inevitably leads toresults that have an uncertain interpretation because the relationship between linear distance from thefacility and tritium dose has not been established, and the relationship between distance and level of
exposure could be very complex and far from being directly proportional if, say, the wind rose issignificantly anisotropic. Further, such studies must take other relevant exposure into account,
including other sources of radiation. As a consequence, studies of public exposure to tritium have to be
examined carefully and their findings need to be interpreted accordingly.
302. In only one epidemiological study of environmental exposure to tritium has individually assessed
exposure to tritium been used, and that is in a historical cohort study of cancer incidence during
1986-2005 among residents of two areas near the Pickering heavy-water-moderated CANDU reactorsite in Ontario, Canada [W5]. In this study, exposure to tritium at residences occupied in 1985 wasestimated from an atmospheric dispersion model that used discharge and meteorological data, and theexposure data were used in an analysis of cancer risk in relation to tritium exposure. Assessed annual
individual effective doses from tritium were very low (maximum for an adult, 2.36 µSv), and thelimited number of cases of cancer available for study presented problems for some of the analyses, so
the lack of detection of an effect of tritium exposure upon cancer risk must be viewed in this light.
Nonetheless, this study does illustrate what may be done to address tritium-specific risk in a study ofenvironmental exposure, but it also demonstrates the problems of achieving reasonable statistical power
in a practicably sized study of such exposure.
303. With only one epidemiological study of environmental exposure to tritium that takes account ofassessed tritium-specific doses to individuals, very little can be reliably inferred from studies of healtheffects in the vicinities of installations producing, processing or storing tritium. Nuclear installations
that include nuclear reactors or reprocessing plants inevitably discharge tritium to some limited extent
because, for example, ternary fission product tritium will be produced/processed during the operation atsuch installations, although such discharges are likely to lead to very small doses. Studies aroundinstallations that include heavy-water-moderated reactors or tritium production or processing plants arelikely to address higher doses from tritium discharges, although even then tritium-specific doses will, in
general, be low; few of these studies have been conducted, and they will be briefly considered inappendix A, table A5.
304. One possibility that can be eliminated by environmental studies is that the risk of childhood
leukaemia from exposure to tritium has been grossly underestimated, and that this is the cause ofexcesses of childhood leukaemia incidence that have been reported around certain nuclear installations,as has been suggested by Fairlie [F1]. Substantial quantities of tritium were released into theenvironment by atmospheric nuclear weapon testing during the late 1950s and early 1960s, particularly
by thermonuclear weapon testing in the early 1960s [U8]. If there has been a serious underestimation of
the tritium-specific risk of childhood leukaemia then it would be apparent in the rates of childhood
leukaemia incidence following this period of intense nuclear weapon testing, particularly in the
Northern Hemisphere where most of the testing took place and tritium-specific doses were highest.

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Examination of childhood leukaemia incidence rates from around the world has not revealed any
evidence for an increase of childhood leukaemia risk that might be attributed to tritium fallout, or thatthe risk of childhood leukaemia has been greater in the Northern Hemisphere than in the Southern
Hemisphere after this period of intense nuclear weapon testing [W1, W2]. It cannot be claimed that theresults of this study show that the risk of childhood leukaemia arising from exposure to fallout from
nuclear weapon testing is less than predicted by standard radiation-induced leukaemia risk models (orthat there is no risk) because the statistical power is insufficient for this purpose, but the study can
exclude a risk that is much greater than predicted, as has been claimed.

C. Summary of epidemiological studies
305. The great majority of epidemiological studies of tritium workers have not used estimates oftritium-specific doses in their analyses, which limits the inferences that may be made about tritiumspecific risk using the findings of these studies. There is no indication, however, from studies of tritium
workers that tritium-specific risks have been seriously underestimated. Some results from a few studies
that have used tritium-specific doses are available, which represents progress in the epidemiologicalapproach to tritium-specific risks, but the conclusions that may be drawn about tritium-specific risks tohealth from these few studies are limited.
306. For scientific purposes, it should not be assumed that the effect per unit absorbed dose for tritiumis the same as that for an external source of penetrating gamma rays (i.e. the assumption of an RBE
value for tritium beta particles of unity), since this question is a component of the epidemiologicalstudy. A separate analysis of risk is required using tissue-specific absorbed doses received from tritium
rather than equivalent or effective doses. However, tritium-specific absorbed doses to tissues have only
rarely been used.
307. Owing to the limited numbers of tritium workers in particular countries, and limited exposure ofmost of these workers, it is unlikely that epidemiological studies of individual nuclear facilities, orindeed individual countries, will have sufficient statistical power to have a reasonable prospect ofdetecting the risks predicted by standard models, or of risks not far removed from those predicted.
Consequently, international collaboration is required to provide a study large enough to properly
investigate tritium risk. Some studies of tritium workers have already been conducted in Canada, theUnited Kingdom and the United States. Other candidate countries for studies of tritium workers include
France and the Russian Federation, and it is possible that other countries, such as China and India,
might be able to contribute worker data. It is clear that a coordinated effort is needed if a serious
epidemiological evaluation of tritium risk is to be made.
308. It is unlikely that epidemiological studies of environmental exposure to tritium will produce
meaningful tritium-specific risk estimates because such exposure is, in general, unlikely to produce
tritium-specific tissue/organ doses that are not low or very low. Further, measures of tritium exposureof members of the public based upon monitoring data are rare, and indirect assessments of tritiumspecific doses, particularly those based upon linear distance from a point of discharge, are likely to be
associated with considerable uncertainty. Those measurements of the presence of tritium in humantissue that have been made do not indicate that assessments of the amounts of tritium entering membersof the general public and being retained have been seriously underestimated.
309. The absence of a discernible impact upon global childhood leukaemia incidence rates of tritium
released into the environment by atmospheric nuclear weapon testing demonstrates that the risk ofchildhood leukaemia posed by tritium has not been grossly underestimated, as has been proposed bysome commentators.

322 UNSCEAR 2016 REPORTIX. RESEARCH NEEDS(a) Heterogeneity, concept of mean dose
310. The biodistribution of organic forms of tritium is heterogeneous within tissues and cells. The
issue of the relevance of the mean organ dose concept as a risk indicator therefore arises. There is no
appropriate dosimetric model for use in human risk assessment and radiation protection for tritiatednuclear acid precursors. The development of an appropriate microdosimetric approach to better
understand the distribution of dose from various organic forms of tritium within a cell and withintissues and organs is of importance. The uptake and long-term retention of heterogeneously distributedorganic forms of tritium in tissue and cells is particularly of concern in assessing doses to germ cells,
the embryo, the fetus and the infant. Another physico-chemical form of tritium that is a factor ofhuman, mostly occupational, exposure and deserves further study of biodistribution in body organs andtissues and biokinetics is tritium dust and flakes formed in nuclear fusion reactors on carbon, beryllium
or tungsten basis.

(b) RBE studies
311. Up-to-date methods should be used to gather further knowledge on the RBE of tritium beta
particles, especially as OBT, focusing not only on aspects relating to carcinogenesis but also on noncancer effects. In particular, research is required for the various stages of in utero and early childhood
exposure. Research aimed at addressing real or practical situations should be prioritized—for instancelong-term intake of HTO/OBT as food. Data regarding the potential induction of heritable/transgenerational effects should be critically assessed. New approaches should be investigated, in thelight of the latest advances in biology.

(c) Mechanistic studies
312. Mechanistic studies should emphasize cellular damage and notably the types and frequencies of
DNA damage caused by tritium beta radiation. Of particular interest is the complexity of DNA damage(that might affect efficiency and fidelity of repair) induced by either DNA synthesis precursors (tritiatedthymidine) or tritium-labelled amino acids in chromatin binding proteins, the triggering of (DNA)damage signalling pathways and activation of protective processes (e.g. repair, cell cycle arrest,
apoptosis, differentiation) in the context of toxicity, and genome instability. Data are also lacking on themetabolism and biological/mechanistic effects associated with organic tritium (tritiated biochemical
substances) in situations of chronic environmental exposure of the public.

(d) Environmental considerations
313. Given the still incomplete knowledge on tritium accumulation and behaviour in sediment,
targeted multidisciplinary studies with rigorous protocols need to be used to provide experimental
verification of the hypotheses regarding the possible influence of the activity of microorganisms inaquatic sediments when organic tritium is remobilized in aquatic animals. In general, the scientific data
regarding the conversion of HTO into organic tritium along the food chain should be enhanced.
Reliable quantitative estimates are required.

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314. Some information on the levels of tritium in the environment is available, but it would be of value
to enhance this database. Measurements in targeted environmental media, including those giving ahistorical record of exposure, such as tree-rings, particularly in the vicinity of tritium-handlingfacilities, would be desirable to provide additional confidence in the current understanding of tritium
behaviour in the environment and its transfer to food.
315. Some measurements of tritium in humans exposed in the environment are available, fromurinalysis and measurements in tissue, for example, at autopsy. It would be desirable to expand thisrather sparse database by gathering data from urine sampling programmes, particularly in residents neartritium-handling facilities, and from measurements made on various tissues as the opportunities arise.

(e) Epidemiological issues
316. Epidemiological studies are currently very limited in the robustness of their conclusions due toinsufficient statistical power and lack of information on tritium-specific doses. It should be noted thatrecorded tritium doses for workers in presently available reports are of the order of ten milligray, which
implies that very large cohorts would be needed to demonstrate a statistically significant increase incancer risk at conventionally predicted increases in risk. However, there may be facilities not currently
included in epidemiological studies at which workers were exposed to relatively high levels of tritiumin the early years of tritium production or processing, and it will be important to include such facilities
in any international collaboration. Such studies should use available tritium monitoring data, andexposure information may be improved by dose reconstruction methods, such as the job exposure
matrix methodology, which has already been used at one tritium production/processing facility. A
coordinated international approach based on standardized dosimetric assessments would be required inorder to make progress in this field. Such international collaboration is the only realistic prospect ofobtaining tritium-specific risk estimates from occupational epidemiological studies.
317. Since tritium will continue to find large-scale use, especially if commercial fusion reactors comeinto operation, it would be advisable to seriously consider international collaborative projects toinvestigate risks posed by tritium to workers that make full use of relevant occupational data, and thatthese studies should be based on a common protocol for the determination of tritium-specific doses
from occupational monitoring data.
318. Unless circumstances can be identified in which relatively large numbers of members of the public
have been exposed to relatively high levels of tritium, it is difficult to see how tritium-specific risk
estimates can be obtained from epidemiological studies. Should a biomarker of tritium-specificexposure be identified and developed, this could provide opportunities for studies but, even then, thevery low exposure from industrial tritium releases would be difficult to distinguish from thebackground presence of tritium.
319. A more realistic approach to assessing exposure and inferring risks to members of the public is
careful monitoring for the presence of tritium in the environment in the vicinity of facilities thatdischarge tritium, and the possible monitoring of tritium in selected members of the public in theseareas through, for example, urinalysis. However, monitoring programmes conducted in areas aroundfacilities handling tritium must also record data on the general levels and variability of tritium awayfrom such facilities in order that results can be appropriately interpreted.

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X. GENERAL CONCLUSIONS
320. Tritium is a radioactive isotope of hydrogen and thus behaves chemically like hydrogen. Humans
are exposed internally to beta radiation emitted by tritium either in occupational settings or as membersof the public. Workers are generally exposed to higher levels of tritium, as HTO, HT, metal tritides anddust, luminous compounds, tritiated biochemical substrates and some other anthropogenic chemical
forms. The general public is exposed to environmental HTO and OBT in food.
321. Absorbed doses arising from the intake of tritium cannot be measured directly and recourse has tobe made to the use of biokinetic and dosimetric models such as those of ICRP for dose assessment
based on environmental measurements and of bioassay methods (such as the tritium measurement inurine) combined with models for retrospective determination of doses from individual measurements.
322. The ICRP has developed models for estimating the dose from the intake of HTO, or other tritiatedcompounds that partially convert to HTO after being taken into the body, for inhalation of tritiatedgases and low soluble particulate tritium, and for OBT. The current ICRP biokinetic models for tritium
intake by workers and members of the public are reasonably consistent with experimental results, andimproved models are under development. There is a practical need for the development of biokinetic
models for intake of tritiated biochemical substrates, including nucleotropic forms.
323. Doses, and hence risk, from some tritiated biochemical substrates and OBT are greater than thosefrom HTO, due to their longer residence in the body and potentially also as a result of their localization
within cells, specifically their proximity to DNA. Direct studies of OBT-related biological effects are
not generally feasible because of low tritium concentrations in OBT. However, there are a few studies
looking specifically at biological effects related to tritiated biochemical substrates, most of them using
DNA precursors and amino acids. There is no appropriate dosimetric model for use in human risk
assessment and radiation protection for tritiated DNA precursors. It should be noted, however, that the
number of workers dealing with these forms of tritium is rather limited.
324. In laboratory studies of mammals, tritium has been shown to induce both stochastic anddeterministic biological effects, consistent with the effects induced by other types of ionizing radiation,and consistent with its generally uniform distribution throughout body tissue, particularly as HTO. The
severity of deterministic effects increases with increasing tissue dose above thresholds, as observedwith other radiation. Exposure to tritium can also induce stochastic effects, such as cancer or heritableeffects, in laboratory mice and rats. However, to date, there is no epidemiological evidence ofstochastic health effects being induced by tritium exposure in humans.
325. A review of the RBE studies of tritium beta radiation indicates a range of values from about unity
to several-fold higher compared to gamma rays and orthovoltage X-rays depending on many factors
such as biological end point, test system, dose and dose rate and choice of reference radiation. RBE
values derived from about 50 in vivo and in vitro experiments on mammals, for different end points,
ranged from 1.0 to 5.0 (centred around 2–2.5) and from 0.4 to 8.0 (centred around 1.5–2) with regard to
gamma rays and orthovoltage X-rays, respectively. Studies also showed a general tendency of RBE
values to increase with lower doses. However, the Committee emphasizes that the ability to draw
specific conclusions for carcinogenic effect in mammals is limited because of the lack of pertinent data.
326. A number of epidemiological studies have been conducted of workers or members of the publicpotentially exposed to tritium. Unfortunately, the majority of these studies do not use the results oftritium monitoring to calculate tritium-specific doses for use in the analyses. This makes it very difficult
to reliably interpret the findings of these studies in terms of tritium-specific risk as distinct from risk
from other types of exposure, principally external sources of penetrating radiation and other internal

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emitters. These limitations apply particularly to studies of members of the public, but also extend tostudies of tritium workers, because any occupational tritium monitoring data that might be available
have not been fully utilized. Consequently, at present, little information of substance on tritium-specificrisk can be derived from epidemiological studies of tritium workers or members of the publicpotentially exposed to tritium beyond the conclusion that tritium-specific risks have not been seriously
underestimated.
327. Epidemiological studies of occupational exposure to tritium offer the best prospect ofinvestigating tritium-specific risk to health, but certain requirements have to be fulfilled if this prospect
is to be realized. First, tritium-specific doses derived from tritium monitoring and other occupational
data need to be calculated from existing records or reconstructed. Second, exposure to other sources of
exposure to radiation—such as external sources of penetrating gamma radiation and intake of otherradionuclides—needs to be taken into account so that analysis in terms of tritium-specific doses may be
adjusted for the presence of other exposure. Third, since the numbers of tritium workers and the tritiumspecific doses they receive are limited, international collaboration is required to achieve reasonablestatistical power in studying tritium workers to meaningfully investigate tritium-specific risk, and suchinternational collaboration should use a common protocol for the determination of tritium-specificdoses. Fourth, the success of an international collaborative study will depend on the numbers of tritium
workers available and the doses they have received, so the chance of achieving meaningful results will
depend on the participation of as many countries as possible in such a study.
328. As far as the effects of tritium exposure of the public are concerned, there is effectively no
information on tritium-specific risk that can be obtained from presently available epidemiologicalstudies. It is unlikely that epidemiological studies of members of the public potentially exposed totritium will produce results that are interpretable in terms of tritium exposure with any acceptable
degree of reliability. This is because environmental tritium exposure is generally very low, and any
effect of such exposure against the background of other risk factors will provide a very small signal oftritium risk against this large background noise.
329. Suggestions that reports of excesses of childhood leukaemia incidence near certain nuclear facilities
could be due to releases of tritium from these installations because of a serious underestimation of the riskof childhood leukaemia from exposure to tritium are implausible. Large quantities of tritium were released
into the environment by atmospheric nuclear weapon testing in the early 1960s and there is no evidence
from childhood leukaemia registration rates following exposure to tritium fallout of any major
underestimation of the risk of childhood leukaemia from exposure to tritium.

XI. ACKNOWLEDGEMENTS
The Committee wishes to acknowledge with gratitude the help of the experts, J. Harrison (United
Kingdom), B. Lambert (United Kingdom) and R. Wakeford (United Kingdom), who were directly
involved in conducting the evaluation, and J. Chen (Canada) for critically reviewing the manuscript. In
particular, the Committee would like to thank M. Balonov (consultant) who mainly elaborated thisscientific annex based on a first draft prepared by K. Bundy (Canada), D. Chambers (Canada), R. Lane
(Canada) and B. Thériault (Canada). The views expressed in the scientific annex remain those of theCommittee and do not necessarily represent the views of individual experts, the United Nations or its
Member States.

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

327

APPENDIX A: TABLES SUMMARIZING STUDIES OF
OCCUPATIONAL AND ENVIRONMENTAL EXPOSURE TO TRITIUM
Tables A1 to A4 present four groups of studies that provide increasing levels of information on tritiumspecific risk.
Table A5 presents studies of environmental exposure to tritium released from heavy-water-moderated
nuclear reactors or tritium production/processing plants.

Table A1. Studies of workforces that include workers (potentially) exposed to tritium though not explicitly identified
Occupational exposure to tritium at installations with heavy-water-moderated nuclear reactors and/or tritium production/processing plants will generally be greater than exposure at installations where there arevery low ambient levels of tritium resulting from, for example, the release of tritium generated in ternary nuclear fission. Consequently, studies of workers exposed to very low levels of tritium at, for example,
nuclear power stations with only light-water-moderated reactors are not included in this table

Summary of study

Summary of findings relating to tritium

Relevance for this report

Cragle et al., and Cragle
and Watkins [C31, C32]

Historical cohort study of mortality among whitemale workers at SRS, USA, first employed
1952−1975. Total 9,860 workers; 1,722 had died
before 1987. SMRs, stratified by employee pay code,
calculated using general population of US whitemales as reference. Mortality trends, particularly forvarious cancers, with recorded external doses.
Annual tritium whole-body doses available, but not
used. Various subgroups examined, but not tritium
workers

About 5,000 workers exposed to tritium, and about
800 workers had recorded dose >0.5 mSv fromtritium. SMRs for all causes and all cancers were lessthan 1.0, which was the case for most causes ofmortality, with no SMR significantly above 1.0.
Marginally significant positive trend of leukaemia
mortality with external dose

Not possible to derive tritium-specific risk, because
tritium workers not analysed separately, andtritium-specific doses not used

Richardson et al., andRichardson and Wing
[R2, R3, R4]

Historical cohort study of mortality before 2003
(5,098 deaths) among 18,883 workers at SRS before
1987. Cumulative recorded whole-body doses fromexternal exposure combined with tritium-specificdoses from urinalysis. Tritium-specific doses alone
not used, and workers monitored for potential
tritium exposure not considered separately

SMRs significantly <1.0 for all causes and all cancers.
Using nested case–control approach, marginallysignificant positive trend of leukaemia (excludingCLL) mortality with cumulative whole-body dose;strongest for myeloid leukaemia. Adjusting
indirectly for smoking, some evidence for positive
trend of lung cancer mortality with cumulative
whole-body dose

Not possible to derive tritium-specific risk, because
tritium workers not analysed separately, andtritium-specific doses not used

McGeoghegan and Binks[M5]

Historical cohort study of mortality (1946–1995) andcancer incidence (1971–1991) among 12,540
workers at Capenhurst, UK. Tritium had beenprocessed before 1988. Annual external wholebody radiation doses were included in the analysis.
Unclear how many workers exposed to tritium.
Tritium workers not analysed separately andinternal doses, such as tritium, were not considered

Most SMRs and SRRs <1.0; several significant
association between bladder cancer incidence andcumulative external radiation dose when dose
lagged 20 years

Tritium workers not analysed separately and tritiumspecific doses not used, so not possible to derive
tritium-specific risk

McGeoghegan and Binks[M6]

Historical cohort study of mortality (1955–1995) andcancer incidence (1971–1995) among 2,628 workers
at Chapelcross, UK. Tritium production from 1980.
Annual external whole-body radiation doses used.
Unclear how many workers exposed to tritium.
Tritium workers not analysed separately andinternal doses not considered

Most SMRs and SRRs <1.0; several significant trends
with cumulative external radiation dose found forprostate cancer; no case monitored for tritium andonly 2 at Chapelcross after tritium production
started

Tritium workers not analysed separately and tritiumspecific doses not used, so not possible to derive
tritium-specific risk

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

Study references329

Summary of study

Summary of findings relating to tritium

Relevance for this report

McGeoghegan et al.,Douglas et al., Omar etal., and Smith andDouglas [D8, M7, O5,
S15]

Historical cohort study of mortality (1947–1992) andcancer incidence (1971–1986) among 14,319
workers at Sellafield, UK. Tritium released from
reprocessing from 1952, and tritium production
during 1955–1962. Doses from external radiation
and Pu included in the analyses, but not tritiumspecific doses. Tritium workers not identified oranalysed separately

SMRs for all causes and all cancers <1.0 for radiation
workers. Significant trends of mortality andincidence for leukaemia (excluding CLL) withincreasing cumulative external dose

Tritium workers not analysed separately and tritiumspecific doses not used, so not possible to derive
tritium-specific risk

Carpenter et al. [C7]

Historical cohort study of mortality during 1946–1988(13,505 deaths) among 75,006 workers at Sellafield,
UKAEA and AWE; 40,761 monitored for externalradiation (6,900 deaths). Tritium doses at Sellafield,
AWE and Harwell generally included with wholebody external dose, but excluded for Winfrith andDounreay. Tritium workers not considered as aseparate group

SMRs for radiation workers and other workers
significantly <1.0 for all causes and all cancers. The
mortality from leukaemia (excluding CLL) increased
significantly with increasing cumulative wholebody dose. No analysis performed for tritiumspecific doses and tritium workers not considered
as a separate group

Although tritium workers included, not consideredseparately and tritium-specific doses not analysedalone, so not possible to derive tritium-specific risk

Johnson et al. [J7]Historical cohort study of mortality before 1997
(6,516 deaths) among 22,543 workers at AWE, UK
during 1951–1982. Tritium-specific doses combinedwith external radiation doses; no analysis of tritiumspecific doses alone. Tritium workers notconsidered separately from other internal dose
workers

SMRs for all causes and all cancers significantly <1.0
for all workers, all radiation workers, and all workers
monitored for internal emitters. For all internal dose
workers, the SMR for kidney cancer was significantly
raised. Significant positive trends with cumulative
external (plus tritium) dose for mortality frommultiple myeloma, bladder cancer and lung cancer(doses lagged 11 years). No mention of prostatecancer [B18]

Neither tritium workers nor tritium-specific doses
considered separately, so not possible to derive
tritium-specific risk

Muirhead et al. [M22]Historical cohort study of mortality (26,731 deaths)
and cancer incidence (11,165 cases) before 2002
among 174,541 radiation workers included in the
UK National Registry for Radiation Workers (NRRW).
Analyses for recorded external doses. Internal doses(including tritium) not included; internal dose
workers identified and either excluded or presence
adjusted for in subsidiary analysis. Tritium workers
not identified and not separately analysed

SMRs significantly <1.0 for all causes and all cancers.
Significantly raised SMR for pleural cancer.
Leukaemia (excluding CLL) and all other cancers
significantly increased with increasing externaldose. Exclusion of internal dose workers increased
slopes for other cancers, but alternatively adjustingfor internal monitoring had little impact

Tritium workers included but not analysed separately
and tritium-specific doses not used, so not possible
to derive tritium-specific risk

Gribbin et al. [G11]Historical cohort study of cancer mortality during
1956–1985 (227 deaths) among 8,977 male workers
employed by Atomic Energy of Canada Limited(AECL) during 1956 to June 1980. Analyses forrecorded external doses. Tritium-specific doses notconsidered. Tritium workers not analysed separately

SMRs significantly <1.0 for all causes and all cancers.
Marginally non-significant positive trend of
mortality from leukaemia excluding CLL andcumulative external dose

No tritium-specific doses and no separate analysis oftritium workers, so not possible to derive tritiumspecific risk

330 UNSCEAR 2016 REPORT

Study references

Study references

Summary of findings relating to tritium

Relevance for this report

Sont et al. [S19]Historical cohort study of cancer incidence during
1969–1988 among 191,333 Canadian workers
monitored for exposure to radiation during 1951–
1988, using the Canadian National Dose Registry(NDR). Tritium doses estimated from routine
urinalysis and added to whole-body externalradiation doses. Tritium-specific doses not analysed
separately and tritium workers not considered as aseparate group

Collective cumulative dose of 1,144.5 man Sv fromexternal gamma radiation compared with that of
122.6 man Sv from tritium. For nuclear workers,
mean cumulative tritium dose received from tritiumof 5.56 mSv compares with 26.43 mSv received fromexternal exposure; for analysis these twocomponents were combined. SIRs for all cancers
significantly <1.0 for males and females. ERR/Sv
whole-body dose for all cancers significantly raised

Tritium-specific doses combined with external doses
and not considered separately, and tritium workers
not considered as a separate group, so not possible
to derive tritium-specific risk. Caution needs to be
exercised because of problems with the NDR data
(see also [Z2])Metz-Flamant et al. [M11]

Historical cohort study of mortality during 1968–2004(6,310 deaths) among 59,021 nuclear workers inFrance during 1950–1994. Analyses of recorded
external photon doses. Tritium-specific doses (andother internal doses) not included. Tritium workers
not analysed separately

Significant trend with external dose for myeloid
leukaemia

Tritium workers not treated as a separate group andtritium-specific doses not used, so not possible toderive tritium-specific risk

Azizova et al., Hunter etal., Moseeva et al., andSokolnikov et al.
[A7, H20, M21, S18]

Historical cohort study of mortality and disease
incidence among workers of the Mayak complex in
the Russian Federation. Tritium-specific doses not
used and tritium workers not considered separately

Study principally concerned with risks from externalradiation and from Pu; doses from this exposure
were high in the early years of operations at the
Mayak complex

Tritium-specific doses not used and tritium workers
not identified, so not possible to derive tritiumspecific risk

Cardis et al., Fix et al., andIARC [C3, F6, I5]

Historical cohort study of mortality (15,825 deaths), in
particular cancer mortality (3,976 deaths), among
95,673 workers at Sellafield, UKAEA and AWE in the
UK, Hanford, Rocky Flats and Oak Ridge National
Laboratory (ORNL) in the USA, and AECL in Canada(the “IARC 3-country study”). For workers at all sites(except Winfrith and Dounreay) tritium-specificdoses included, but combined with recorded
external doses. Separate analysis with tritiumspecific doses not done and tritium workers notconsidered separately

ERR/Sv cumulative dose for all cancers excluding
leukaemia was 0.07 (90% CI: −0.39, 0.30) while
ERR/Sv for all leukaemia excluding CLL was 2.18
(90% CI: 0.13, 5.7)

Neither tritium worker nor tritium-specific doses
considered separately so not possible to derive
tritium-specific risk

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

Summary of study331Summary of study

Summary of findings relating to tritium

Relevance for this report

Cardis et al., Thierry-Chef
et al., and Vrijheid et al.
[C4, C5, T11, V2]

Historical cohort study of mortality among 407,391
nuclear industry workers from 15 countries (the“IARC 15-country study”). Total of 18,993 deaths
included, 5,233 from cancer. Photon doses obtained
from records at individual installations. Tritiumspecific doses included with whole-body doses, andno separate analysis conducted. Tritium workers notconsidered separately

Significant association between cumulative (lagged)mortality
radiationdose
andall-cause(ERR/Sv=0.42), mainly due to a dose-related
increase in all cancer mortality (ERR/Sv=0.97).
Among 31 specific types of malignancies, significant
association for lung cancer (ERR/Sv=1.86) andborderline non-significant associations for multiple
myeloma (ERR/Sv=6.15) and ill-defined andsecondary cancers (ERR/Sv=1.96). Stratification on
duration of employment had a large effect on
ERR/Sv estimates, reflecting a strong “healthy
worker survivor effect” in the contributing cohorts

No information on tritium-specific doses provided
and no separate analysis of tritium workers
conducted, so not possible to derive tritium-specificrisk. 15-country study has to be treated with caution
because of problems with the Canadian worker data
(see also [Z2])Hamra et al., Leuraud etal., Richardson et al., andThierry-Chef et al.
[H5, L9, R5, T12]

Historical cohort study of mortality (66,632 deaths)
among 308,297 radiation workers from nuclear
installations in USA, UK and France. Doses usedderived from recorded external radiation doses.
Tritium-specific doses not included in the study(although in some tritium doses included withexternal doses). Tritium workers not analysed
separately

ERR/Gy for all cancers excluding leukaemia 0.39 (90%
CI: 0.12, 0.67), all leukaemia excluding CLL 2.96 (90%
CI: 1.17, 5.21). Intended that tritium-specific doses
excluded from doses used in analysis; seems likelythat for some installations and for some years
tritium doses included in external dose records

No analysis conducted for tritium-specific doses andtritium workers not considered as a separate group,
so not possible to derive tritium-specific risk

Daniels et al., andSchubauer-Berigan et al.
[D2, S9, S10]

Nested case-control (1:4 matching) study ofleukaemia mortality before 2006 (369 deaths) andradiation exposure among 105,245 workers at six
nuclear sites in the USA. 66 (17.9%) cases and 227(15.4%) controls were exposed to tritium. Internal
doses to red bone marrow (RBM) from urinalysis
results were included with recorded external
photon doses. No separate analysis using tritiumspecific doses and tritium workers not treated as aseparate group

ERR per 100 mGy (total low-LET radiation dose toRBM) 0.09 (95% CI: −0.17, 0.65) for leukaemia
excluding CLL (ERR per 100 mGy for exposure 6–14
years prior to diagnosis 1.9 (95% CI: <0, 8.0)).
Tritium-specific doses to RBM available (mean
0.2 mGy; maximum, 85.1 mGy) but not used alone

Tritium-specific doses alone not used, and tritium
workers were not considered as a separate group,
so not possible to derive tritium-specific risk

332 UNSCEAR 2016 REPORT

Study references

Table A2. Studies of workers monitored for (potential) exposure to tritium, identified and investigated as such, but tritium-specific doses not available or not used
Study reference

Summary of findings relating to tritium

Relevance for this report

Beral et al. [B17]

Historical cohort study of mortality (3,373 deaths
during 1946–1979) in 39,546 workers at UKAEA.
Exposure to tritium occurred at UKAEA sites. Of
20,382 workers exposed to external radiation,
1,418 also potentially exposed to tritium, andtritium workers were considered separately.
Only doses from external radiation used, nottritium doses (apart from Harwell from 1977
when tritium included in whole-body
exposure)

Tritium doses not quantified. All causes SMR for tritium
workers significantly low at 0.59, and all cancers SMR nonsignificantly low at 0.77. On the basis of 6 deaths, prostatecancer SMR significantly raised, at 8.89; all 6 deaths withcumulative external doses ≥50 mSv (SMR=12.77)

Absence of tritium-specific doses does not permit
a reliable conclusion on the potential role oftritium in the raised prostate cancer SMR for thegroup of tritium workers

Carpenter et al. [C8]Historical cohort study of mortality (4,149
deaths) during 1946–1988 among 40,761
external radiation workers at Sellafield, UKAEA
and AWE. 4,111 workers monitored for tritium
analysed separately, but tritium-specific doses
not used, only a flag indicating tritium
monitoring

For tritium workers, SMRs for all causes and all cancers
significantly <1.0, and compatible with other radiation
workers. SMR and RR for prostate cancer non-significantly
raised; testicular cancer the only cancer with significantly
elevated RR buccal cavity and pharynx the only cancer withsignificantly reduced RR. Little evidence that RRs varied withperiod since, or age at, or calendar year of first monitoring orwith age at first monitoring. For UKAEA or AWE workers, data
available on duration of monitoring; among tritiummonitored workers, for prostate cancer, significant variation
with number of years monitored (highest in workers
monitored in 2–4 years). Significantly raised RR for lungcancer among tritium workers with external dose <10 mSv,
but not ≥10 mSv. Non-significantly raised RR (=1.39) for
prostate cancer among tritium workers with external doses≥l0 mSv

Absence of tritium-specific doses does not permit
a reliable conclusion about tritium-specific risk

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

Summary of study333Summary of study

Summary of findings relating to tritium

Relevance for this report

Gillies and Haylock, andMcGeoghegan et al.
[G4, M8]

Historical cohort study of mortality (19,613
deaths) and cancer incidence (10,411 cases)
before 2006 among 64,956 workers employedduring 1946–2002 by British Nuclear Fuels plc(BNFL) at Sellafield, Springfields, Capenhurst
and Chapelcross. 42,431 external radiation
workers, and 22,675 workers also monitored fortritium, Pu and U. 1,757 workers weremonitored for tritium, 1,062 for tritium only.
Analyses with external doses only, not internal
doses (including tritium), but workers flagged
for tritium monitoring. In some analyses,
tritium workers considered separately

SMRs for all causes and all cancers significantly <1.0 for bothradiation and non-radiation workers, but RRs <1.0,
significantly for all causes of death. All causes and all cancers
RRs for internal radiation workers vs. other radiation workers
did not differ significantly from 1.0. Mortality from bothcancer and non-cancer increased significantly with increasing
cumulative external dose; for cancer trend significantly lessfor internal radiation workers than external only radiation
workers, and same pattern for cancer incidence. For tritium
workers, SMR for all causes significantly <1.0, but SMR for all
cancers non-significantly >1.0 (contrasting with SMR of 1.0
for Pu workers and SMR significantly <1.0 for U workers).
Tritium worker SMRs significantly >1.0 for pleural and female
breast cancer. RR comparing all cancer SMR for tritium
workers vs. external only radiation workers significantly >1.0.
Mortality RR for female breast cancer significantly >1.0
(2 deaths among tritium workers) and also for all smokingrelated cancer. All cancer SIR for tritium workers nonsignificantly >1.0 (and non-significantly greater than SIRs for
Pu and U workers). SIRs for pleural, testicular and nonmelanoma skin cancer significantly >1.0. Incidence RR for all
cancers significantly >1.0 for tritium workers vs. external only
radiation workers, and incidence RR significantly raised fornon-smoking-related solid cancer and for solid cancer
excluding lung, liver and bone. Tritium workers stated tohave significantly increasing incidence of digestive cancer
with increasing cumulative external dose

Tritium workers considered separately, and allcancer risk (both mortality and incidence)
significantly greater vs. external only radiation
workers. However, tritium-specific doses not
used so not possible to draw reliable conclusions
about tritium-specific risk

Boice et al. [B22]Historical cohort study of mortality during 1944–
2009 (3,681 deaths) among 7,270 workers at
the Mound nuclear facility, USA, during 1944–
1979. 4,509 workers monitored for externalradiation. Tritium processed at Mound during
1954–1997; 4,134 workers monitored fortritium (1,125 with positive urinalysis result).
Tritium doses estimated on assumption that
intake was HTO, and added to the doses fromother sources. Tritium-specific doses alone notused in analysis

Mean tritium dose for workers with positive monitoring result
8 mSv and maximum 195.5 mSv. (Mean external dose 26 mSv
and maximum 939.1 mSv.) For workers with non-zero tritium
dose, SMRs for all causes and all cancers significantly <1.0.
For specific cancers, SMRs were generally <1.0, including
prostate cancer and (significantly) lung cancer. For dose–
response analysis, with tritium doses included in totalorgan/tissue doses, significant positive trend for oesophageal
cancer and negative trend for liver cancer

Tritium workers did not have unusual patterns ofmortality. Tritium-specific doses incorporated into
total organ/tissue doses in analysis, so that
tritium-specific risk cannot be derived

334 UNSCEAR 2016 REPORT

Study referenceTable A3. Studies of workers monitored for (potential) exposure to tritium, identified and investigated as tritium workers, and occupational dose records used, but
tritium-specific doses not available or not explicitly used
Study reference

Summary of findings relating to tritium

Relevance for this report

Beral et al. [B18]

Historical cohort study of mortality during 1951–1982(3,115 deaths) among 22,552 workers at AWE, UK,before 1983. 9,389 workers monitored for externalradiation; 1,562 also monitored for tritium. Wholebody doses from tritium added to recorded external
doses. Tritium workers considered as a separate
group for some analyses, but tritium-specific doses
not used<2% of workers had a tritium dose >10 mSv. SMRs fortritium workers vs. other radiation workers not
unusual: RRs of 1.02 for all cancers and 0.97 for othercauses (did not vary notably with different doselagging periods). Tritium workers had a significant
trend of prostate cancer with cumulative wholebody dose compared with workers not monitored
for tritium exposure (on the basis of 3 deaths; trend
driven by 1 death with cumulative whole-body dose≥100 mSv). For tritium workers, SMR for prostatecancer non-significantly elevated at 2.50 with
RR 1.27 vs other radiation workers

Very few prostate cancer deaths for tritium workers,
and no analysis with tritium-specific doses alone, so
reliable conclusions cannot be drawn about tritiumspecific risk

Fraser et al. [F12]Historical cohort study of mortality during 1946–1986(5,509 deaths) and cancer morbidity during 1971–
1984 (1,594 cases) in 39,718 UKAEA workers,
including 1,702 workers monitored for tritium,
considered separately. Only recorded external doses
used (except tritium doses at Harwell from 1977
included in whole-body doses)

SMRs for tritium workers compared with SMRs forother radiation workers significantly raised RR only
for prostate cancer, with SMR in tritium workers 2.82
(7 deaths) (see also [B17]). Significant association of
prostate cancer mortality with cumulative externalradiation dose for tritium workers – all 7 deaths withcumulative external doses ≥100 mSv (SMR 5.31).
High case fatality for most cancer led to cancer
morbidity results similar to mortality

Tritium-specific doses not quantified, so no reliableconclusion on tritium-specific risk of prostate cancer(or other cancer) can be reached. Cannot assume
that external dose gives an acceptable measure oftritium dose

Rooney et al. [R14]

Nested case–control study (1:3 matching) of 136
UKAEA workers diagnosed with prostate cancer
between 1946 and 1986, and 404 matched controls.
65% of study subjects had been monitored forexternal radiation (a matching criterion). Monitored
and assessed exposure to tritium (and a number ofother radionuclides) included in analysis. Althoughassessed level of tritium exposure used in someanalyses, tritium-specific doses derived frommonitoring data were not used

Risk of prostate cancer significantly increased in men
monitored for tritium, with RR 14.26 (95% CI: 3.09,
133.16). Significantly raised RR for men working>10 years with heavy-water-moderated reactors.
Significant trend of RR with increasing external dose
for those men likely to have been exposed to tritiumor one of four other radionuclides (Cr-51, Fe-59,
Co−60 or Zn-65). Assessed potential for tritium
exposure, and likely level of exposure, gavesignificant trend of risk with assessed degree of
exposure. When only men assessed as potentiallytritium exposed, but not monitored for exposure,included in analysis, no significant increase inprostate cancer RR. Owing to multiple, not possible
to disentangle the independent effects of tritium,
Cr-51, Fe-59, Co-60 and Zn-65

Exposure to tritium could not be separated fromexposure to four other radionuclides. Absence oftritium-specific doses prevents a reliable estimate of
the tritium-specific risk of prostate cancer

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

Summary of study335

Summary of study

Summary of findings relating to tritium

Relevance for this report

Atkinson et al. [A5, A6]

Historical cohort study of mortality before 1998
(10,249 deaths) among 51,367 UKAEA workers
before April 1996; 26,395 monitored for externalradiation. Some tritium-specific doses included withrecorded external doses, although not for Dounreay
and Winfrith. Tritium workers considered separately
in some analyses. Tritium-specific doses not used

SMRs for all causes and all cancers for non-radiation
workers, radiation workers and internal dose
workers all significantly <1.0. For tritium workers, no
SMR significantly raised, but prostate cancer SMR for
tritium workers significantly higher than for otherradiation workers. Previously reported significant
positive trend of prostate cancer risk withcumulative external dose for tritium workers [F12].
No longer significant and no trend with external
dose for 1980–1997

Absence of tritium-specific doses, so little can bederived about tritium-specific risk

Ashmore et al. [A4]Historical cohort study of mortality during 1951–1987(5,426 deaths) among 206,620 Canadian radiation
workers during 1951–1983, using Canadian NDR.
Tritium doses from urinalysis added to whole-bodyexternal radiation doses. Tritium-specific doses andtritium workers not considered separately

For all causes of death, when tritium doses included
with whole-body external doses, ERR per 10 mSv,
2.6 (90% CI: 1.6, 3.6), with tritium doses excluded (i.e.
external doses only) and, ERR per 10 mSv, 2.5 (90%
CI: 1.5, 3.5). Risk estimates for tritium doses alone
not given. The group of workers monitored forpotential exposure to tritium not consideredseparately

ERR/Sv for all causes of death did not change
substantially when tritium doses included withexternal doses. Separate results for tritium-specificdoses alone not presented, so little can beconcluded about tritium-specific risk. Caution needs
to be exercised because of problems with the NDR
data used (see [Z2])Zablotska et al. [Z1]Historical cohort study of mortality during 1957–1994(1,599 deaths) among 45,468 Canadian nuclear
industry workers (from AECL, Ontario Hydro, Hydro
Québec and New Brunswick Power) during 1957–
1994, using Canadian NDR tritium doses fromurinalysis and added to recorded external radiation
doses. Tritium-specific doses and tritium workers notconsidered separately

Mean cumulative dose 13.5 mSv; among workers with
non-zero doses, mean 19.7 mSv. SMRs for all causesof death and all deaths from cancer significantly<1.0. ERR/Sv total cumulative dose marginally nonsignificantly positive for all solid cancers combined
and marginally significantly positive for all
leukaemia excluding CLL; ERR/Sv show small
increases (from 2.67 to 2.80, and from 16.3 to 18.9,
respectively) when tritium doses included withexternal doses

ERR/Sv for all solid cancers and all leukaemia did notmaterially change when tritium doses excluded
from whole-body doses, but tritium-specific doses
alone not used, so not possible to draw reliableconclusions about tritium-specific risk. Caution
needs to be exercised because of problems with the
NDR data used (see also [Z2])Schubauer-Berigan et al.
[S11]Historical cohort study of mortality (41,508 deaths)
among 119,195 radiation workers at five nuclear
sites in the USA. Tritium-specific doses fromurinalysis added to recorded external doses but
separate analysis with tritium-specific doses alone
not done – separate analysis for all cancers and all
haematopoietic and lymphatic cancers with bothneutron and tritium doses excluded fromcumulative doses. Tritium workers not consideredseparately

SMRs for all causes and all cancers significantly <1.0.
ERR per 10 mSv whole-body dose 0.14% (95% CI:−0.17%, 0.48%) for all cancers combined and 2.0%
(95% CI: 0.71%, 3.5%) for all haematopoietic andlymphatic cancer. When neutron and tritium doses
together excluded from the whole-body dose, ERR
per 10 mSv became 0.18% (95% CI: −0.14%, 0.53%)
and 2.0% (95% CI: 0.73%, 3.7%), respectively

Absence of tritium-specific doses alone, and tritium
workers considered separately, so little can bederived about tritium-specific risk

336 UNSCEAR 2016 REPORT

Study referenceTable A4. Studies of workers monitored for (potential) exposure to tritium, identified and investigated as tritium workers, and tritium-specific doses available and used inanalyses so that tritium-specific risk may be examined explicitly
Study reference

Summary of findings relating to tritium

Relevance for this report

Hazelton et al. [H8]Historical cohort study using a biologicallybased analysis of lung cancer incidence during
1969–1988 (400 cases, 322 in men) among
191,042 Canadian radiation workers during
1951–1988, making use of Canadian NDR.
Tritium-specific doses used in analysis, mainly
in combination with whole-body gamma-ray
doses

Of collective dose from external gamma radiation andtritium combined, tritium contributes 9.0%. 95,430 males,
60,677 with non-zero doses (mean cumulative gamma
plus tritium dose, 18.2 mSv), 9,013 with non-zero tritium
doses (2 253 with tritium doses >14.95 mSv). Significant
dose–response for men with gamma-ray and tritium
doses combined, and with gamma-ray doses alone, but
for tritium doses alone dose–response only marginallysignificant. Allowing RBE for the tritium absorbed dose tovary did not improve fit significantly. When dataset
restricted to 69,826 men not flagged for neutron
exposure, dose–response for tritium doses alone notsignificant. For 95,603 women (44,238 with non-zeroradiation doses, mean cumulative gamma plus tritium
dose 3.8 mSv), dose–response non-significant, but
consistent with dose–response for men. For men,modelling predicts ~31 cases attributable to gamma
radiation and ~2 cases to tritium exposure

Analyses performed with tritium-specific doses alone,although most results for gamma and tritium doses
combined. Modelling predicts ~2 lung cancer cases
attributable to tritium exposure. Caution needs tobe exercised because of problems with the NDR
data used (see [Z2])Zablotska et al. [Z2]

Review by Canadian Nuclear Safety Commissionof employment and dose records for Canadian
nuclear workers led to corrections andimprovements to records. Mortality (489 cancer
deaths during 1956–1994) in revised historicalcohort of 45,468 nuclear workers (originallystudied by [Z1]) reanalysed. Particular attention
paid to accuracy of records for AECL workers
before 1965. Tritium doses from urinalysis.
Summary doses for whole-body external(gamma) and internal (mainly tritium) radiation
used for risk analysis. Workers with neutron orhigh internal exposure excluded. Tritium
workers not considered separately

42,228 workers first monitored since 1965 had no
significant ERR/Sv for solid cancer (-1.20) or leukaemia(14.4, p=0.28). This contrasted significantly with ERR/Sv
for solid cancer in 3,088 AECL workers first monitored
during 1956–1964: ERR/Sv=7.87 (p <0.01) (but no doserelated risk of leukaemia). Very likely that dose
information in Canadian NDR incomplete for early AECL
workers. Mean cumulative tritium dose in revised cohort
3.02 mSv, 14% of mean cumulative total dose;
809 workers had tritium doses >50 mSv. When in revised
cohort tritium doses added to risk model for solid cancer
with gamma doses as second linear term, fit of model did
not improve; in model with two separate dose terms, forgamma component ERR/Sv was 2.56 (95% CI: −0.11, 6.79),
while for tritium component ERR/Sv was −4.71 (95% CI:<−5.92, 8.58). For individual cancer types, adding tritium
doses to risk model did not improve fit, and risk duesolely to gamma doses

Analysis of mortality from all solid cancers, individual
solid cancers and leukaemia showed risk due solely
to whole-body gamma doses and that the additionof tritium doses did not improve fit of the model.
Study provides estimates of tritium-specific risk, but
uncertainties associated with estimates are large
due to generally low tritium doses received by alimited number of workers

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

Summary of study337

Summary of study

Summary of findings relating to tritium

Relevance for this report

Hamra et al. [H3, H4]

Reconstruction of tritium-specific annual doses
for workers at SRS, USA, using tritium urinalysis
data, recorded external doses, and a jobexposure matrix approach [H3]. Reconstructed
tritium-specific doses used in a study oftritium-specific leukaemia risk [H4], using
Bayesian approach informed by experimental
studies of tritium

From 75,523 dose records for 1954–1978 the proportion of
the whole-body dose from tritium was calculated forvarious jobs, areas and time periods, and tritium doses
assigned for 43,590 person-years. Under a strong
assumption that the leukaemia risk per 10 mGy fromtritium is always greater than that from external gamma
radiation, authors derived ERR per 10 mGy of 0.298 (90%credibility interval: 0.027, 0.702) for leukaemia and 0.344(90% credibility interval: 0.049, 0.817) for leukaemia
excluding CLL. Risks obtained without the restriction thatthe leukaemogenic effect of tritium is always greater thanthat of external gamma radiation leads to values of 0.141(90% credibility interval: −0.323, 0.649) and −0.281 (90%credibility interval: −1.136, 0.548), respectively

Illustrates what may be done using occupational
records to derive tritium-specific doses for use inepidemiological analysis. Tritium-specific leukaemia
risk is uncertain because of limited data, and thedependence of leukaemia ERR/Sv upon the strong
assumption that the risk per unit absorbed dosefrom tritium is always greater than that for gamma
radiation means that the findings must be treatedwith caution

HSE [H17, H18]

Case-control study investigating exposure offathers working at Sellafield before theconception of their children and an increased
risk of leukaemia and non-Hodgkin’s
lymphoma (LNHL) in these children. Exposure
considered included tritium, both assessed
potential for exposure and doses derived fromcontemporary tritium monitoring data

Highly significant association between risk of LNHL in
offspring and assessed potential for paternal exposure totritium in preconceptional period. Association not
reproduced when doses derived from measured
exposure to tritium based on contemporary monitoringdata used

Paternal preconceptional tritium doses derived frommonitoring data do not indicate raised risk of LNHLin offspring. Risk indicated by assessed potential fortritium exposure must be treated with caution given
difficulties of retrospective assessment of potential
for historical tritium exposure

McLaughlin et al. [M9]Matched case-control study (1:8 matching) ofchildhood leukaemia (112 cases during 1950–
1988) and paternal preconceptional exposureof workers in Ontario, Canada. Analyses
included external and tritium doses fromoccupational records

No association between childhood leukaemia risk andpaternal preconceptional radiation dose found for either
recorded external whole-body dose or tritium dose. For
tritium exposure, no father of an affected child wasrecorded as exposed while 14 control fathers had
preconceptional exposure

No indication of measured paternal preconceptional
tritium exposure increasing the risk of childhood
leukaemia in offspringGreen et al. [G10]

Matched case-control study (1:1 matching) ofcongenital anomalies in offspring of workers of
Ontario Hydro. Preconceptional tritium doses
based on monitoring data

No association between risk of congenital anomalies andthe tritium dose received during the preconceptional
periodTritium doses received prior to conception did notincrease the risk of congenital anomalies in the child

338 UNSCEAR 2016 REPORT

Study referenceTable A5. Studies of environmental exposure to tritium released from heavy-water-moderated nuclear reactors or tritium production/processing plants
Study reference

Summary of findings related to tritium

Relevance for this report

Wanigaratne et al.
[W5]Historical cohort study of cancer incidence 1986–2005
using data from the Ontario Cancer Registry, amongpeople living near Pickering heavy water CANDU
nuclear reactor site in Ontario in 1985 and incomparison area (north Oshawa) further from site.
Atmospheric tritium concentrations estimated foreach 1985 residential location using atmosphericdispersion model with tritium discharge andmeteorological data. Model predictions comparedwith monitoring results. In addition to all cancers
combined, leukaemia and cancer of the lung, thyroid,
female breast and of childhood examined

More than half of Pickering and all north Oshawa
residents experienced modelled average tritium
concentration levels <2.9 Bq/m2, representing annual
effective dose of 0.47 µSv for average adult. The allcancer SRR significantly less than 1.0 for both cohorts.
Effect of emigration from Ontario not assessable. Only
for female childhood cancer in Pickering, SRR
significantly raised. For Pickering residents living at thesame address in 1985 as in 1979 (“non-movers”),
assessed tritium exposure not associated with risk oflung cancer or female breast cancer (other cancer casenumbers too small for this analysis)

This study used assessed tritium-specific exposure
residents at the same address for at least six
years for analysis of cancer incidence. However,doses based on residential histories notreconstructed. Tritium-specific dose estimatesvery small (maximum annual effective dose,
2.36 µSv), and number of cancer cases (available
for just one installation) limited, so power ofstudy to detect any risk was low. Nonetheless,study attempts to address tritium-specific cancer
risk, not done in any other environmental
exposure study

Grosche et al. [G13]Comparison of childhood leukaemia incidence aroundKrümmel nuclear power station, Germany and SRS,
USA. Study conducted because of suggestion that
tritium discharges from Krümmel (a boiling (light)
water reactor) responsible for marked excess ofchildhood leukaemia cases in the vicinity of the site. In
contrast, SRS produced/processed relatively large
quantities of tritium for weapons

If releases of tritium from Krümmel site responsible forexcess childhood leukaemia cases in immediate vicinity,
then the much larger releases from SRS would
detectably increase childhood leukaemia risk inneighbourhood. Around SRS, however, statistically nonsignificant deficit of childhood leukaemia incidence
found. Therefore, theory of childhood leukaemia excess
around Krümmel due to tritium exposure not supportedby this study

Absence of human monitoring data and assessed
tritium-specific doses from this study limits
possible conclusions on tritium-specific risk.
Nonetheless, absence of a detectable increased
risk of childhood leukaemia around SRS does notsupport serious underestimation of the tritiumspecific risk of childhood leukaemia

Cragle et al., Cragle
and Watkins, andMcLaughlin et al.
[C31, C32, M9]Geographical study of childhood leukaemia incidence(1964–1986) and mortality (1950–1987) within 25 km
of nuclear installations in Ontario, Canada, inparticular, the heavy-water-moderated CANDU reactor
power stations

For children born within 25 km of the nuclear power
stations, (marginally) non-significant excess ofchildhood leukaemia mortality and a non-significant
excess of childhood leukaemia incidence. Nonsignificant excess of childhood leukaemia mortality
among those resident within 25 km of the nuclear
power stations

In the absence of assessed tritium-specific doses
to individuals, little may be concluded from thisstudy about the tritium-specific risk of childhood
leukaemia

Johnson and Rouleau[J6]

Geographical study of birth abnormalities, perinatal andinfant mortality during 1971–1988 within 25 km of
Pickering heavy-water-moderated CANDU reactor site
in Ontario, Canada. Health outcome data analysed forairborne and waterborne tritium emissions from thesite, and using ground-monitored airborne tritium
concentration data

No unusually high mortality or abnormality rates found instudy area. Only association between tritium releaselevels and birth abnormalities was for CNS
abnormalities, but not reproduced using groundmonitoring data. Although some evidence of elevatedrisk of Down’s syndrome around Pickering, noconsistent associations with tritium release levels andground monitoring data foundLack of assessed individual exposure to, or doses
from, tritium means no substantial conclusions
on tritium-specific risk may be drawn from thisstudy

ANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

Summary of study339

Summary of study

Summary of findings related to tritium

Relevance for this report

Richter and Stockwell[R10]

Cancer mortality (during 1980–1991) among residentsof Lamar County, Mississippi, near the Salmon(underground) nuclear test site, following two nuclear
tests in 1964 and 1966. Residents worried that tritium
released due to these two explosions detectably
increased cancer risk in vicinity

No increase in environmental tritium levels as a result of
the nuclear tests detected. Observed cancer mortality
rates for Lamar County no different from those
expected for all Mississippi. No association between
cancer mortality rate and distance from detonations

Lacking measured increases of exposure to tritium
resulting from the nuclear tests, this study
provides no information on tritium-specific risk.
Conducted as a public reassurance exercise

340 UNSCEAR 2016 REPORT

Study referenceANNEX C: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—TRITIUM

341

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Corrigendum to Sales No. E.17.IX.1
18 May 2018

Sources, Effects and Risks of Ionizing Radiation:
UNSCEAR 2016 Report to the General Assembly,
Scientific Annexes A, B, C and D
Corrigendum
Annex D, page 438, paragraph 319, last line
should read
(18.6 mBq/d) and not (18.6 mBq/L)

V.18-03158*1803158*

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM

363

CONTENTSI.

INTRODUCTION ................................................................................................................................. 365

II. SOURCES AND LEVELS .................................................................................................................... 367
A. Natural sources .......................................................................................................................... 367
B. Artificial sources ........................................................................................................................ 374
III. PHYSICAL, RADIOLOGICAL AND CHEMICAL CHARACTERISTICS ..................................... 378
A. Physical and radiological characteristics .......................................................................... 378
B. Chemical characteristics......................................................................................................... 379
IV. HUMAN EXPOSURE .......................................................................................................................... 380
A. Exposure of members of the public ................................................................................... 380
B. Occupational exposure .......................................................................................................... 382
C. Measurement of uranium ...................................................................................................... 382
V. BIOKINETICS........................................................................................................................................ 383
A. Inhalation..................................................................................................................................... 384
B. Ingestion ...................................................................................................................................... 385
C. Absorption .................................................................................................................................. 386
D. Systemic distribution, retention and excretion ............................................................. 386
E. Materno-fetal transfer ............................................................................................................. 387
VI. DOSIMETRY ......................................................................................................................................... 388
VII. BIOLOGICAL EFFECTS ...................................................................................................................... 390
A. Chemical versus radiological toxicity ................................................................................ 390
B. Kidneys ......................................................................................................................................... 391
C. Bone............................................................................................................................................... 397
D. Lungs ............................................................................................................................................. 399
E. Liver ............................................................................................................................................... 400
F. Brain............................................................................................................................................... 401
G. Reproduction and development ........................................................................................ 405
H. Other organs .............................................................................................................................. 408
I.

Relative biological effectiveness ......................................................................................... 414

J.

Lethal effects .............................................................................................................................. 414

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VIII. EPIDEMIOLOGICAL STUDIES......................................................................................................... 416
A. Studies of occupational exposure ...................................................................................... 416
B. Studies of Gulf War veterans ................................................................................................ 433
C. Studies of environmental exposure................................................................................... 434
IX. RESEARCH NEEDS ............................................................................................................................. 437
X. GENERAL CONCLUSION ................................................................................................................. 438
XI. ACKNOWLEDGEMENTS .................................................................................................................. 439
APPENDIX A. TABLES SUMMARIZING URANIUM LEVELS IN WATER, STUDIES OF
OCCUPATIONAL AND ENVIRONMENTAL EXPOSURE TO URANIUM ...................................... 441
REFERENCES............................................................................................................................................... 467

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 365

I. INTRODUCTION
1.
This annex provides a review of the scientific literature on characteristics of uranium, its
biokinetics and dosimetry within the human body for various physical and chemical forms and routes ofintake into the body, radiobiological and toxicological effects of exposure to uranium, andepidemiological studies of nuclear workers and the public who have been exposed to uranium.
2.
Uranium was discovered by Martin Heinrich Klaproth in 1789 and its radioactive properties by
Antoine Henri Becquerel in 1896. Uranium, element 92 in the periodic table, is present naturally in allrock and soil. Levels of uranium content in soil depend on local geology and range widely from a
few mg/kg up to levels of several per cent in ore bodies. The Committee in its UNSCEAR 2008 Report[U10] reported a median activity concentration of around 30 Bq/kg (1.2 mg/kg) for uranium in rock andsoil. Uranium is released to the environment through natural events such as forest fires and volcanoes
and released from rock and soil through natural processes. It is distributed through mechanisms such asleaching to ground and surface water and through wind erosion of soil. In turn, uranium in water, soil
and air is taken up by plants and animals. People may be exposed to uranium by inhalation of airborne
particulates, through skin uptake and through ingestion of uranium in food and water.
There are three naturally-occurring, alpha-particle emitting, isotopes of uranium: 238U, 234U and235
U. Two of these, 238U and 235U, with radioactive half-lives of 4.47 × 109 and 7.04 × 108 years
respectively, are the parents of radioactive decay chains that are major contributors to the backgroundradiation exposure of the human population. Uranium-238 supports 14 decay products. The isotope
234
U, with a half-life of 2.45 × 105 years, is a member of the 238U decay chain. In natural uranium, 238U
is the most abundant isotope in terms of mass (99.2742%), while 234U and 235U constitute only 0.0054%
and 0.7204%, respectively [N8, S15]. Figure I shows a simplified radioactive decay chains for 238U.
Other isotopes, such as 232U, may be produced in thorium breeder reactors. Further, 236Uranium, with ahalf-life 2.35 × 107 years, is present in spent nuclear fuel and in reprocessed uranium [W24], and occurs
naturally as a very small component of natural uranium (<10−11% by mass).
3.

4.
Both 238U and 234U, when in secular equilibrium, contribute 48.9% of the total alpha particle
activity of natural uranium, while 235U contributes 2.2%. Some nuclear reactors require fuel that isenriched to the fissionable isotope 235U. Current technologies for enriching natural uranium are gaseous
diffusion and centrifugation. Enrichment increases the proportion of 235U from its natural levels(0.72%) to 2–5%, depending on the design requirements of nuclear power reactors. In addition, higher
enrichment levels (>90% of 235U) are achieved for use in weapons. The term depleted uranium (DU)
refers to isotopic mixtures that contain a lower percentage of 235U than is present in naturally occurring
uranium. It is recovered as a by-product of the enrichment process. The proportion of 235U in DU is
between 0.2 and 0.3%. Reprocessed uranium (especially from earlier military reprocessing) may also be
contaminated with traces of fission products and transuranic elements [W24].
5.
Uranium compounds exhibit differences in their chemical and physical properties and, as a result,also differ in their toxicological properties. For example, uranium compounds vary widely in theirsolubility and this can result in differences in bioavailability following intake (via inhalation oringestion) into the body [A31, L9, S37, U16]. The biological and health effects of uranium are due toits chemical and radiological toxicity. In general, this toxicity, as demonstrated in animal studies, iscaused by chemical rather than radiological components, excepting that effects induced by the isotopesof higher specific activity and by enriched uranium are more probably due to radiation exposure.

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6.
Since 1949, many animal studies have indicated that the toxicity of uranium is due mainly tochemical damage to the kidneys [A25]. Other systems or organs may also be affected by exposure touranium, such as the skeleton [A26], the lungs [L18], the gonads [A19] and the liver [P6].
Figure I. Radioactive decay chain for 238U [I17]
Half-life is expressed in a = year; d = day; h = hours; m = minutes; s = seconds

7.
Uranium concentrations in environmental media may be measured in terms of radioactivity(measured in Bq/L, e.g. by alpha spectrometry) or mass (measured in µg/L, e.g. by high-resolution
inductively coupled plasma mass spectrometry). Consequently, data on uranium levels in soil, air, waterand food are indicated in Bq/L and in µg/L.
8.
A general concept is the relation between radioactivity and mass. As mentioned above, theactivity of each member of a chain headed by a parent radionuclide would be the same under conditionsof secular equilibrium, but the mass of each member of the chain would be quite different. The
relationship between activity, A, and mass, M, of a radionuclide is given by:
2
/

where



A =
A0 =
AW
 =
T1/2
M =

activity of a radionuclide, Bq;
Avogadro’s constant, 6.023 × 1023 atoms/mole;=atomic weight of the radionuclide, kg/mole;decay constant, dis/(atom s);=
half-life of the radionuclide, s;mass of the radionuclide, kg.

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 367

For natural uranium, the activity is 25,400 Bq/g (table 1). However, as shown in table 1, the natural
relative abundance of 234U, 235U and 238U can be expressed in terms of either numbers of atoms or
weight, giving slightly different values [K3, M31, M32].
Table 1. Mass activities of the three natural isotopes of uranium [K3]
Relative abundance
Natural uranium
238
235
234

atoms% (wt%)

mBq/μg U

U

99.274 (99.284)

12.40

U

0.720 (0.711)

0.60

U

0.0054 (0.0053)

12.40

II. SOURCES AND LEVELS
A. Natural sources
1.

Levels in soil

9.
Naturally occurring radionuclides in the environment affect the levels of background radiationencountered at different locations around the world [N2, U8]. As mentioned by Cuney [C38], three
types of deposits contain more than three quarters of the worldwide uranium resources: unconformityrelated deposits, iron oxide–copper–gold (IOCG) deposits, and sandstone-hosted deposits [L10].
10. The concentration of uranium in soil varies with location and local geology. Its concentration isrelatively low in basic rock, such as basalt, and higher in acid rock, such as the sedimentary rocksaturated with silica. The uranium content of granites is higher still [U10]. For example, the nominal
activity concentration of uranium in soil is about 15 Bq/kg of 238U (1.2 mg/kg) with a typical activity
range of 10–50 Bq/kg (0.4 to 2 mg/kg) [N2, U10]. Much higher concentrations are found in uranium
mining areas such as the Northern Saskatchewan in Canada, the Colorado Plateau and central Florida
where phosphate is mined. The uranium content of phosphate rock used for phosphate fertilizers ranges
from about 50 to 2,400 Bq/kg of 238U (4–190 mg/kg) [A31, N2, R20, U8]. One of the highest activity
concentrations worldwide is localized in the region of Recife in Brazil, with sedimentary rock that
contains 30–500 mg/kg with an average of 150 mg/kg (1,860 (range 372–6,200 Bq/kg of 238U)) [S18].
However, since uranium in soil may be more or less tightly bound depending on soil characteristics, theuranium speciation in soil has an impact on bioaccessibility in the gut.

2.

Levels in air

11. Soil particles containing uranium may be transferred into the atmosphere through natural
mechanisms. The natural uranium concentration in air is typically very low, varying from location tolocation according to local ground sources [H19]. Airborne uranium can deposit on soil, plants andopen water as dry or wet deposition [A31].

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12. Golchert et al. [G16] measured airborne concentrations of 238U of around 0.3 Bq/m3 at a site near
the Argonne National Laboratory (Illinois, United States). Average levels of natural uranium in ambient
air have been reported to be 0.25 µBq/m3 of 238U (0.02 ng/m3) in Tokyo [H19]. Tracy and Prantl [T17]found the average concentration of 238U in air in a southern Ontario rural environment to be about
1.25 µBq/m3 (0.1 ng/m3), on the basis of measurements of 226Ra in dust and an assumption ofequilibrium between 238U and 226Ra. Taken together, these different values indicate an average uranium
level in air of around 1 µBq/m3.
13. The World Health Organization (WHO) [W14] estimates that an adult of average size inhales
20 m3 of air per day with a nominal natural uranium concentration of 0.6 µBq/m3 of 238U (0.05 ng/m3),
corresponding to 12.4 µBq (1 ng) of 238U. These values lead to a calculated annual intake throughinhalation by adults of approximately 0.0045 Bq of 238U. For comparison, tobacco smoke (from two
packages of cigarettes per day) contributes to 0.11 Bq of 238U (corresponding to 9 µg) of inhaled naturaluranium per year [L49].

3.

Levels in water

14. As long as uranium is inside undisturbed crystalline rock in secular equilibrium with its progeny,
the ratio of 234U to 238U is expected to be one. Nevertheless, disequilibrium can be observed when rockis disturbed by chemical or physical processes involving water. As a result, water from any source may
contain 234U/238U ratios greater than unity because of the greater mobility and increased availability of
234
U, generally due to a reducing environment [D1]. Two isotopes in the decay series, 234Th and 234mPa,
separate the two uranium isotopes and their different solubilities in the source rock, permitting 234U to
be released preferentially and leading to variations in the ratio 234U/238U [O9]. In addition, the total
uranium activity present in water may influence the 234U/238U ratio. Indeed, Ortega et al. found that
~80% of samples with a high ratio of disequilibrium (>1.6) were linked to the lowest uranium activities<50 mBq/L, when the samples with a low activity ratio (<1.5) corresponded to samples with highconcentrations of uranium (>200 mBq/L) [O8].
15. Uranium is present in different water sources (surface water, groundwater and drilled water) at
variable levels [W16]. In oxygenated surface water, uranium levels were found at around 0.02 to 6 µg/L
(0.25–76 mBq/L of 238U). In sea water, its average content is 3.3 µg/L (42 mBq/L of 238U), often boundby ligands or associated with suspended particles [B42]. Natural uranium levels were found to behigher in Precambrian rock aquifers (average, 115.6 µg/L (1.45 Bq/L of 238U)) than in Palaeozoic
sedimentary rock aquifers (average, 3.5 µg/L (0.045 Bq/L of 238U)). Cothern and Lappenbusch [C34]reviewed the available data on the occurrence of uranium in surface and groundwater supplies in the
United States and reported that surface water samples derived from about 35,000 sources had an
average uranium concentration of 18 mBq/L of 238U (1.45 µg/L) (range from 0.18 to about
12,500 mBq/L of 238U (0.014–982 µg/L) and that about 55,000 samples of groundwater supplies had an
average uranium concentration of 55 mBq/L of 238U (4.4 µg/L) (range −0.018–12,000 mBq/L of 238U
(0.0014–942 µg/L)).
16. The WHO indicated values for uranium levels in water generally less than 12.4 mBq/L of 238U(<1 µg/L) [W15, W18]. The Agency for Toxic Substances and Disease Registry (ATSDR) reported anaverage concentration of 14.4 mBq/L of 238U (1.16 µg/L) [A31]. This was much higher than thepreviously reported value by the Committee in its UNSCEAR 1977 Report 0.54 mBq/L of 238U
(0.044 µg/L) [U7]. Thus, it might be more appropriate to report median rather than mean values
because of the large variation of uranium concentration in water. The Committee reported a variation ofnatural uranium concentrations measured in drinking water samples in 16 countries of about eightANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 369

orders of magnitude [U10]. A global overview of uranium concentration (µg/L) and activity (mBq/L) is
given in appendix A, in table A1 for the values measured in groundwater, in table A2 for the values
measured in surface water, in table A3 for the values measured in public water supplies and in table A4
for the values measured in bottled mineral water.
17. In Canada, different surveys aimed to measure uranium levels in drinking water in different
provinces. The mean natural uranium concentration in surface water and groundwater (some treated)
supplies was about 50 mBq/L of 238U (4 µg/L) in southern-central British Colombia [P32], 124 mBq/Lof 238U (10 µg/L) in south-eastern Manitoba [B21], 65 mBq/L of 238U (5.2 µg/L in the Kitigan Zibi First
Nation community in Quebec) [M58], and 5 mBq/L of 238U (0.40 µg/L) in Ontario [O3]. In summary,the analysis of the different values measured in these surveys indicates that mean uranium levels in
drinking water were extremely variable in the different Canadian counties/provinces, from 5 to
750 mBq/L of 238U (0.4–58.3 µg/L), including great internal variations depending on the precise
location. Furthermore, behind these average uranium levels in drinking water, more extreme values
were measured in some Canadian provinces or some counties in the United States as indicated inappendix A, table A1. In fact, natural uranium concentrations as high as 8,680 mBq/L of 238U
(700 µg/L) were found in private groundwater supplies [M64, M65]. A value of 25,048 mBq/L of 238U
(2,020 µg/L) was measured in groundwater in south-eastern Manitoba [B21].
18. Concerning water in the United States, official reports indicated an average natural uranium
concentration of 31.6 mBq/L of 238U (2.55 µg/L) in drinking water from 978 sites in the 1980s [U14].
These values are higher than those mentioned in a study by Fisenne et al. with mean activity of naturaluranium in drinking water in New York City ranging from 0.62 to 1.25 mBq/L of 238U (0.05 to
0.09 µg/L) [F5]. In this study, New York city tap water had 234U, 235U and 238U activities of 1.040.19,
0.0350.010 and 0.870.18 mBq/L, respectively [F5]. Maximum values were measured in Connecticut
86.472 mBq/L of 238U (7,780 µg/L). As for Canada, differences may relate to geographical variations
and local geology (appendix A, table A1).
19. In Finland, national and local surveys of uranium content in water distributed by Finnish
waterworks have been conducted. The median value was of 1.9 mBq/L of 238U (0.15 µg/L) [T22](appendix A, table A1). An extreme value of 114,100 mBq/L of 238U (9,200 µg/L) was measured in theSouth of Finland [M66]. It is noteworthy that although the uranium concentrations in Finnish wells
drilled in bedrock are among the highest in the world [K26, M66, M67, P28], the uranium
concentration in water distributed by the waterworks is generally low [M67, T22].
20. In France, periodic reports address the levels of the radiological quality of drinking water [I21].
Measurements performed during 2008 and 2009 indicate that 226Ra and uranium isotopes constitute themain contributors to a total alpha activity above 0.1 Bq/L; in this case, the mean value of uranium
concentration was 2.22 µg/L (27.5 mBq/L of 238U) with a range from 0.14 to 114 µg/L
(1.8-1,450 mBq/L of 238U) [I21].
21. Activity concentrations of natural radionuclides in soil, food, natural and drinking water were alsomeasured in China [P2]. However, measurements of uranium were not reported for food and drinking
water. Except for salt water lakes that presented higher uranium levels (22 µg/L or 272.5 mBq/L of
238U), the values for uranium in freshwater lakes, reservoirs, rivers, hot and cold springs, well water andsea water were similar (2.2 µg/L with a range from 0.87 to 3.82 (27.3 mBq/L of 238U with a range from10.78 to 47.3 mBq/L of 238U).
22. The guidance levels of radionuclide concentration provided in the WHO Guidelines for DrinkingWater Quality are based on an individual dose criterion (IDC) of 0.1 mSv committed effective dosefrom one year’s consumption of drinking water. They are expressed as activity concentration for a

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UNSCEAR 2016 REPORT

given isotope (Bq/L) and were calculated by dividing the IDC of 0.1 mSv per year by the product of theisotope dose conversion factor (Sv/Bq) and an assumed water consumption of 2 L per day (i.e. 730 L
per year). The guidance levels of radioactivity concentration for 238U and 234U have been rounded to
10 Bq/L and 1 Bq/L, respectively [W18].
23. Due to the fact that the uranium chemical toxicity is generally of greater importance thanradiological effects, several national and international guidelines refer to concentrations of uranium indrinking water, as indicated in table 2. Guideline values (in mg/L or µg/L) were derived from the total
tolerable daily intake (TDI) expressed in mg/kg or µg/kg of body weight (e.g. 60 kg for an adult usedby WHO), itself based on the no observed adverse effect level (NOAEL)1 or lowest observed adverse
effect level (LOAEL) for kidney toxicity, divided by an uncertainty factor of 100 (for intra- andinterspecies variation), and taking the daily drinking water consumption into account (~2 litres) [W16].
Table 2. National and international guidelines for uranium content in drinking water
Only chemical aspects of uranium toxicity are addressed in these guidelines

Organizations/countries

Uranium in drinking water (μg/L)

Reference

Australia

17

[N5]

Bulgaria

60[E2]

Canada20[H16]

Finland100[E2]

Germany

10[B27]

Slovenia

6.8

[E2]

USA

30

[U17]

WHO

30

[W18]

24. The WHO chemical guideline value for uranium in drinking water significantly increased from
2 µg/L in 1998 up to 15 µg/L in 2004 and then to 30 µg/L in 2011 [W18]. The current WHO chemical
guideline of 30 µg/L is still designated as provisional because of scientific uncertainties regarding
uranium toxicity, notably with regard to possible carcinogenic effects of uranium [A17] and specificsensitivity of some groups, such as children or people with hypertension or osteoporosis [F15].
25. Tables A1–A4 in appendix A show that uranium concentrations may exceed guideline values inseveral countries, including those of water from public supplies. In a study of 476 Norwegiangroundwater samples, 18% had natural uranium concentrations in excess of 20 μg/L (0.25 Bq/L of 238U)[F12]. Natural uranium concentrations in groundwater in excess of 20 μg/L (0.25 Bq/L) have also beenreported in parts of New Mexico, the United States [H5], central Australia [F9] and France [I21]. Some
Finnish studies noted a median uranium concentration of 28 μg/L (0.35 Bq/L of 238U) and 285 μg/L
(3.5 Bq/L of 238U) in drinking water [K26, P28], respectively. In Canada, one study also reported highlevels of uranium concentration up to 845 µg/L (10.5 Bq/L of 238U) in private wells [Z8].
26. In a Canadian study, Zamora et al. [Z6] found that water contributed 31–98% of the total daily
intake of uranium from food and water for individuals whose drinking water contained uranium atconcentrations ranging from 2 to 780 µg/L (25 to 10,000 mBq/L of 238U). This was similar to values
1
NOAEL: the greatest concentration or amount of a substance that causes no detectable adverse alteration in morphology,functional capacity, growth, or life span of the target organism under defined conditions of exposure [W13].

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obtained in studies by the United States Environmental Protection Agency, which reported that uranium
in drinking water contributed about 31% [U13, U14] of the total daily uranium intake.
27. In summary, uranium average levels in water worldwide are: 2 µg/L (15 mBq/L of 238U) for
groundwater (appendix A, table A1), 1 µg/L (12.4 mBq/L of 238U) for surface water (appendix A,
table A2), 1 µg/L (12.4 mBq/L of 238U) for public water supplies (appendix A, table A3), 0.5 µg/L
(6.5 mBq/L of 238U) for bottled mineral water (appendix A, table A4) (for natural uranium of
25,400 Bq/g). These median values hide great variability, notably for groundwater (0.0005–7.780 µg/L
(0.0063–96,472 mBq/L of 238U)). However, overall only a small proportion of few drinking water
samples (generally <3%) exceed the national or international guidelines. As expected, the values foruranium content in bottled mineral water are not so scattered.

4.

Levels in food

28. The measurement of the bioaccumulation of uranium in animals and plants shows thatconcentration factors are dependent on organism characteristics (e.g. species, life stage, physiology),
exposure pathways, and the chemical and physical characteristics of the environment [G4, Q2]. Various
publications have reported that the most available forms of uranium in plants were phosphate,
carbonate, sulphate or citrate forms [L14, L15].
29. Most available data relate to transfer through plants from their roots and the direct contamination of
aquatic organisms. Within plants, uranium concentrates mainly in the roots. Uranium found in meat anddairy products results from livestock feeding on plants and on food supplement made from natural
phosphates and supplied to dairy cows. Ingestion of soil particles, either directly or through consumptionof grass contaminated with soil, is likely to be a significant component of the total intake by livestock.
Transfer parameters of natural uranium are known for the main meat-producing species (cattle, sheep andpigs) and also for cow’s milk [I1]. They vary between 3.9 × 10−4 and 7.5 × 10−1 day/L in bovine meat andpoultry, respectively.
30. Uranium has been detected in a variety of foodstuffs, with great variability. The 238U activity has
been estimated to be 100-fold higher in root vegetables than in fruit or leafy vegetables as shown, forexample, by measurements for beets and tomatoes (100 vs. 1.13 mBq/kg, respectively) [I2]. A synopsis
of the activity of 238U measured by different authors in several types of foodstuffs is contained in theCommittee’s 1977 Report and in its 2000 Report [U7, U8]. Meat products have the lowest uranium
activity (between 0.08 and 20 of 238U mBq/kg). A recent report indicated activities of 238U between
1 and 49 mBq/kg for meat products [R11].
31. An estimate of daily uranium ingestion of food was made in Japan for urban residents [K30].
Concentration of 238U varied between 9.9 × 10−5 and 5.9 Bq/kg depending on food types: grain vinegar
and boiled and dried hijiki (a brown sea vegetable), respectively. When prepared diets were analysed,
the uranium concentrations observed were, on average, about four times higher than those seen in raw
foodstuffs. Hamilton explained this to the possible addition of uranium in seasonings and to transferfrom the cookware [H8]. It was unclear whether these dietary intakes included those from drinking
water and it was emphasized that the latter had sometimes been found to be equal to that from the diet
[H8]. Wrenn et al. have suggested that in regions where treated surface water was used for cooking anddrinking, food appeared to be the major source of uranium intake [W26].
32. In aquatic animals (crustaceans, molluscs and fish), the bioconcentration factor from water is very
low [I3]. Concentration factors for fish vary from 0.01 to 20. The values depend on the behaviour of theorganisms (pelagic species accumulate approximately 10 times less than benthic species) and the372

UNSCEAR 2016 REPORT

tissues considered (bone 200–8,000 and kidneys > liver and gills > muscles 1.5–24 > digestive system> gonads). A study was performed in Japan to determine uranium concentrations in marine organisms(soft tissues) [M15]. This study showed large differences in uranium levels depending on the marine
species, with a minimum value of 0.077 µg/kg (0.97 mBq/kg of 238U) measured in rockfish (kichiji) anda value of 5,040 µg/kg (63.8 Bq/kg of 238U) found in octopus. The values reported by Belles et al. also
indicated that fish and seafood showed the highest uranium concentrations (90 µg/kg, 1.1 Bq/kg of
238U), followed by dairy products (40 µg/kg, 0.5 Bq/kg of 238U) [B12].
33. The concentrations of uranium in fish are also dependent on uranium levels in water. The uranium
concentrations in the muscle (dry weight) of fish caught in a Canadian lake receiving effluents from auranium mill were 7–11 times higher than those from fish caught in uncontaminated lakes [S50].
Uranium mines and mills operating between the 1940s and the late 1970s have left behind legacy
contamination due to historic mining and milling practices and incomplete site remediation duringdecommissioning. In Beaverlodge Lake, Northern Saskatchewan, Canada, elevated concentrations ofuranium are still present with a mean concentration for the period 2013–2015 of approximately
135 µg/L (range 130–142) corresponding to 1.69 Bq/L of 238U (range 1.63–1.78) [C2, C3, C4].
34. Human health risk assessments for environmental contaminants take soil ingestion rates into
consideration. The value recommended by Richardson and Stantec Consulting Ltd. [R13] of 20–40 mg/d
for children is based on mechanistic assessments made by Wilson et al. [W21] and Ozkaynak et al. [O10].
Another report suggested assessing health risk for children with the value of 100 mg of soil per day [U15].
However, uranium bound to soils is not completely bioavailable. Reported bioaccessibility (in vitro
estimate of bioavailability) values are quite variable ranging from 2112% in the gastric phase and4817% in the gastric and intestinal phase [J4] to less than 5% in the gastro-intestinal phase [T16].
35. The urinary concentrations of several metals, including uranium, were found to be higher thaninternational reference values in a study of schoolchildren and working children in Lahore, Pakistan[S44]. The measured urinary concentrations of uranium corresponded well with uranium concentrations
in drinking water.
36. A study of 19 categories of food was performed by Fisenne et al. [F5, F6]. Potatoes, meat, fresh
fish and bakery products were found to contribute more than 70% of the average uranium intake(1.2 µg/d or 14.9 mBq/d of 238U). Dietary levels of uranium in the United Kingdom were reported in astudy of typical diets using both raw and prepared foodstuffs. Analysis of the raw foodstuffs indicatedthat 83% of the daily intake of uranium derived from starchy roots, vegetables and fruit, and cereals.
37. The Committee in its 1977 Report [U7] included a summary of 238U concentrations in foodstuffs
in France, Japan and the former Soviet Union along with the results given above for the United States
and the United Kingdom. In areas with typical uranium concentrations in soils, the daily dietary intake
fell within a relatively small range, ~0.9 to 1.5 μg natural uranium (11.5 to 19 mBq of 238U). This range
is consistent with values given in several publications calculated from the mass activity of naturaluranium (25,400 Bq/g): 1.32 µg/day corresponding to 16.4 mBq/day of 238U for typical diets of adults
in New York City, Chicago and San Francisco in the United States [W11], 1.14 µg/day (14.5 mBq/dayof 238U) [K30] and 1.46 µg (18.6 mBq/day of 238U) in different cities in Japan [N9]. In the UnitedStates, the average daily per capita intake of natural uranium in foodstuffs was estimated to range fromabout 1 to 33 µg (i.e. from about 12 mBq 238U/day to 405 mBq 238U/day) determined from excretionmeasurements [S26]. The values given by ATSDR indicated that the daily intake of uranium from foodsources ranged from 0.9 to 1.5 μg/day [A31, W11]. Similar values were given for European countries[W15]: intakes ranged from 6 to 22 mBq/day of 238U corresponding to 0.47 and 1.77 µg/day,
respectively. This daily intake contributed to a body burden of around 50 µg (0.62 Bq of 238U) in
humans [F6].

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 373

38. The Codex Alimentarius gives guideline levels applied to radionuclides contained in food,destined for human consumption and traded internationally, which has been contaminated following a
nuclear or radiological emergency [C29]. These guideline levels apply to food after reconstitution or as
prepared for consumption, i.e. not to dried or concentrated foods, and are based on an intervention
exemption level of 1 mSv in a year. A value of 100 Bq/kg is given for 235U. However, these guideline
levels exclude radionuclides of natural origin such as 238U.
39. In summary, uranium is present in a variety of foodstuffs, with great variability. Potatoes, meat,
fresh fish and bakery products were found to contribute more than 70% of the average uranium daily
intake. The total daily intake in food was found to be around 1.5 μg (18.6 mBq of 238U), about twicethat via drinking water, recognizing that levels in diet and drinking water can vary greatly.

5.

Levels in milk

40. Some publications report uranium levels in milk, notably from cattle. The uranium concentrations
measured in milk were in a wide range, from 0.012 to 0.41 µg/L (0.15–5.2 mBq/L of 238U), dependingon the species and the technical methodology (table 3). A mean value of 0.26 µg/L is given,corresponding to 3.3 mBq/L of 238U (range 0.001–1.20 µg/L; 0.012–15 mBq/L of 238U), if the highest
values of Santos et al. [S3] are excluded. This mean value is above the reference value of 1 mBq/L of238
U given in the Committee’s UNSCEAR 2000 Report [U8]. Furthermore, transfer coefficients ofuranium into milk are also available in the literature [A5, K1, T8, W9, W28] and presented for different
species in table 4.
Table 3. Uranium content (mass and activity) in milk (animals)
Values are expressed in kg of fresh matter, the numbers in italics correspond to calculated data obtained from the mass
activity of natural uranium of 25,400 Bq/g and from the relative proportion of 238U (12.4 mBq/μg) in natural uranium of 48.3%[U] in μg/L

238

U in mBq/L

Reference

0.14–0.24

1.74–2.98

[A14]

1.20

14.8

[S43]

0.10 (0.03–0.24)

1.24 (0.38–3.0)[F16, M34]

0.72±0.35

8.9±4.3[A5]

0.001–0.01

0.01–0.12[R11]

0.21±0.02

2.56±0.25[P19]

0.25±0.06

3.07±0.74[P20]

3.09 (0.18–9.6)

38 (2.2–118)

[S3]

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UNSCEAR 2016 REPORT

Table 4. Transfer coefficients of uranium into milk (animals)
n.i.: not indicatedSpecies

Transfer factor (d/L)

Range (d/L)

Reference

Cattle2.0 × 10−4

6.0 × 10−5–6.0 × 10−4

[T8]

Cow

2.9 × 10−3

5.0 × 10−4–6.1 × 10−3

[K1]

1.0 × 10−3

3.0 × 10 –3.0 × 10

Goat

1.4 × 10

−3

n.i.

[K1]

Camel

4.2 × 10−3

1.1 × 10−3–1.5 × 10−2

[A5]

Sheep and goat

−4

−3

[T8]

B. Artificial sources
41. The main application of uranium is for energy production and military use. Uranium-235 isnaturally a fissile isotope. Uranium is used primarily in most nuclear power plants. Its utilization inmost reactors requires enrichment of natural uranium containing 0.72% by weight of 235U to a 235U
content of 2–5%. Weapons use high enriched uranium with over 90% 235U. Some research reactors andnaval reactors also use high enriched uranium. Depleted uranium is used as a metal in kinetic energy
penetrators and tank armour.
42. The nuclear fuel cycle leading to the production of electricity from uranium in nuclear powerreactors includes mining, milling, conversion, enrichment, fabrication of nuclear fuel and reprocessing[G18]. Mining for the extraction of uranium involves both conventional open pit (where deposits areclose to the surface) and underground mining (used for deeper deposits). Currently, most uranium
mining worldwide uses the in situ leach mining process [S4]. The milling process produces a uraniumoxide concentrate, named Yellowcake, which contains more than 80% uranium. This uranium oxide isthen converted to uranium hexafluoride (UF6). It contains only natural uranium, which is enriched via
one of the two major types of enrichment technologies, gaseous diffusion or gas centrifuge. The
enriched uranium hexafluoride is then converted to uranium dioxide (UO2) powder and processed into
ceramic pellets. Finally, these pellets are inserted into tubes of corrosion-resistant metal alloy, called
fuel rods, which are grouped in fuel assemblies for the nuclear fuel core of a power reactor.
43. The uranium remaining after removal of the enriched fraction is DU, containing 0.3% 235U or less
[B22, B26]. This uranium has various civilian applications, such as in counterweights or ballast in
aircraft or counterweights for rudders and flaps [B22], for X-ray radiation shielding in medical
equipment and also for containers for the transport of radioactive material. Moreover, DU has also beenused in glassware, ceramics and dentistry.
44. In addition to exposure to natural uranium in the environment, anthropogenic activities have led toincreasing uranium exposure for humans. For instance, uranium was found to leach into water fromuranium-bearing glass items (maximum uranium in water, 30 µg/L (0.38 Bq/L of 238U) and from ceramicglazed items in which natural uranium is used as a colouring agent (300 µg/L; 3.7 Bq/L of 238U) [L6].
45. Uranium is present in water as a result of leaching from natural deposits and waste from themining of uranium and other minerals, releases from the nuclear fuel cycle and the combustion of coaland other fuel [D22, E6, S26, T1]. Phosphate fertilizers, which may contain uranium at concentrations
as high as 150 mg/kg, may also contribute to the uranium content of groundwater [S26].

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 375

46. Contamination of surface water and groundwater by effluents from uranium mining, milling, andproduction operations due to in situ leaching methods has been documented [A31, E1, H14, S50].
Table 5 shows some recent data. Except for some specific locations, i.e. at the pit, the values ofuranium concentrations in surface water or groundwater are usually below the WHO guideline of
30 µg/L (372 mBq/L of 238U).
47. Since the 1970s, DU is used for kinetic energy penetrators and tank armour, because of its
pyrophoricity [B22]. The military applications of DU led to the significant release of this radionuclide
into the environment during the conflicts in Iraq and Kuwait (321 tons of DU), in Bosnia-Herzegovina(3 tons of DU), and in the Kosovo (10 tons of DU). More details are given in reports by the United
Nations Environment Programme (UNEP) and the National Defence Research Institute (NDRI) [H13,U3, U4, U5, U6].
48. The use of reprocessed DU in mixed oxide (MOX) fuels (constituted by 8–9% of plutonium and~90% of DU) has been used as a recycling strategy by countries including France and Japan as an
option to reduce the necessity for storage of spent fuel [R6].
49. In a study in Tajikistan, uranium concentrations were shown to vary from more than 1,600 µg/L(>20 Bq/L of 238U) at the pit lake to 90 µg/L (1 mBq/L of 238U) in tube supplies and 6.3 µg/L
(80 mBq/L of 238U) in drinking water from the neighbouring village [S28]. Another study documented
uranium contamination of groundwater in Arizona, the United States after uranium mining [D13].
Approximately 20% of total uranium concentrations in the water samples exceeded the maximum
concentration level for drinking water of 0.37 Bq/L of 238U [U17]. In another study, the uranium
content was measured in drinking water samples from locations near the uranium mining site at
Jaduguda, India [P8] with uranium concentrations between 0.03±0.01 and 11.6±1.3 µg/L – values
below the WHO guidelines of 30 µg/L for uranium level in drinking water [W18].
50. Uranium conversion, uranium enrichment and fuel fabrication facilities are other steps in the
nuclear fuel cycle, which also release small amounts of uranium to the environment [A31]. Tracy andMeyerhof showed that concentrations of uranium in the air near a uranium refinery were 200 times
higher than background concentrations [T18]. For monitoring stations in Port Hope, Canada, where auranium refinery is operating, the annual average concentration varied between 0.001 and 0.0158 µg/m3
in 1988 and 1989. Uranium concentrations subsequently decreased and varied between 0.005 and
0.00028 µg/m3 in the early 2000s [C27]. In 2009, elevated levels of uranium were registered in storage
reservoirs of liquid radioactive waste at the Mayak facility, Russian Federation: the concentration was
370 mBq/L to 520 Bq/L for 234U, and 260 mBq/L to 520 Bq/L for 238U [T21].
51. Concentrations of uranium in surface waters downstream from currently operating uranium mines
and mills are relatively low and decrease with distance from the point of effluent discharge. For
example, between 2000 and 2012 in the vicinity of Canadian uranium mines and mills within 1 km
from the discharge points the mean concentration values ranged from 17 to 0.92 µg/L and decreased tomean values in the range of 1.44 to 0.096 µg/L at distances greater than approximately 10 km fromdischarge points [C28].
52. Environmental contamination by uranium caused by DU in ammunition used in military conflicts
was reported in several studies and in UNEP reports [U3, U4, U5, U6]. Uranium in agriculture soil in
Kosovo and Bosnia–Herzegovina averaged 1.8 and 3 mg/kg, while concentrations in public drinking
water averaged 0.5 μg/L (16.3 mBq/L of 238U) and 0.4 μg/L (5.1 mBq/L of 238U), respectively [C12].
The average uranium concentrations in soil and water were consistent with natural levels, although
localized areas of greater concentration were measured in the immediate surroundings of the DU

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UNSCEAR 2016 REPORT

penetrators [D4, E5, S2]. In the UNEP report on Kosovo, a great variability was observed in uranium
concentrations in water samples, with a range between 0.006 and 2.15 µg/L [U4].
53. Carvalho and Oliveira [C12] found high, localized contamination of soil with DU at Djakovica(4,662 Bq/kg of 238U; 376 mg/kg). The water samples collected from public water distribution networks
and river water ranged from 0.2 to 0.76 µg/L (2.5–9.7 mBq/L of 238U) and the air samples ranged from
0.8 to 7.2 µBq/m3. Consistent results from a number of studies indicate that environmental
contamination by DU has been very localized and confined to the areas of ammunition impact.
54. Concentrations of uranium were measured in surface and groundwater at the Semipalatinsk
nuclear test site, where more than 400 nuclear tests were conducted [L28]. The measurements showedthat 238U concentrations in well water within the study area were in the range of 74–213 mBq/L. The
results of these studies suggest that diverse human activities involving uranium (from extraction toapplication) have led to some localized increase of the concentration in the environment.

Table 5. Overview of uranium content in water close to nuclear fuel facilities worldwide
The numbers in italics correspond to calculated data obtained from the mass activity of natural uranium of 25,400 Bq/g and from the relative proportion of 238U (12.4 mBq/μg) in natural uranium (48.3%); * mean
value

Country

Location

Type of plants

Samplenumber238
U(mBq/L)

Total uranium(μg/L)

Reference

Lagoa Real

Mining and ore processing

26

5.45 (0.1–259)

67.6 (1.2–3 212)

[C13]

CanadaSaskatchewan, Beaverloke Lake

Mining

22

100.8

1 250

[Y1]

India

Jaduguda

Mining and ore processing

33

3.2 (0.03–11.6)

39.7 (0.37–144)

[P8]

India

Jharkhand, Narwapahar

Mining103

0.63 (0.10–3.75)

7.8 (1.24–46.5)

[R7]

India

Jharkhand, Bagjata

Mining10

3.22* (<0.5–11.2)

40* (<6.2–139)

[G12]

India

Jharkhand, Banduhurang

Mining10

2.15* (<0.5–27.5)

26.7* (<6.2–341)

[G12]

India

Jharkhand, Bagjata

Mining40(<61–55.9)(<12.6–693)[G13]

Kazakhstan andKyrgyzstan

Kurdai (Kazakhstan) and
Shekaftar, Kavak and Kadji-say (Kyrgyzstan)

Mining1028.2 (1.9–35.9)

350 (23.6–445)[U12](1 525 in lake at mining pit)

(18 910)

Kyrgyzstan

Mailuu Suu

Mining and milling

0.28 (0.27–0.34)

3.47 (3.35–4.2)

[C32]

25

(6 820 for tailings)
Nigeria

Jos plateau

Abandoned mining

5

0.10 (0.03–0.27)

1.24

[A22]

Nigeria

Abakaliki

Mining20

2 (0–7)24.6 (0–87)[O2]

Portugal

Viseu, Quinta do Bispo & Cunha Baixa

Mining1217.7 (1 km)

220 (1 km)[C11]

0.5 (7 km)

6.2 (7 km)

Tajikistan

Taboshar & Digmai

Mining and milling66.95 (3.4–92)

86.2 (42.2–1 138)

[S28]

USA

Karnes County, Texas Pana Maria

Mining and milling6

19.7 (14.8–95)

244 (183–1 178)

[M18]

USA

California Juniper

93.35 (0.02–52.37)

42.2 (0.25–649)[K8]

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 377

Brazil378

UNSCEAR 2016 REPORTIII. PHYSICAL, RADIOLOGICAL AND CHEMICAL CHARACTERISTICS
A. Physical and radiological characteristics
55. Uranium is an actinide and has one of the highest atomic numbers (92) of any naturally occurring
element. It is a silvery-white metal that is malleable, ductile, slightly paramagnetic, with a very highdensity. In its natural state, crustal uranium occurs as a component of several minerals, including
carnotite, uraninite (pitchblende) and brannerite, but is not found in the metallic state in nature. In
addition, uranium metal is pyrophoric. Due to its pyrophoricity, it is used in military applications,particularly in armor-piercing projectiles.
56. The three naturally occurring isotopes of uranium (234U, 235U and 238U) behave the same way
chemically but have different radiological properties (table 6). Uranium-238 has the longest half-lifeand consequently the lowest specific activity. It is the most abundant naturally occurring uranium
isotope. Among the natural isotopes of uranium, 234U has the highest specific activity and the shortest
half-life [L35, S15]. However, other isotopes of uranium may be produced, such as 233U that has a very
high specific activity (3.57 × 108 Bq/g) [L43].
57. Natural uranium is made of a mixture of the three isotopes described above with about 0.72% of
U in mass (table 7). Depleted uranium refers to isotope mixtures that contain a lower percentage of235
U (from about 0.2–0.3%) while enriched uranium contains typically 3–5% 235U in mass and may
contain up to 90% 235U for military applications (table 7) [L35, S15].
235

Table 6. Radiological properties of uranium isotopes [N8]nt: nuclear transformation

Half-life(years)Isotope234

Emitted energy (MeV/nt)

Daughter
nuclideAlpha

Electron

PhotonTotal

Th

4.84300.01370.00204.8587231

Th

4.46930.05300.16694.6891234

Th

4.25840.00920.00144.2691U

5

2.46 × 10

230

235

U

7.04 × 108

238

U

4.47 × 109

Table 7. Typical isotopic composition in mass and activity of different types of uranium [L35, S15]
Mass (%)
Type of uranium
NaturalDepleted
Enriched (3.5%

235

U)

238

U

235

U

Activity (%)
234

U

238

U

235

U

Activity (Bq) for 1 g
234

U

238

U

235

U

234

U

99.2840.711

0.0053

48.2

2.3

49.5

12 400580

12 47499.8070.0.1990.0008

86.1

1.1

12.8

12 400158

1 843

96.481

3.46

0.02831

14.7

3.4

81.8

12 0052 800

66 703

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 379

58. Uranium in rock and soil is in secular equilibrium with the daughters of the decay chain.
Uranium-238 decays to 234Th and 234mPa reaching secular equilibrium within about one year. The 238U
decay chain ends with the stable isotope 206Pb. However, disequilibrium between the uranium isotopes
can occur through physical and chemical changes involving water. For example, combinations ofphysical and chemical processes can lead to a separation of 238U and 234U in groundwater [N2].
Uranium-235 and 238U decays contribute to subsequent formation of 10 or more emitters of , and (figure I). Due to their short half-life, 234Th and 234mPa (24.1 days and 1.17 min, respectively) are
generally present together with 238U.

B. Chemical characteristics
59. Uranium has four valencies, which represent the number of valence bonds that uranium can form
with other atoms. The most common valencies of uranium in ores are IV and VI. The conditions oftransition from valency IV to VI depend on the redox potential of the medium. Compounds containing
hexavalent uranium are much more soluble than those containing tetravalent uranium. Hexavalenturanium forms complexes such as uranyl carbonates (UO2CO3) and uranyl sulphates (UO2SO4).
60. Uranium can take many other chemical forms. In nature, it is generally found as uranium dioxide(UO2) with other compounds, such as in pitchblende. Uranium dioxide (UO2) is the final product in the
manufacture of nuclear fuel pellets used in most reactors, and is also present as DU in MOX. Uranium
metal is generally alloyed with other elements (Si, Cr, Al, Fe, Mo, Sn, Al).
61. Uranium metal is pyrophoric and extremely reactive. It oxidizes readily to form triuranium
octaoxide (U3O8) and uranium dioxide (UO2). Uranium trioxide (UO3.xH2O) and uranium peroxide(UO4.2H2O) also exist. Triuranium octaoxide (U3O8) is the most stable oxide of uranium and is the form
most commonly found in nature. Both triuranium octaoxide (U3O8) and uranium dioxide (UO2) are solids
that have low solubility in water and are relatively stable over a wide range of environmental conditions.
Through reactions with acids, bases or chelating agents, compounds such as uranyl nitrate, uranyl
carbonate, uranyl chloride, uranyl sulphate and uranyl acetate may also be formed. Ammonium diuranate(NH4)2U2O7 is a basic product in the uranium fuel cycle, a component of Yellowcake, which is produced
during milling and consists of magnesia or ammonium diuranate (respectively MgU2O7 and (NH4)2U2O7).
These compounds are converted to uranium hexafluoride (UF6) prior to enrichment. During conversion,diuranate is converted to uranium trioxide (UO3), then UF4, and finally UF6 is then enriched from 235U
~0.7% to 235U at ~4%. This enriched UF6 is then converted into UO2 [D6].
62. As indicated above, different chemical forms of uranium are produced throughout the nuclear fuelcycle [G18]. Uranium-fluorine compounds are encountered in uranium processing, with uranium
hexafluoride (UF6) and uranium tetrafluoride (UF4) being the two most common. The compound UF6 isused in the enrichment process of uranium to increase the proportion of 235U, either by gaseous diffusion
or by gas centrifuge. It is prepared industrially by the reaction of UF4 powder with fluorine gas. Uranium
tetrafluoride is obtained by treating UO2 with gaseous fluorhydric acid. It is a non-hygroscopic, nonvolatile compound and very soluble in water. In the presence of water vapour, it undergoes pyrohydrolysis
and becomes UO2. When gaseous UF6 is released into air or as it enters the respiratory tract, it hydrolizes
with moisture in air to produce hydrofluoric gas and particulate UO2F2. The oxidation states andcrystallographic forms of uranium in DU particles have been determined from selected samples collected
at different sites in Kosovo and Kuwait contaminated by DU ammunition during conflicts [L39]. Oxidized
uranium (+6) was found in large, fragile and bright yellow DU particles released during a fire at DU
ammunition storage facilities in Kosovo and Kuwait and crystalline phases such as schoepite

380

UNSCEAR 2016 REPORT(UO3.2.25H2O), dehydrated schoepite (UO3.0.75H2O) and metaschoepite (UO3.2.0H2O) were identified[L39]. These DU particles were rapidly dissolved indicating a high degree of potential mobility andbioavailability. Crystalline phases such as UO2 and metallic uranium or U–Ti alloy were also determined
in impacted DU particles from Kosovo and Kuwait.

IV. HUMAN EXPOSURE

A. Exposure of members of the public
63. Natural uranium exposure of humans occurs through food, water and inhalation. Reviews of
autopsy data show that the skeleton is the main site of accumulation of uranium (~80% of total) [B53,
K6, S25, U18]. This result, observed in a United States population, was corroborated by measurements
made in the United Kingdom [H7, H8] and Japan [I18]. Publications differ about whether the
distribution of uranium in the skeleton appears to be uniform [H12] or not [S24]. The age dependenceof uranium concentration in the skeleton was also investigated in humans (from six months to 65 years)
from vertebrae bones collected in Canada from population exposed to high uranium levels in drinking
water [L11]. The data indicated higher uranium accumulation at six months.
64. Fisenne et al. [F6] summarized numerous publications from 12 countries concerning uranium
concentration in human tissues, including blood, soft tissue and bone. This analysis demonstrated small
differences in uranium concentration in soft tissue and bone. The authors calculated an average of 30 µg
for the skeletal burden. The body burden of uranium in humans was estimated between 50 and 60 µg, with
57% in the skeleton, 20% in muscles, 16% in fat, 4% in blood, 2% in the lungs, 1% in the liver and 0.36%
in kidneys as presented in figure II [F6]. One publication mentioned uranium levels in the human brain[K6] with values between 0.18 and 0.77 µg of uranium/kg (n=3), corresponding to 0.4 to 0.99%.
65. The Committee in its UNSCEAR 1977 Report [U7] provides further reference concentrations of
238
U in various human tissues expressed in mBq/kg: 20 mBq/kg in the lungs, 3 mBq/kg in the liver,
30 mBq/kg in the kidneys, 5 mBq/kg in muscles and 100 mBq/kg in the bones. Despite somedifferences, these values and those reported by [F6] are similar. High uranium concentrations were alsomeasured in the kidneys in humans [A2, D20, S25]. Calculations made for four age groups (infant, oneyear-old, ten-year-old, adult) indicated that long-term chronic uranium ingestion would result in akidney burden of 6.6% of daily uranium intake for all age groups [C21].
66. Fisenne and Welford [F4] measured an average kidney content of 0.13 g 238U from 12 New York
City dwellers, aged 20–60. An average daily New York City diet of 1.3 g yields a blood uptake of
0.026 μg. The kidney-to-blood uptake ratio was estimated at 4 and the bone-to-blood uptake ratio at
3.3. Figure III shows the average daily dietary intake of 238U measured in 16 cities in 12 countries [F7].

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 381

Figure II. Distribution of 238U in human body [F6]

Figure III.

Average daily 238U dietary intake measured in 12 countries [F7]

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UNSCEAR 2016 REPORT

B. Occupational exposure
67. Exposure to uranium may also be relevant to occupational exposure, notably for workers involvedin electricity production in nuclear power reactors or at any stage of the nuclear fuel cycle. Each stageof the nuclear fuel cycle is associated with distinct exposure characteristics [D6, D7]. During mining,
workers may be exposed to various uranium compounds exhibiting different solubility such as
Yellowcake. Further, workers in metal mines, such as underground gold mines, may also be exposed touranium.
68. Several tissues were collected at the autopsy of workers by the United States Transuranium andUranium Registries (USTUR) [F3]. Some cases have been analysed following uranium exposure [A34,
R31]. Accidental exposure to uranium hexafluoride (UF6) led to long-term (65 years) retention of
uranium in the deep lungs and in thoracic lymph nodes [A34]. High concentration of uranium intracheobronchial lymph node was also found in other cases without accidental exposure [R31].
69. Often, data on worker exposure are restricted to measurements of external radiation exposure. For
instance, Anderson et al. reported that only 16% of the Oak Ridge Gaseous Diffusion Plant (ORGDP)
workers were monitored between 1948 and 1988 for internal exposure to uranium by urinalysis [A12].
The difficulty is that monitoring programmes for internal exposure need a combination of bioassay
techniques, e.g. urine and faecal analysis, especially in workplaces where compounds of different
solubility are handled and also in cases of accidental intakes [J6].
70. Following inhalation of insoluble forms of uranium, the lungs may retain the highest
concentrations of uranium [A2, I19]. Adams et al. reported measurements of uranium in kidney andbone following lifetime occupational exposure to uranium aerosols [A2], showing greater retention inbone than kidneys, as seen in data for natural uranium in tissue samples from members of the public.

C. Measurement of uranium
71. The amount of uranium taken into the body can be assessed from external radiation measurements
or by bioassay sampling (urine or faeces). The choice and efficacy of each procedure is dictated by theroute of intake, the pattern of exposure, the physical and chemical form of the uranium, the time between
intake and measurement, and the detection limit of the analytical procedure used [A31, L9, S30].
72. Exposure to uranium can be assessed through the detection of uranium in the urine [B4, C18,
C31, D21, L18, L21, M62, S37, S39, W12, W23]. After absorption through oral, dermal and inhalationroutes, uranium is excreted in urine mostly as uranyl ions. Uranium urinalysis data have been shown tocorrelate with airborne uranium exposure when averaged over time and the contribution from ingested
uranium is insignificant. Thus, urinalysis can be used to verify the adequacy of air sampling and as a
non-invasive method for the estimation of exposure [C18, D9, T7].
73. In vivo external radiation measurements can be used to determine the amount of uranium in the
respiratory tract or whole body. The energy of the main gamma emission from 235U is 186 keV.
Specialized counting systems (e.g. using germanium semiconductor detectors) are required. The
minimum detectable activity (MDA) of uranium in the chest depends on the isotopic mix of 235U and
238
U and also on the total amount of uranium in the chest or in the whole body. However, it is of theorder of 4 Bq of 235U and 100 Bq of 238U [L25]. When 234Th has reached secular equilibrium with 238U,
photon emissions from 234Th may be used in addition to those from 235U to further reduce the detection

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 383

limit, by summing the 234Th and 235U photopeaks. Kramer et al. have reported an MDA of 4 mg (i.e.
49 Bq of 238U) [K13]. The detection limit is a function of the chest wall thickness of the measured
individual. This parameter must be measured in order to interpret in vivo measurements of uranium inthe respiratory tract.
74. Measurement of uranium excreted in urine after exposure is potentially a more sensitive methodthan chest monitoring to determine the amount of inhaled uranium. The limit of detection by alpha
spectrometry is approximately 0.1 mBq of 234U, 235U or 238U in a 24-hour urine sample. Counting times ofapproximately one week are required to achieve this sensitivity. For natural uranium, a measurement of
0.1 mBq of either 234U or 238U would correspond to about 8 ng of total uranium. Specialized mass
spectrometric techniques (such as multicollector inductively coupled plasma mass spectrometry; highresolution inductively coupled plasma mass spectrometry; or thermal-ionization mass spectrometry) can
provide isotopic analysis at levels lower than can be achieved by alpha spectrometry [K12, L20].
Synchrotron-based X-ray techniques (e.g. X-ray fluorescence microscopy, X-ray absorption fine structure,
X-ray diffraction may be also used for uranium measurements [C15, P5].
75. The use of faecal assay is confined to intakes by inhalation of relatively insoluble forms ofuranium, and dose assessments using these data are subject to substantial uncertainties [D8, J6].
Measurement of uranium in hair could be used as an indicator of body burden in contaminated subjects[B16, I23, J3].
76. Many studies have reported urinary excretion of uranium in humans. The National Report on
Human Exposure to Environmental Chemicals [C16] gives uranium concentrations in both µg/L and µg/g
creatinine. Expressed in µg/L, mean levels of uranium in the general United States population range from0.005–0.009 μg/L according to surveys conducted from 1999 to 2012 on 18,266 individuals. Oeh et al.
[O1] measured uranium content in 113 urine samples from 63 occupationally unexposed persons inGermany. The urinary excretion of uranium per day was in the range of 1.4 to 77.5 ng with a geometricmean of 13.9 ng and median of 14.4 ng. Höllriegl et al. [H21] measured the uranium content in urine in
the general public of Nigeria with creatinine normalized values from <10.4 to 150 ng/L (median
13.8 ng/L) and from 2.52 to 252.7 ng/g creatinine (median 33.4 ng/g). Malátová et al. [M8] measured
daily excretion of 238U in urine in the general population (mean 0.311 mBq, range 0.011–2.88 mBq). The
measured urinary excretion per day among 40 active uranium miners indicate a mean value of 0.56 mBq
with a range of 0.08–2.77 mBq of 238U normalized to 1.7 g daily creatinine excretion [M7]. A study
performed in Italy indicated that the daily excretion for the Italian volunteers ranges from 8.2 ng to 59 ng
uranium [B1]. The lowest daily excretion was observed for the youngest volunteer (seven years old).

V. BIOKINETICS
77. The main routes of intake of uranium into the human body are ingestion and inhalation. Transfer
through intact skin is a minor route. Wounds require consideration in occupational exposure. In general,
occupational exposure arises primarily through inhalation of dust containing uranium or following
injury. Exposure of the general public arises mainly through ingestion of water or foodstuffs containing
uranium. The extent of transfer of ingested or inhaled uranium to blood depends on its chemical form
[I10, N2].

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A. Inhalation
78. The intake of radionuclides is determined by the air concentration and by the respiratory
characteristics of the subjects, particularly the ventilation rate, which changes according to the level of
exercise and determines the volume of air inhaled and the deposition of inhaled radionuclides in the
airways of the lungs.
79. For radionuclides inhaled in particulate form, regional deposition in the respiratory tract isgoverned mainly by the size distribution of the aerosol particles [I9]. Deposition fractions of gases andvapours are determined by their chemical form. After deposition in the respiratory tract, absorption andtransport of radionuclides involve three general processes. Material deposited in the anterior nasal
passage is removed by extrinsic means such as nose blowing. In other regions, clearance is competitive
between upward particle transport out of the lungs and dissolution and absorption to blood from thelungs. Particles escalated out of the lungs are subsequently swallowed and pass through the alimentarytract where absorption can occur.
80. The ICRP human respiratory tract model describes the biokinetics and dosimetry of inhaledmaterial and is used to calculate the inhalation dose coefficients that are in general use for radiologicalprotection and scientific purposes [I8]. This model represents the deposition of inhaled radionuclides inthe different regions of the respiratory tract, and the clearance of the deposited activity by mechanicaltransport and absorption to blood. Mechanical particle transport rates are taken to be the same for allmaterial, but are altered by factors such as smoking and disease. Absorption into body fluids depends
on the physical and chemical form of the deposited material [I7]. Absorption is modelled as a two-stage
process: dissolution (dissociation of material into body fluids) and uptake of soluble material. Uptake
into blood is usually treated as instantaneous while dissolution is time dependant and modelled by threeparameters: a fraction (fr) of the activity is rapidly dissolved at a rate (sr), the remaining fraction (1 – fr)
is dissolved at a slower rate (ss).
81. The absorption rate of a given compound may vary greatly depending on its method of production
and history. The ICRP recommends that the absorption rate of any material should be determined fromthe study of the material itself. In the absence of data, ICRP recommends default parameters for threereference absorption types: Type F (fast), corresponding to rapid and complete absorption of theradionuclide with a half-time of about 10 min; Type M (moderate), corresponding to the absorption of
20% of the activity with a half-time of 10 min; Type S (slow), corresponding to the absorption of 1%
with a half-time of 10 min [P4].
82. Hodgson et al. [H20] and Ansoborlo et al. [A15] reviewed the absorption kinetics of uranium
compounds handled in the British and French nuclear industry. In vivo experiments in rats and in vitro
dissolution studies led to the classification of UO2 and U3O8 as Type S; mixed oxides, UF4, UO3 and(NH4)2U2O7 as Type M; and UO4, UO2(NO3)2 and UO2F2 as Type F. In addition, these studies provided
specific absorption parameter values for each of these compounds. Duport et al. [D32] studied thesolubility of radioactive dust present in the workplace atmosphere from three types of Canadianuranium ores—Yellowcake, UO2 and UO3—using simulated lung fluid and determined their solubility
classification. Solubility studies were conducted in Canada on material from various uranium mining
and production facilities [R17]. Bečková and Malátová [B10] studied the solubility of dust samples
from the underground uranium mine, Rožná, using simulated lung fluid and calculated specificabsorption parameter values for 238U, 234U and 230Th. The dissolution parameters calculated for UF6
mixture were higher than the current ICRP dose coefficient for Type F uranium (factor 2–7) [A34].
83. Models allow dose calculation from the inhalation of uranium particles in different chemical forms
and in several sizes. However, specific data might be needed for remediation and decommissioningANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 385

activities potentially generating uranium nanoparticles [T20]. A study of inhaled nanoparticles of uranium
in rats [P18] showed that 97% of inhaled particles were deposited in the deep lung and partly translocatedto the pulmonary interstitium. Approximately 22% of these deposited particles were rapidly cleared (lungretention half-life of 2.4 h) and for the 78% remaining, the lung retention half-time was estimated at 1,412
days. The ICRP is currently developing material specific absorption parameter values for compounds ofuranium and is revising the default parameter values for Type F, M and S compounds.

B. Ingestion
84. Material may reach the alimentary tract either directly by ingestion or indirectly by transfer fromthe respiratory tract or from the systemic circulation. Absorption takes place largely in the specialized
absorptive region of the small intestine [L2]. The extent of absorption of individual radionuclides isdependent on the chemical properties of the element, and also the specific form of the intake. It isquantified by the fraction of element reaching blood following entry in the alimentary tract. The
absorption and retention of radionuclides in the human alimentary tract are described in the human
alimentary tract model produced by ICRP [I14]. This model depicts transfer of ingested material
between alimentary tract regions, faecal elimination and absorption into blood [I14]. In the ICRP
model, the fraction f1 represents the small intestinal absorption, and the symbol fA refers to the total
absorption from the different sectors of the human alimentary tract, including the walls [I14].
85. Soluble uranium is absorbed throughout the small intestine [D25, K9]. Comparative data between
species (rabbit, rat, hamster, dog, baboon, pig and human) have been provided by several authors [F11,T19, W26]. Tracy et al. reported a gastrointestinal absorption factor of 0.06% in rats and rabbits [T19]
while Frelon et al. reported a value of 0.4% in rats [F11] for uranyl nitrate administrated in drinking water.
Leggett and Harrison [L23], Zamora et al. [Z5, Z7] and Wrenn et al. [W27] reviewed the uptake ofingested uranium from the alimentary tract in environmentally exposed human subjects and in volunteer
studies. The distribution of fA values was in a range of 0.001 to 0.063, with daily uranium intake varying
from 0.37 to 573 µg. The authors estimated that the best estimate for fA was 0.009, with no correlation
with age, sex, duration of exposure, and total uranium intake. These values are in accordance with those ofanother study that found fractional absorption (fA) values in a range from less than 0.1% to about 6% forindividual subjects, with the central values from the different studies falling in the range 1–2% and 4%
from water for both 234U and 238U [S30]. On the basis of available data, the ICRP Publication 69 [I10]takes the fractional absorption of uranium from diet to be 2% in adults.
86. In newborn infants, fractional absorption may increase by a factor of about two due to the higher
intestinal permeability [I14]. This higher absorption in newborns was measured in different species(rats, guinea pigs, pigs, dogs) [S45, S47]. On the basis of animal data, ICRP recommended an f1 valueof 0.04 for infants and 0.02 for anyone more than one year of age [I10]. Chen et al. measured uranium
concentrations in 73 bone ash samples of young children residing in a Canadian community known tohave an elevated level of uranium in its drinking water supply [C22] and estimated fractional
absorption with confidence intervals as wide as 0.0930.113 for infants and 0.0500.032 for children of
1–7 years. In another extended study by Chen et al. [C23], the absorption fractions were estimated to be
0.0300.022 for children and youths of 7–18, and 0.0210.015 for adults of 18–25 years of age.
87. Experimental studies have shown that fractional absorption depends strongly on the ingested
chemical forms. The ICRP is currently revising its Publication 69 [I10] and will adopt an fA value of
0.002 for relatively insoluble compounds (e.g. UO2, U3O8) and an fA value of 0.02 for all other moresoluble chemical forms [P4]. A biokinetic model was recently developed to describe uptake andretention in hair following ingestion [L34].

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C. Absorption
88. Few human data are available on uranium transfer through skin in a report of the National Council
on Radiation Protection and Measurements [N3] and in a review of the Armed Forces Radiobiology
Research Institute [M17]. Accidents involving workers with extensive skin exposure to uranium havebeen reported in the other reviews. Lu and Zhao indicated a rapid increase in the uranium level in urine
followed by severe kidney dysfunction [L48] with a return to normal values at post-accident day 30.
These results were similar to the clinical follow-up made on workers with acute uranium compound
intoxication [S16]. Acute renal failure occurred for several days post-accident with recovery one month
later. The result of a 33-year follow-up showed that kidney and liver functions were normal [S49].
89. In the use of DU munition, small particles originating from DU dust can contaminate open
wounds, and embedded DU fragments may be implanted in muscles. Uranium urine concentrations
following accidental intramuscular implantation of metal DU fragments were measured in United States
service members exposed to DU through incidents involving DU munition and vehicles protected by
DU armour [M25]. Uranium values were from 0.001 to 39.955 µg/g creatinine.
90. Several in vivo studies of rodents with uranium exposure of intact skin [D10, O4, T23] demonstratethat very soluble forms of uranium such as uranyl nitrate and ammonium uranyl tricarbonate are able todiffuse through the skin layer [L47]. The LD50 of uranium depended on the species as follows: rabbits>rats >guinea pigs >mice. This toxicity increased with the time and the area of exposure. In vitro Franz
diffusion chamber model [T23] and in vivo hairless rat model may be used for evaluation of uranium
passage through intact skin [P16, P17]. An in vivo study showed that a significant uptake of uranium froma uranyl nitrate solution could occur within the first six hours of exposure after skin contact. Furthermore,as high uranium concentration remained present at the deposit site for up to 24 hours after contamination,skin provided a reservoir for uranium that remained bioavailable [P17].
91. Percutaneous diffusion of uranium was also studied on damaged skin from hairless rats, following
an abrasion (stratum corneum removal) [P15, P17]. In vitro study of biopsies showed that thepercutaneous absorption of uranium increased with the impairment of the stratum corneum. Ex vivo
studies with biopsies showed an increase in the diffusion of uranium through skin after abrasioncompared with that through intact skin. These results have been confirmed in vivo in hairless rats [P16].
92. The NCRP’s biokinetic model describes the mobilization of radionuclides, including uranium,
entering the body through a wound to blood [N3]. Three wound retention categories were described,
corresponding to contamination with soluble forms, colloids, particles or fragments. The uranyl ion(UO22+) was classified as a weakly-retained radionuclide, and UO2 oxide particles behaved rather asstrongly-retained radionuclides.

D. Systemic distribution, retention and excretion
93. The translocation of uranium to blood strongly depends on the physical and chemical form of theinitial compound [A15]. After its absorption to blood, uranium is present mainly as uranyl ionscomplexed to proteins (e.g. transferrin, albumin) or bicarbonate anions [A16]. The main sites ofdeposition of uranium from the circulation are the skeleton, kidneys and general soft tissues. Humanand animal data show that urinary excretion is rapid with about two thirds of uranium reaching blood
excreted in the first 24 hours and a further 10% over the next 5 days. Most of the remaining uranium is
excreted over a period of a few months, but a small percentage of the amount injected may be retained
for a period of years [L22].

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 387

94. The work of Leggett [L22] was used by ICRP [I11] to propose a reference model for uranium
biokinetics. This model is constructed within a physiologically-based framework that is also applied tothe alkaline earth elements. Rates of uranium transfer between plasma, red blood cells, skeleton, liver,kidneys, other soft tissue and excretion pathways are based on measurement of uranium in humans andanimals and consideration of the physiological processes. The model considers age-related changes inorgan and tissue uptake and retention of uranium. The ICRP will use the same model in ongoing
revisions of dose coefficients [P4].
95. The United Kingdom Royal Society [R28] used the ICRP’s biokinetic models to estimate theconcentrations of uranium in the kidneys following chronic exposure over one year at a constant daily
uptake of 1 µg of uranium to blood. The estimated uranium concentration in the kidneys was
0.0056 µg/g kidney after one year and 0.011 µg/g kidney after 50 years of contamination. This result is
not in accordance with the result of other studies. For instance, a study measured 238U in diet andkidney tissue in residents of New York and showed that a daily intake of 1.27±0.03 µg resulted in auranium concentration in the kidneys of 0.00043±0.26 µg/g that was constant over ages <20 years to>60 years old [F4].
96. Several studies have found that uranium can be incorporated into brain tissues [B11, G10, L30,
O7]. It has been demonstrated by an in situ rat brain perfusion method that uranium is able to cross the
blood-brain barrier [L27] but the mechanism by which uranium is transferred to the brain is unknown.
After acute, subchronic or chronic exposure, low uranium concentrations are found in the various
structures of the brain [B6, F8, H24, L38, P11, P12]. The level of uranium in the brain, notably inolfactory bulbs, was greatest after inhalation [H26, T15]. Such uranium concentration in olfactory bulbs
as compared to other cerebral structures is attributable to the direct transfer of uranium via olfactory
receptor neurons [I6, T15].
97. Mean values of uranium concentrations in human breast milk were reported as 0.03 µg/L
(0.76 mBq/L) [W9] and 0.30 µg/L (7.6 mBq/L) [C1]. The daily intake of uranium of mothers is
0.030.019 µg/kg (meanSD) [W9]. These values are in accordance with values measured in several
animal studies, which suggest similar transfer processes (see also table 3).

E. Materno-fetal transfer
98. Few data are available on materno-fetal transfer in humans and in animals. Some publications
have reported uranium concentrations in fetal samples [B46, B47, L26, S17, W10], in placenta [B47,
S17, W10], in amniotic fluid [C14, S17] and in cord blood [G22].
99. Ham et al. [H7] reported uranium concentrations measured in human fetus and Bradley andProsser in placenta samples [B47]. The activities of 238U ranged from 0.1 to 9 mBq/kg in the fetal
samples of human fetus and from <1 to 11 mBq/kg in placenta samples. Uranium concentrations ranged
from 0 to 0.2 µg/L (mean=0.024 and median=0.005) in human amniotic fluid [C14] and from 0.003 to
0.834 µg/L (mean=0.104 and median=0.057) in cord blood [G22].
100. An experimental study performed by Legrand et al. on pregnant rats exposed to uranium via
drinking water (40 and 120 mg/L) indicated no elevated uranium concentration in exposed fetuses ascompared to control animals [L26]. Another study performed by Sikov and Mahlum on pregnant rats
after injection indicated that only a small fraction of the injected nuclide (148 kBq of 233U) entered thefetus and the distribution within the fetus was dissimilar from that observed in the dam [S17]. Only
0.06% of injected dose/g body weight was measured in fetal kidneys at 20 days as compared to 5.18 in

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UNSCEAR 2016 REPORT

the dam. The authors also reported 0.03% injected dose/g body weight in fetus, 0.05% in placenta and0.001% in the amniotic fluid [S17].
101. The ICRP [I12] has provided biokinetic and dosimetric models and dose coefficients for theembryo, fetus and newborn as a result of intake of radionuclides by the mother. The term fetus refers to
the developing human offspring after the eighth week of pregnancy. The equivalent dose to the embryo
is assumed to be the same as that to the uterus wall and proportional to the concentration of uranium inmaternal soft tissue. The ICRP uses a simple approach for the calculation of fetal doses for the majorityof elements and their radioisotopes, including uranium, considering data collected from studies ofanimals and humans [I12, I13]. Thus, fetal doses from uranium are calculated on the basis of relative
concentrations of uranium averaged over the whole body of the fetus (CF) and that of the mother (CM).
A conservative CF:CM ratio of 1 is used for intakes of uranium during pregnancy. The distribution of
uranium in the fetus is assumed to be 80% to skeleton, 2% to kidneys, and 18% to other tissues [I12].

VI. DOSIMETRY
102. Absorbed doses in tissues are calculated using dosimetric models such as those of ICRP (e.g. [I7,I10, I11]) and of the Medical Internal Radiation Dose (MIRD) committee [B39]. The absorbed dose is
the fundamental quantity that is estimated and averaged over particular tissues and organs. The
distribution of absorbed dose from internally deposited radionuclides, here the uranium isotopes,
depends on a number of factors, including the distribution of the radionuclides within organs andtissues, and the penetration and range of radiation emitted from the radionuclides. Dosimetric models
have been developed for this purpose.
103. Absorbed doses from photons and electrons are calculated by applying Monte-Carlo codes ofradiation transport to anthropomorphic computational phantoms representing a reference person [I15].
Alpha particles are considered to be absorbed in the region where they are emitted, except for theskeleton, lung, urinary bladder and alimentary tract where the respective positions of the alpha-emitting
radionuclides and of the sensitive target cells are taken into account to assess the dose absorbed by thetarget cells. In most organs and tissues, local activity and also the radiosensitive target cells are
assumed to be uniformly distributed. However, in these few specific tissues, the identification of theputative radiosensitive cells allows a more precise definition of the source and target geometry ofirradiation, which is of concern for the short-range alpha particles emitted by uranium isotopes. The
target cells identified in the thoracic airways include basal and secretory cells in the bronchial
epithelium, endothelial cells such as those of capillary walls, and type II epithelial cells in the alveolarinterstitial region [I8].
104. Following the ICRP, the equivalent dose in a region T, HT, is defined as:

w D
R

T ,R

R

where DT,R is the average absorbed dose in region T, due to radiation of type R; wR is the radiation
weighting factor for radiation R and is equal to 1 for photons and beta particles and 20 for alphaparticles [I10, I15].
The effective dose E is defined as a weighted average of equivalent doses to radiosensitive tissues ofthe body:

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 389

E   wT H T   wT  wR DT , R
tTR

where wT is the tissue weighting factor for tissue T, with the sum of wT being 1. The committed
effective dose ET(τ) is defined as the effective dose delivered over the time τ following intake of aradionuclide. The τ is usually set to 50 years for adults and for children up to the age of 70 years so as
to cover life-long irradiation [I10].
105. Doses from the inhalation or ingestion of unit mass of uranium can be determined by multiplying
the corresponding dose coefficient (i.e. for 238U, 235U and 234U) by the isotopic activity for each level ofenrichment [I7, I11].
106. The ICRP provides reference dose coefficients per intake of uranium isotopes for workers [I7]and members of the public [I10] in accordance with the international safety standards of the IAEA [I4].
The values for inhalation by workers and ingestion by members of the public are given in tables 8 and
9, respectively. The f1 values of the ICRP model depend on age, and the model for the systemic
behaviour of uranium is also age-dependent, and so the committed effective dose per unit intake
calculated for ingestion of soluble uranium by members of public exhibits age-dependence (table 9).
Table 8. Committed effective dose per intake (Sv/Bq) for inhalation by workers (for 5 μm AMAD
particulates) [I7]
Solubility group of uranium compound: F (fast soluble), M (moderately soluble) and S (slowly soluble)

Absorption type

Isotope
F
234

U

235
238

M
–7

S

6.4 × 10−6

2.1 × 10

6.8 × 10−6U

6.0 × 10−7

1.8 × 10−6

6.1 × 10−6U

5.8 × 10−7

1.6 × 10−6

5.7 × 10−6

Table 9. Committed effective dose per unit intake (Sv/Bq) for ingestion of soluble uranium bymembers of public [I10]Age at intake
Isotope3 months

1 year

5 years

10 years

15 years

−7

−7

−8

−8

7.4 × 10−8

7.5 × 10

5.0 × 10−8

8.5 × 10−8

7.1 × 10−8

7.0 × 10−8

4.7 × 10−8

8.0 × 10−8

6.8 × 10−8

6.7 × 10−8

4.5 × 10−8234

U

3.7 × 10

1.3 × 10

8.8 × 10235

U

3.5 × 10−7

1.3 × 10−7

238

U

3.3 × 10−7

1.2 × 10−7

Adult

390

UNSCEAR 2016 REPORTVII. BIOLOGICAL EFFECTS

A. Chemical versus radiological toxicity
107. Uranium represents a particularly difficult problem for internal emitter studies because of its
chemical and radiological toxicities. It is a radioactive heavy metal, and it is difficult to characterizedifferences in responses to the metal component alone, the radioactivity alone, or the possible combined
effects of both. Some studies on DU have attempted to define the metal component, however these arelimited by the presence of radioactivity.
108. Despite the dual toxicity of uranium, few studies investigate the respective parts of its chemical
and radiological toxicities. As with all chemicals, the chemical toxicity of uranium is linked to its
ability to interfere with compounds and biochemical processes in living organisms. The chemical actionof all isotopes and isotopic mixtures of uranium is independent from the specific activity because
chemical action depends only on chemical properties [W13]. However, it is dependent on its physicaland chemical forms. For instance, the NOAEL values measured for uranium effects vary depending on
the absorption type of uranium compound, i.e. F (fast soluble), M (moderately soluble) and S (slowly
soluble). Different reports gave NOAEL values as a function of the administration mode and theexposure duration [A31, W14].
109. Concerning the radiological hazard of uranium, alpha particles do not penetrate beyond the outerlayer of skin, except in regions of thinner skin (the depth varies typically in the range 20–100 µm). The
impact on health of alpha particles of uranium is expected mainly after internal contamination anddepends partly on the route of exposure (inhalation or ingestion) [W14].
110. Both natural uranium and DU pose primarily chemical rather than radiological hazards in the
short term. The toxicity of uranium depends on its chemical form and the route of exposure [A31]. The
potential health effects arising from uranium exposure are discussed below. Some effects are related tothe chemical toxicity of uranium, notably renal effects, while other effects are mainly due to theradiological toxicity of uranium, such as tumorigenic effects.
111. The relative importance of chemical and radiological toxicities of uranium thus depends on the
degree of enrichment of 235U (and 234U), the compound solubility, the chemical speciation and the modeof incorporation [A31, S46, T9]. Chemical toxicity from uranium exposure appears mainly in thekidneys and is assumed not to occur below a threshold concentration. The thresholds given in theliterature are most often derived from a NOAEL in animal experiments with the application of an
appropriate safety factor for transposition to humans [A31]. Human autopsy data are used to confirm
these observations [A31, S43].
112. Stradling et al. discussed anomalies between radiological and chemical limits for uranium after
inhalation by workers [S41]. As a consequence of the different procedures used in their calculation,
they are incompatible and adherence to one limit may result in a breach of the other. They concluded
that for chronic intake by members of the public, it can be deduced that a unified exposure level of
0.5 µg/kg per day or a daily intake of 35 µg would be acceptable in most cases. More recently, Thorne
and Wilson proposed higher limits corresponding to 2 µg/kg per day or a daily intake of 140 µg [T9].

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 391

B. Kidneys
113. The limited human studies suggest that damage to the kidneys can be detected following chronic
exposure that results in uranium concentrations as low as 0.1 µg per gram kidney [R28]. Human studies
also suggest that acute intake which leads to a peak uranium level of about 1 µg per gram kidney canlead to detectable kidney dysfunction [R28]. Some authors have aimed to predict renal concentrations
in human populations exposed to uranium for dose assessments [C21, S31]. In the epidemiologicalsection some human data are presented of populations exposed to high uranium concentrations indrinking water.

1.

Acute exposure

114. Morphological renal modifications induced by uranium (from 0.1 mg/kg after injection) werereported in rodents in several experimental studies, suggesting that the kidneys are the major target
organ of acute uranium toxicity independent from the route and duration of exposure [B25, D17, G27].
Histopathological changes, including degenerative changes or necrosis of the proximal tubularepithelium and glomeruli, were reported after acute exposure in rats [D12, M16]. Further, somehistological alterations were noted in renal tubules of rats following chronic exposure with elevatedconcentration of uranium [G10, O7]. Thus, acute exposure may lead to alteration of glomerules andtubules, whereas chronic exposure to uranium seems to affect only tubular functions. Following
intramuscular injection of DU, rats undergo dose-related tubular necrosis and glomerulonephritis. After30 days, glomerular damage is reversible in rats at all doses (0.1–1 mg/kg) but there is a dose dependent
delay in the initiation of the regeneration, seen first in the low-dose group (0.1 mg/kg) [Z17].
115. An experimental study performed by Shiquan et al. on rats considered a wide range in uranium
concentrations of 5, 2.5, 1.0, 0.5, 0.25, 0.1, 0.05, 0.025, 0.01, 0.0075, 0.005, 0.0025 and 0.001 mg/kg,
administered by intraperitoneal injection as uranyl nitrate [S14]. Doses from 0.01 to 0.05 mg/kg
induced slight renal damage. The lowest dose that induced renal damage in all rats of one group was
0.1 mg/kg. At the highest dose (5 mg/kg), necrosis of the proximal tubular epithelium was observed
after 6–12 hours, and half the rats died after 6–8 days. Primary damage to the kidneys was necrosis of
the proximal tubular epithelium. After repair with regeneration, different degrees of fibrous scarring
were found in the kidneys.
116. The third segment of the proximal convoluted tubule is the most affected site in the kidneys ofrats [G11] and humans [M9]. Experimental studies performed on dogs [M63] demonstrated that thecomplexed uranium (VI) is filtered from blood via the glomeruli. In the proximal tubule, water isreabsorbed and uranium is concentrated. As urinary flow acidifies, uranium (VI) complexes are
dissociated and uranium can bind epithelial membrane proteins. Uranium kidney retention thusincreases with urinary acidity and decreases with uranium complexation.
117. Table 10 summarizes studies of the toxic effects of uranium on the kidneys of rats following acuteexposure. Biological effects indicating toxicity of uranium to the kidney were noticed starting from
0.1 mg/kg [Z17] to 126 mg/kg of body weight [F18]. Most of the studies were done on rats and showeda decreased creatinine clearance and an increased electrolyte or protein excretion, indicating tubular
alterations of the kidneys [B5, F18] that can also be associated with liver alteration as shown by
increased transaminase level [G29].
118. Shude and Suquin et al. reported dose evaluation and medical follow-up of a case of acuteuranium compound intoxication [S16, S49]. A nuclear worker received both thermal and acid burns. A

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UNSCEAR 2016 REPORT

solution of uranyl nitrate hexahydrate was spurted on his body, and uranium was absorbed through theskin to blood. Acute renal failure and toxic hepatitis occurred during several days post-accident andrecovery occurred one month later. The amount of uranium taken into the body as a result of thisincident was calculated to be 116 mg [S16, S49]. The result of 33 years’ follow-up showed that thefunction of kidney and liver were normal. Chromosome aberrations of peripheral lymphocytes wereobserved, including ring, fragment and dicentric. The results of other examinations were at normal
levels, including peripheral blood, bone marrow, immune system, cardiovascular and respiratory
function, endocrine and metabolism [S49].
Table 10. Summary of studies of toxicity of acute exposure in kidneys of adult rats
ip: intraperitoneal; im: intramuscular; iv: intravenous; BUN: blood urea nitrogen; GFR: glomerular filtration rate; LDH: lactate
dehydrogenase; AST: aspartate aminotransferase; ALT: alkaline phosphatise; BW: body weight

UraniumcompoundUranylnitrateExposureconditions
ip 0.001–5 mg/kg

Postexposure
follow-up
8 daysSex
Biological effects
Male

Necrosis of the proximal tubularepitheliumStudyreferences[S14]

Damage repair with renal fibrous scar
Half died after 6–8 days in the highestdose group
Uranylacetate

ip 0.1–1 mg/kg

30 daysMale

Increased creatinine, BUN and albuminat high dose in serum[Z17]

Uranylnitrate

ip 0.5 mg/kg

5 daysMale

Increased creatinine, urea, cholesterol,phospholipids in plasma[B5]

Uranylfluoride

ip 0.66 mg/kg

110 days

Male

Decreased kidney weights[D12]

Uranylnitrateim 0.2–2 mg/kg

28 daysMale

Decreased plasma ALT, AST, protein
and increased urea, creatinine, ALP[F18]

Uranylnitrateip 10 mg/kg

28 daysMale

Increased concentrations of sodium
and protein in urine[H6]

Increased LDH, AST in plasma andprotein, albumin in urine

Decreased GFRUranylnitrateip 11.5 mg/kg

3 daysMale

Increased creatinine, urea, ALT, AST
plasma levels[G25]

Uranylnitrateim 7.9–126 mg/kg

28 daysMale

Decreased BW and died after 3–7 days

[F18]

Uranylnitrateip 5/10/20 mg/kg

2 daysMale

Histopathological lesions[A18]

Uranylnitrate

ip 50–500 μg/kg

21 daysFemaleRenal morphological alterations (at thefirst dose, becoming increasingly
prominent with higher doses)[B17]

Uranylnitrateiv 15/25 mg/kg

17 hours

FemaleChanges in endothelial cellmorphology for the higher dose

[A33]

Increased plasma creatinine and urea
plasma levels

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 393

119. Other high human exposure—accidental or deliberate—were reported. Roszell et al. reported
27 cases of human exposure to uranium and the resulting kidney effects [R24]. The uranium burden
estimated in the kidneys ranged between 10 µg/g kidney after accidental inhalation of UF4 powder[L48] and 100 µg/g kidney after deliberate ingestion of 15 g uranium acetate [P9]. Early symptoms ofrenal failure were noted in the first weeks after exposure, with renal dysfunction observed in somecases until 6 [P9] or 18 months [L48] after exposure. Kathren and Moore re-evaluated the intake anddeposition of soluble natural uranium compounds in three men accidentally exposed in an explosion in
1944. One of the three exposed individuals showed an altered clearance pattern for uranium shortly
after the accident, possibly from pulmonary oedema associated with concomitant exposure to acidfumes. However, medical examinations of two of the men 38 years after the accident revealed no
detectable deposition of uranium [K4].
120. Table 11 provides a summary of the effects of chronic uranium exposure on the kidneys ofexperimental animals. Conversely, chronic exposure did not clearly induce a toxic effect on the kidneysof mammals, with duration of exposure from three to twelve months and exposure level from 0.02 to
200 mg/kg. Tissue alteration of the tubules or glomerules was observed only for uranium doses above
400 mg/kg [G10, Z13]. After a nine-month chronic exposure to uranium via ingestion of uraniumcontaminated drinking water, the kidneys of rats did not show signs of histological lesions for uranium
renal levels >3 µg/g (3 µg/g for 120 mg/L [D29] and 6 µg/L for 600 mg/L [P23]).
121. Gilman et al. [G10] noted that effects on the kidneys could be seen at uranium levels of 0.06 mg ofuranium per kilogram of body weight per day for male rats. Nevertheless, no rise in histopathologicalseverity with increasing dose was reported: histological lesions starting from the lowest concentration(0.96 mg/L, 0.09 mg/kg body weight) were not significantly different from the kidney lesions observed forthe highest concentration (600 mg/L) [G10, G11]. No clear dose-dependent effect was observed after
chronic exposure. Indicators of kidney function after acute exposure (0.1–10 mg/kg) included theconcentrations of blood urea nitrogen (BUN), creatinine in blood or protein in urine [B5, F18, G23, H6,
O7]. Conversely, only limited changes were identified for chronic exposure to uranium in drinking water(2–16 mg/kg) [G11, O7] or from implanted DU pellets (200 mg/kg) [P12, Z13, Z14].
122. Various urinary parameters—including levels of urea, glucose, creatinine, total protein, and albumin
and also activities of LDH and N-acetyl-beta-D-glucosaminidase (NAG), and glucose excretion—had themost persistent effect after chronic exposure [G10]. Complementary studies showed that chronic low doseuranium exposure did not modify the nephrotoxic effects of gentamicin renal response as evidenced by therenal tissue levels of kidney injury molecule (KIM-1), osteopontin, and kallikrein [R27]. The results
observed with osteopontin were different from those found in clinical studies, which indicated a decreased
osteopontin level in urine [P30]. Gentamicin-induced increase in renal levels of KIM-1 was augmented inrats previously exposed to uranium as compared to uncontaminated animals.
123. In a dose-response study (0.27–40 mg/kg) of chronic oral exposure of rats, nephrotoxic andpro/anti-oxidant effects were analysed after three- and nine-month exposure [P23]. The uranium
content of the kidneys was proportional to uranium intake after three and 9 months of contamination. It
reached 6 µg/g of kidneys for the highest uranium exposed group, a nephrotoxic level for acute intake.
Uranium microdistribution analyses showed that it was found mainly in the nucleus of renal proximal
tubular cells and to a lesser extent in other renal structures. Nevertheless, no renal impairment wasobserved according to histological analysis or measurements of sensitive kidney biomarkers such as
KIM-1, 2-microglobuline or retinol binding protein. Uranium contamination appeared to reinforce theantioxidant system in the kidneys as the glutathione pool increased dose-dependently up to tenfold.

394

UNSCEAR 2016 REPORT

Table 11. Summary of studies of biological effect of chronic exposure to uranyl nitrate on kidneysof male adult animals
PCT: proximal convoluted tubule

Follow-up

Biological effects

Studyreferences

Species

Exposure conditions

Rat

Oral
100 mg/kg

27 days

Histopathological lesions (increased when
exposure duration increased)

[G15]

Rat

Oral
0.02–40 mg/kg

91 days

Decreased haemoglobin, erythrocytes,
glucose not correlated to the concentrationin drinking water[G11]

Histopathological lesions in the lowest group
Rat

Oral
2.7 mg/kg

275 days

Decreased plasma vitamin D and vitamin D
target genes in the kidney tissue

[T11]

Rat

Oral
2.7 mg/kg

275 days

Following single exposure toacetaminophen (paracetamol), increased
PCT necrosis of the kidney

[G26]

Rat

Oral
2–16 mg/kg

28 days

Increased glycaemia whatever the dose inplasma

[O7]

Mouse

Oral
13–26 mg/kg

122 days

Decreased urea and increased creatinine inplasma[T4]

RabbitOral
0.02–400 mg/kg

91 daysNo biochemical changes

[G10]

Rat

Muscle implantation
200–600 mg/kg

91 and 360 days

Increased plasma urea, creatinine andurinary beta-2 microglobulin and albumin

[Z14]

Rat

Oral
0.014-8 mg/kg

275 days

No biochemical changes

[D29]

Oral
0.27–40 mg/kg

91 and 275 days

Rat

Kidney histopathological lesions in thehighest group

No renal histopathological lesions
Glutathion increase dose-dependently andlipid peroxidation decrease in kidney[P23]

No nephrotoxity (biomarkers and histology)

124. A study by Silver et al. provided information regarding non-malignant chronic kidney disease[S19], which proved to be non-significant. There was a trend in chronic renal failure deaths in the
Colorado uranium miller cohort with duration of uranium milling employment (SMRs of 1.27, 1.33 and1.53 for 1–2, 3–9 and ≥10 years of employment, respectively) [P21]. However, when treated end-stage
renal disease for those receiving renal dialysis or kidney transplants, the incidence was evaluated using
the ESRD program (End Stage Renal Disease) date, there was no excess (SIR=0.71, 95% CI:
0.26-1.65). In the Fernald cohort, mortality from chronic renal disease was not related to uranium
exposure [S19] (see also table 18). Other studies of uranium worker cohorts indicated no significant
overall excesses of chronic kidney disease death [B38, C20, D33, M26, M27, P24]. Some renal effects
were studied in humans following chronic exposure to uranium via drinking water. These data areshown in appendix A, table A11, which summarizes the published literature on the health effects ofhuman exposure to uranium through ingestion of surface or groundwater.

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 395

2.

Influence of age

125. Few studies have investigated the influence of age at exposure on renal toxicity. The first results
considering this issue were published before 1920 for studies using dogs [M3]. The author
demonstrated that older animals developed more severe uranium poisoning than young animals,
associated with more marked histological alterations, leading to an impaired functional capacity [M3].
This result was corroborated by a more recent experiment, also performed on dogs [P10]. Uranium
effects on renal function (glomerular filtration rate) were more severe in older (3/4 weeks) than inyounger dogs (1/2 weeks). Recent publication of an experiment performed on rats with 0.1–2 mg/kg ofuranium acetate indicated a higher uranium concentration in the kidneys of neonatal animals than in
prepubertal or adult animals [H22].
126. Magdo et al. studied the effects of the high uranium concentration in drinking water of a private
wells used by a family (two adults and five children). The authors measured concentrations of up to(1,160 µg/L) in the groundwater [M5]. The authors evidenced a nephrotoxicity in the youngest family
member (a three-year-old child), demonstrating the highest sensitivity to uranium exposure. This caseshows potential for significant residential exposure to naturally occurring uranium in well water. It
highlights the special sensitivity of young children to residential environmental exposures, a reflectionof the large amount of time they spend in their homes, the developmental immaturity of their kidneys
and other organ systems, and the large volume of water they consume relative to body mass [M5].
However, this result appears inconsistent with observations made in young rodents.

3.

Biological mechanisms

127. According to Leggett [L21], uranium binds to the luminal brush-border membrane of tubules.
This binding decreases reabsorption of sodium and other compounds, resulting in an increased urinary
excretion of proteins, glucose, catalase, phosphate, citrate and sodium without causing cellular damage[B25]. Uranium can then separate from the luminal membranes by a number of mechanisms including
association with complexing ligands from the urinary tubular flow, detachment of uranium-boundmicrovillosities, and elimination of dead cells. The mechanism by which uranium then enters tubularcells has been investigated. Once in the cytoplasm, uranium accumulates in lysosomes, where it
precipitates with phosphate, forming microcrystals at high concentrations [M47]. This process induced
destruction of the lysosome.
128. The mechanisms of uranium effects on the kidneys have not been fully elucidated. The relations
between uranium penetration into and distribution within cells and its toxicity was analysed in kidney
proximal tubular cells. Some authors posited that uranium did not need to penetrate cells to exert its
toxic effect [L21, M68] while others posited that it did [L1, M47]. Rouas et al. have suggested that thephysical form of uranium (soluble or precipitate) and its intracellular localization play a role in cell
toxicity [R26]. They suggested that uranium may be visualized in the nuclei in kidney cultured cells
exposed to uranyl nitrate. Experimental studies have indicated that uranium exposure of animals (rats ormice) at 5–25 mg/kg may induce changes in solute transport [G11, T5], protein biosynthesis-related
genes [T4, T5] or cell signalling [P26, T5].
129. Several studies reported effects of uranium on renal transporters, Na+,K+-ATPase [B48], sugartransporters [G17, N4, R19], sodium-dependent phosphate cotransporters [M68, M69] and organic
cation transporters [M4, S13] but these effects were studied only in vitro using relatively high uranium
concentrations.

396

UNSCEAR 2016 REPORT130. Mitochondrial dysfunctions have recently been demonstrated in rats after injection at 0.5, 1 and
2 mg/kg per intraperitoneal injection [S9]. Isolated mitochondria from the uranyl acetate-treated rat kidney
showed a marked elevation in oxidative stress accompanied by mitochondrial membrane potential
collapse as compared to the control group. In addition, direct incubation of isolated kidney mitochondria
with uranyl acetate (50, 100 and 200 µM) suggested that uranyl acetate can disrupt the electron transfer
chain at complex II and III.
131. To investigate the influence of uranium speciation on its toxicity, cells representative of rat kidney
proximal epithelium (NRK-52E) have been exposed to uranyl-carbonate and -citrate complexes,
because they are the major complexes transiting through renal tubules after acute in vivo
contamination. When citrate is added to the exposure medium, the predominant species is uranium(VI)-bicarbonate. Nonetheless, citrate increases uranium (VI) toxicity and accelerates its intracellularaccumulation kinetics without inducing precipitation [C10].
132. Uranium can induce cell death but the exact mechanisms are still unclear. Some proposedmechanisms include apoptosis or genotoxicity [M54, T6, V5]. The mechanisms of acute toxicity ofuranium (from 50 to 500 µM) have been studied in renal cell lines and have shown a specific uranium
signature characterized by the downregulation of tubulin and actin [P27]. The most investigatedmechanism to explain uranium toxicity is the oxidative stress response, investigated both in vitro on
cell cultures and in vivo following acute or chronic exposure in rats and mice [B5, S9, T4, T5, T6]. The
results of these studies suggested uranium-induced oxidative stress imbalance with an increased
reactive oxygen species production associated with a depletion of endogenous cellular antioxidants forelevated concentrations administrated to cells (>400 µM) or to animals (>0.5 mg/kg).
133. Several studies have focused on the interaction between iron metabolism and uranium, in
conjunction with the affinity of these two elements for some proteins such as transferrin, ceruloplasmin
and ferritin [V8]. These studies revealed changes in gene expression and protein carriers of iron, DMT1
(Divalent Metal Transporter Type 1) and Fpn (ferroportin), in the liver and kidneys [B19]. A similar
study [D19] performed on rats chronically exposed to uranium in drinking water (2.7 mg/kg/day)showed the appearance of iron granules (aggregates) in the kidneys, indicating that chronic
contamination by uranium could cause long-term changes in the regulation of iron metabolism. A
recent study using surface plasmon resonance techniques has shown a strong affinity of fetuin-A
protein for uranyl ions even though this protein is present in a very small amount [B7].

4.

Conclusion

134. The kidneys are known as the most sensitive target organ after acute exposure, on the basis ofwell-documented studies. Studies on rodents (mice and rats) indicated that injury to the kidneys occurs
from 0.1 mg/kg whole body and renal concentrations of >3 µg/g, targeting the third segment of theproximal convulated tubule. Conversely, data on chronic exposure to low uranium doses are more
recent and have not shown clear nephrotoxic effect at 2–20 mg/kg and no specific biomarkers havebeen identified to date. New blood plasma or urinary biomarkers of the renal function or integrity havebeen investigated in experimental or clinical studies, with a view to allowing a precise diagnosis of thekidney function [G27]. Mitochondrial dysfunction seems to be involved in toxic mechanisms ofuranium, but the mechanisms of its toxicity are not fully elucidated to date, especially after chronic
exposure to low and high doses. Further experimental studies are necessary for a better comprehensionof the renal alteration process and the identification of more specific biomarkers of kidney alteration.

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 397

135. Chemical effects on the kidneys are usually assumed not to occur below certain thresholdconcentrations of uranium; most often, these findings have been derived from animal experiments withthe appropriate safety factors applied to human exposure [A31]. Maximum concentrations over 3 μg of
uranium per gram of kidney have been used as the basis for occupational exposure limits (e.g. [H20,
L21]). Leggett [L21] suggested that the occupational limit based on 3 μg of uranium per gram of kidney
is about tenfold too high for non-occupational exposure. Indeed, some human studies suggest thatdamage to the kidneys can be detected following chronic exposure that results in uranium
concentrations as low as 0.1 µg per gram kidney [R28]. Thorne and Wilson [T9] suggested 30 µg U/kg
kidney as a level below which effects will not be observed. Further, following a review of the existing
literature, Leggett et al. have recommended that “the concentration of uranium in the kidney should notexceed 1 μg U/g kidney at any time” [L25].

C. Bone1.

Acute exposure136. Several animal experiments demonstrated a high uranium accumulation in bones of rats and dogs
[A23, A24, P31]. Several studies performed on rats have investigated the effects of uranium on bone
physiology. These studies found an impairment of bone growth and bone formation [P33, U1, U2].
These effects were associated with inhibition of endochondral ossification in mice [B45] and in rats
[D15]. Other rat experiments indicated that uranium induced ultra-structural alterations in osteoblasts[T3]. An acute high dose of uranyl nitrate also delayed both tooth eruption and development in rats
[P33]. However, the retardation observed seven days after acute uranyl nitrate exposure was reversed
completely after 27 days [P34].
137. The in vivo results were confirmed by in vitro studies that showed that human osteoblasts weresensitive to uranium effects (increased reactive oxygen species production, decreased alkaline
phosphatase activity, modified osteoblast phenotypes, genomic instability) [M38, M39, M42, T3]. An
increase in oxidant stress shown by an increased lipid peroxidation was also observed in vivo in rat
bone at high doses [G8].

2.

Chronic exposure138. On the basis of measured tissue concentrations and organ weights [F4, F8, L22, W26], bone tissue
may contain up to 66–75% of the body burden of uranium following chronic exposure to uranium.
139. Several rat studies were designed to determine the uranium content of bones following chronic
contamination [A24, P3, R18]. Rodrigues et al. showed that accumulation of uranium in the rat skeleton
following intake via food increased to reach a plateau after one month [R18]. A long-term study ofPaquet et al. demonstrated that uranium accumulation in bones increased until 18 months [P3].
140. Some experiments investigated the long-term accumulation in bones of young male rats (between
the weaning and the post-puberty periods) [A23, R18]. Experiments performed on females during thegrowing period indicate that concentration of uranium in the animals’ femora increased faster in theearly stages of the animal life, then saturating in adult animals.

398

UNSCEAR 2016 REPORT141. Contamination of growing rats with uranium in food led to accumulation in bones (0.1–1.1 µg
U/g bone) that exhibited the same pattern as the skeleton growing curve. Despite this accumulation,
there was no change in the bone mineral density (BMD) [R18]. However, uranium exposure of adults
led to a decrease in the BMD.
142. In contrast, a study of chronic exposure of growing rats to natural uranium for nine months via
daily oral ingestion (uranium-contaminated drinking water at 40 mg/L) affected the skeleton by
decreasing messenger RNA (mRNA) expression of genes involved in bone metabolism and decreasing
femoral cortical bone area, while no changes were observed in adult rats [W1]. Chronic contamination
by subcutaneous implantation of powdered uranium dioxide (125 mg/kg) in rats resulted in an
inhibition of bone formation as has also been described for acute poisoning with uranium [D15].
143. Bone metabolism was investigated in human populations receiving high levels of uranium via
drinking water by measurements of biochemical indicators of bone formation (osteocalcin and amino
terminal propeptide of type I procollagen) and a marker for bone resorption (serum type I collagencarboxy-terminal telopeptide (CTx)) [K27]. The authors showed an elevation of CTx and osteocalcin
that could be associated with increased uranium exposure.

3.

In vitro studies: mechanisms

144. The cytotoxic effect of uranium on rat and human osteoblasts is highly dependent on its
speciation. Exposure to non-toxic doses or non-toxic species of uranium induces the activation of twomarkers of bone formation and mineralization (osteocalcin and bone sialoprotein), while their inhibition
is observed after toxic exposure [M37, M38]. This study highlights the importance of a controlled
speciation of uranium in toxicological studies.
145. In vitro transcriptomic studies performed on several human cell lines taken from kidneys or lungs
as representative targets have highlighted the involvement of osteopontin in the uranyl toxicity
mechanisms [P29]. This major non-collagenous protein involved in the organo-mineral homeostasis of
the bone presents a specific composition in acidic clusters associated with numerous phosphorylations,and also a relative plasticity. Qi et al. showed with in vitro models that native phosphorylated
osteopontin binds uranyl with a nanomolar affinity and that this binding induces conformationalchanges enabling the formation of a very stable complex of uranium [Q1].

4.

Conclusion

146. In conclusion, acute and chronic exposure to uranium-induced biological effects on bonemetabolism such as the impairment of bone growth and of bone formation and the inhibition of
endochondral ossification or the delay of development. Most effects were observed in experimental
models (rodents or isolated humans cells). One study performed on humans exposed to high levels of
uranium in drinking water showed an elevation of type 1 collagen carboxy-terminal telopeptide (CTx),a plasma marker for bone resorption, and of osteocalcin, indicator of bone formation [K27]. All thesestudies—in animals and humans—suggest that uranium affects bone turnover. Osteoblasts appeared tobe the main cell targets of uranium. Some effects were also observed in humans, indicating that bone
may be a target of uranium chemical toxicity in humans.

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 399

D. Lungs
1.

Acute exposure147. In general, the more soluble compounds (uranyl fluoride, uranium tetrachloride, uranyl nitrate) areless toxic to the lungs but more toxic to distal organs due to easier absorption of the uranium from thelungs into the blood and systemic transport [G1].
148. The behaviour of uranium in the lungs following inhalation has been studied in animal
experiments from various industrial settings [A31, C30, C31, D3, E3, S35, S38, S39, S40, W19, W20].
These studies demonstrated that the behaviour of uranium, its distribution, and its clearance in lungs isdependent on the solubility of uranium compounds. Results of these experiments were used to confirm
or improve the ICRP models as described in chapter V on biokinetics.
149. Concerning the clinical effects of uranium in the respiratory system, authors reported alveolar
fibrosis in rats [M60], congestion and haemorrhage in rats and guinea pigs [L19] and bronchopneumonia
in rats and rabbits [D36] for uranium levels of between 10 and 100 mg/m3. The comparison of repeated
and acute uranium exposure via inhalation was performed in rats [M55] with aerosols varying from 116 to
375 mg/m3. The results showed that UO2 repeated pre-exposure by inhalation increased the genotoxic
effects of UO4 inhalation, when UO4 exposure alone had no effect. However, it is not clear if these effects
were due to a potentiation of the effect of UO4 by pre-treatment with UO2 or to a cumulative effect of thetwo types of exposure.

2.

Chronic exposure150. Experimental studies performed on dogs, monkeys and rats with uranium show that inhalation ofnatural uranium dioxide (UO2) at a mean concentration of 5 mg U/m3 for periods as long as five years led
to pulmonary neoplasia and fibrosis [L18]. Pulmonary neoplasia developed in a high percentage of thedogs examined two–six years after exposure. Pulmonary and tracheobronchial lymph node fibrosis,
consistent with radiation effects, was dose dependent and more marked in monkeys than in dogs.
151. Nose only inhalation in rats showed that chronic inhalation of natural uranium ore dust (without
significant radon content) created a risk of primary malignant and non-malignant lung tumour
formation [M49]. The frequency of primary malignant lung tumours was 0.016, 0.175 and 0.328 andprimary non-malignant lung tumours 0.016, 0.135 and 0.131 in the control, low (19 mg/m3 leading toan absorbed dose of 0.87 Gy) and high (50 mg/m3 leading to an absorbed dose of 1.64 Gy) aerosol
exposed groups, respectively, without difference in tumour latency between the groups. Despite lymph
node specific burdens ranging from 1 to 60-fold greater than the specific lung burden in the same
animal, no lymph node tumours were observed.

3.

In vitro studies: mechanisms

152. Inhalation of soluble uranyl nitrate led to uranium uptake in the lysosomes of alveolar
macrophages and precipitation in the form of insoluble phosphate [B20]. A study by Lizon and Fritschon alveolar macrophages with uranium concentrations from 5 × 10−5 to 10−3 M showed that the toxicity
of uranium was concomitant with the presence of insoluble forms in the culture medium [L40].

400

UNSCEAR 2016 REPORT153. Orona and Tasat analysed rat alveolar macrophages to better understand the pathological effects
associated with DU inhalation, metabolic activity, phagocytosis, genotoxicity and inflammation [O6].
The effects of 12.5–200 μM DU seemed to be dose-dependent, most observed from 100 µM. At low
doses, DU induced phagocytosis and at high doses, provoked the secretion of TNFα. Apoptosis was
induced through the whole range of doses tested. The uranium-induced TNFα secretion bymacrophages was consistent with results from previous studies [G5, Z11]. Lung fibrosis was correlatedwith abnormal expression of TNFα and IL−6, which could be antagonized by antibodies against TNFα[Z11, Z12].
154. Orona and Tasat suggested that at low doses (12.5 µM), DU induced O2−, which may act as theprincipal mediator of DNA damage, while at higher doses (200 μM), the signalling pathway mediatedby O2− may be blocked, and the prevailing DNA damage would be by TNFα [O6]. A study by Monleau
et al. indicated that exposure to DU by inhalation resulted in DNA strand breaks in broncho-alveolar
lavage cells and in an increase in inflammatory cytokine expression and production of hydroperoxides
in lung tissue [M56].
155. In addition to effects on pulmonary macrophages, a study showed that uranium (from 0.25 to1 mM) induced significant oxidative stress in rat lung epithelial cells followed by a concomitantdecrease in the antioxidant potential (glutathione and superoxide dismutase) of the cells [P13]. Further,some publications indicated that not only soluble (uranium acetate) but also particulate (uranium
trioxide) uranium induced concentration-dependent cytotoxicity in human epithelial cells [W22].
156. Depleted uranium was clastogenic and induced chromosomal aberrations after 48 hours [L4]. Xie
et al. [X1] found a loss of contact inhibition of these cells after particulate DU exposure, withchromosome instability and a change of cell phenotype, suggesting a neoplastic process.

4.

Conclusion

157. Experimental acute and chronic studies demonstrated possible effects of uranium on fibrosis andtumour formation in lungs. Data obtained in vitro indicated the induction of genotoxic lesions (DNAstrand breaks) and activation of inflammatory pathways on alveolar macrophages with high uranium
concentrations. However, these underlying mechanisms were not investigated in vivo, limiting therelevance of these molecular effects to understand the link between uranium exposure and observed
pathologies (fibrosis and tumour formation).

E. Liver158. Accumulation of uranium was observed in the liver of rats after injection, implantation or oral
administration (uranyl nitrate) but at a lower concentration than in the kidneys for the same level of
exposure [C35, P3, P11]. Acute exposure led to decreased liver weight and increased plasmatransaminase levels, both indicators of liver hepatotoxicity in rats [G24, M57, O7] and in mice [O11].
159. More recently, some results were published on hepatic effects of uranium following in vivo
exposure of rats [D29, G29], which demonstrated that some enzymes of xenobiotic metabolism, notably
the cytochrome P450 of type 3A (CYP3A), were modified by chronic contamination with DU. The
time-course study performed at 1, 3, 6, 9 and 18 months after exposure indicated that significant
changes were observed at six and nine months [G29], with a 50% decrease in the mRNA level ofANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 401

CYP2C11 at six months and an increase in gene expression of CYP3A at nine months. Concerning thedose–response study, the most substantial effects were observed in the liver of rats after nine months of
exposure to 120 mg/L: CYP3A gene and protein expression and enzyme activity all decreased by morethan 40% [D29, G29].
160. Several studies investigated the functional consequences of a co-exposure to uranium (uranyl
nitrate) and the drugs chlorzoxazone [C26, M57], ipriflavone [C25], theophylline [Y4] or paracetamol[G27, R25] in rats. Altered drug pharmacokinetics was observed with high dose and chronic low dose
exposure to uranium, which could be due to liver dysfunction. Functional toxicity of uranium was also
estimated by measuring xenobiotic detoxification enzymes and their gene expression levels, protein
levels or enzyme activities. Some studies reported altered levels of xenobiotic detoxification enzymes[G27, P6] while others reported altered pharmacokinetic metabolism of certain drugs after acute [C25,
C26, M57] and chronic exposure to uranium [G27]. Oxidative stress may also occur in hepatocytes
exposed to uranium, as demonstrated by mitochondrial or lysosomal alterations [P26] ormetallothionein involvement [M47].
161. Chronic contamination with DU (uranyl nitrate, 1 mg/rat/day for 9 months) affected cholesterol
metabolism in the rat liver [R2]. Relative mRNA levels of the enzyme cholesterol storage weremodified, and also the proteins involved in the transport and the regulation of cholesterol homeostasis.
One accident involving workers with extensive skin exposure to uranium has been reported in theliterature [L48, S16, S49]. These results were similar to observations from the clinical follow-up of
workers with acute uranium compound intoxication. Toxic hepatitis occurred for several days postaccident with recovery one month later [S16]. The result of 33-year follow-up showed that liverfunctions were normal [S49]. In conclusion, experimental studies performed in rats indicated thaturanium exposure induced some biological effects on the liver, without a high uranium accumulation inthis organ. These biological effects did not lead to the appearance of pathologies in animals. However,in humans, one study reported a transient toxic hepatitis.

F. Brain1.

Behavioural effects

162. Adult rats exposed to uranium showed subtle but significant behavioural changes. Increases inlocomotor activity, in line crossing and in rearing behaviour were observed after exposure of rats to DU
in drinking water at 2 or 4 mg/kg per day or after inhalation [B52, M54]. Females seemed to be moreresistant: unlike males, they did not show locomotor symptoms [B51]. Exposure to uranium to 0.1, 0.3
or 1 mg uranium/kg also affected working memory with poorer performance (decreased latency) in alight–dark passive avoidance response system [B6]. The spatial working memory measured by thepercentage of spontaneous alternation was significantly lower after exposure to DU by inhalation andafter ingestion of 1 mg per day of 4% enriched uranium [H24, H25, M54]. Lastly, exposure at this doseof enriched uranium had a deleterious effect on anxiety and increased the amount of rapid eye
movement in sleep [H24, H25, H26]. However, DU had no significant effect in the same experimental
conditions [H24, H26, L30]. A recent review summarized the different effects of uranium on behaviour.
It can lead to neurobehavioural impairments, including increased locomotor activity, perturbation of thesleep-wake cycle, decreased memory, and increased anxiety. The mechanisms underlying theseneurobehavioural disturbances are not clearly understood [D16].

402

UNSCEAR 2016 REPORT163. Mouse experiments were conducted on ApoE−/− mice that had been genetically knocked out for
the apolipoprotein E gene, the product of which regulated cholesterol metabolism. These mice had
hypercholesterolemia and expressed biochemical markers of Alzheimer’s disease. Administration of
DU to these mice resulted in impaired memory compared to unexposed ApoE−/− mice [L33]. This
cognitive effect was associated with a trend toward higher total cholesterol content in the cerebral
cortex (+15%). This study demonstrated that some pathological conditions may increase sensitivity touranium. Table 12 summarizes the main studies on animal behaviour after uranium exposure.
Table 12. Summary of studies of uranium exposure on rodent behaviour
im: intramuscular; DU: depleted uranium; EU: enriched uranium

UraniumcompoundSpecies

Exposure conditions

Main biological effects

Reference

UranylacetateAdult rat

2 and 4 mg/kg/day (water) DU
during 2 weeks and 6 months

Increased locomotor activity[B52]

UranylnitrateAdult rat

2.5, 5 and 10 mg/kg/day (water)
DU during 3 months

Increased locomotor activity,
decreased spatial memory[B11]

UranyldioxideAdult rat

Inhalation DU 30 min at
197 mg/m3, 4 days a week for
3 weeks

Increased locomotor activity anddecreased spatial memory one
day post-exposure[M54]

UranylacetateAdult rat

im. 0.1, 0.3 and 1 mg/kg DU

Decreased locomotor activity,
decreased grip strength,
decreased working memory at 6,
13, 20 and 27 days post-dosing[B6]

UranylnitrateAdult rat

1 mg/kg/day (water) 4% enricheduranium (EU) during 1.5 months

Decreased spatial memory,increased of anxiety[H25]

UranylnitrateAdult rat

1 mg/kg/day (water) 4% EU during
1.5 months

Increased paradoxical sleep[L29]

UranylnitrateAdult rati.p. 144 μg /kg

No change in sleep-wake cycle[L30]

UranylnitrateAdult rat

1 mg/kg/day (water) 4% EU during
3 or 9 months

Decreased spatial memory[H24]

Uranylacetate

Fetal
mouse

1, 2 and 4 mg/kg/day DU during
gestation and lactation

Accelerated appearance righting
reflex, forelimb placing, grasping,swimming and weight gain

[B50]

Uranylacetate

Young
mouse

1, 2 and 4 mg/kg/day (water) DU
during 21 days since the birth

Decreased locomotor activity,
decreased working memory[B51]

UranylacetateFetal rat

10, 20 and 40 mg/kg/day (water)
DU during 3 months (male)

Decreased learning in pups

[A10]

UranylnitrateAdult rat

1 mg/kg/day (water) DU during
2 months since the birth

Decreased spatial memory[B15]

Uranylnitrate

Adult
mouse

4 mg/kg/day (water) DU during 3.5
monthsImpaired memory[L33]

1 and 3 days

(ApoE−/−)

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 403

2.

Neurotransmitters164. Several experimental studies suggest that uranium can induce changes in neurotransmitter levels:
acetylcholine levels were unchanged in the hippocampus after exposure to 1 mg DU per kg and day but
were decreased in the cortex [B14]. Chronic DU contamination induced a fall in the rate ofacetylcholine (ACh) and AChE activity in the entorhinal cortex and cerebellum [B15, B55]. This
disturbance of cholinergic function was associated with a decrease in gene expression of several
proteins [B14]. These studies suggest that the modifications of neurobehavioural tasks following
uranium exposure could be linked with a change of AChE activity in the cerebral cortex [B52].
165. A study performed in Sprague-Dawley rats (following 1.5, 6, or 9 months with 2.7 mg U/kg perday) reported that AChE activity was not significantly affected in the striatum, hippocampus, or frontal
cortex at any time point, but it was significantly decreased in the cerebellum at six months [B55].
Depleted uranium exposure at 2.7 mg/kg per day also induced a significant decrease in AChE activity
in the striatum and cerebral cortex [B14, B55]. A dose–response study performed at nine months
following chronic contamination indicated that uranium effects (15% decrease in AChE activity) wereindependent of uranium content in drinking water (from 0.2 to 120 mg/L).
166. A 1.5-month ingestion of 1 mg U/kg per day increased the dopamine level in the hypothalamus[B55]. Chronic exposure produced a significant decrease in the serotonin (5HIAA) level and theserotoninergic (5HT-ergic) turnover ratio in the frontal cortex and also a decrease in the dopamine(DOPAC) level and dopaminergic (DA-ergic) turnover ratio in the striatum [B55]. It appeared that
disruption of these systems differed depending on the brain structure considered, the time of exposureand the degree of uranium enrichment [A1, B14, B55, L29].

3.

Oxidative stress167. One specific mechanism by which uranium leads to neuro-effects could be oxidative stress. The
behavioural changes correlated with lipid peroxidation in the brain induced by uranium [B52, G8]. The
gene expression or enzymatic activity of the main antioxidant enzymes, i.e. superoxide dismutase,catalase and glutathione peroxidase, increased significantly in the hippocampus and the cerebral cortex
after exposure to DU and decreased after exposure to 1 mg of 4% enriched uranium per kg per day[L31, L38]. The cell response to DU could be interpreted as a defensive mechanism towards free
radical damage to cerebral tissue (increase of several antioxidant agents in order to counteract theoxidative stress). The oxidative stress induced by the enriched uranium is possibly too high to be
counteracted by the cell defences.
168. A dose–response study performed nine months after chronic contamination with DU from 0.2 to
120 mg/L indicated that uranium affected the activity of these enzymes differently (diminution for SOD
and increase for GPx or Catalase). The gene expression of inducible nitric oxide synthase, the enzymethat synthesizes nitric oxide, increased significantly after chronic exposure to 1 mg DU per kg and day[L38]. Repeated administration of DU as uranyl acetate during seven days also increased the nitrite
levels in the brains of rats [A1].

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4.

Exposure of developing animals

169. In rodents, exposures at 1, 2 or 4 mg DU/kg per day during development accelerated theappearance of several types of behaviour: righting reflex (for example, when the rat is put on its back, itturns over immediately), placing reactions (for example, the rat is held by the tail over a table until its
whiskers get near, when it puts its paws on the table), grasping (the rat is picked up and the palm istouched with a wire and the response is to grip the wire), swimming and weight gain [B50].
170. Animals exposed to uranium at 1, 2 or 4 mg U/kg per day during development had a decreased
locomotor activity [B51] and their performance was worse on a test of working memory [B15]. The
spatial learning of the offspring of uranium-exposed male rats at 1, 2 or 4 mg U/kg per day for learning
was also affected [A10]. A dose–response study (0, 10 or 40 mg/L of uranium in drinking water)
indicated no significant uranium effect on behaviour at 10 mg /L, and an impairment of object
recognition memory (−20%) at 40 mg/L [L32].
171. Neurogenesis processes during pre- and postnatal brain development were studied in rats byinvestigating the structural morphology of brain, cell death and cell proliferation after chronic exposure
to drinking water contaminated with uranium (40 and 120 mg/L) [L26]. Major changes were observed
at 120 mg/L, both during prenatal and postnatal periods. At the highest dose, DU caused opposite
effects during brain development on cell proliferation and cell death processes, mainly between prenatal
and postnatal development. These modifications did not have a major impact on brain morphology but
they could affect the next steps of neurogenesis and disrupt the organization of the neuronal network.
172. Some studies were published on the comparative effects of depleted and enriched uranium in rats
[H24, L31]. Chronic exposure to 4% enriched uranium for 1.5 months through drinking water increased
the amount of paradoxical sleep, reduced the spatial working memory capacity and increased anxiety
while no effect was observed following exposure to DU [H24]. These cognitive effects were associatedwith imbalance of oxidative stress [L31]. Indeed, lipid peroxidation was increased in brain after
enriched uranium exposure but not after DU exposure. Enriched uranium induced a decrease of antioxidative enzymes, and DU induced an increase of these anti-oxidative enzymes.

5.

Conclusion

173. Animal studies suggest that uranium can have some negative effects on the behaviour of mature
animals that could be explained at a neurochemical level (neurotransmitters, oxidative stress) for
highest doses. In addition, some results obtained in rats and mice demonstrate differential effects
between depleted and enriched uranium, suggesting the importance of the radiological toxicity ofuranium. The data also suggest that the developing animal may acquire a specific sensitivity to uranium
effects. In humans (workers or public), the correlation between behavioural symptoms and exposure touranium was not demonstrated. Thus, data obtained in animals are only suggestive until they can be tied
to meaningful human research.

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 405

G. Reproduction and development
1.

Female reproductive function

174. Results for measurements of uranium concentrations in gonads were contradictory depending on
the study (species, administration pathway, uranium dose and exposure time). In fact, high uranium
accumulation is found in some fish and birds. For instance, significant uranium accumulation wasmeasured in female fish gonads (Danio rerio), corresponding to >20% of the relative burden [S20].
175. A dose-dependence of uranium concentration in ovaries was measured in rats and their offspring
after chronic oral exposure via food [H9, H10]. The accumulation of uranium was higher in the second
exposed generation. Other experimental studies performed on mice failed to report uranium
accumulation in the ovary after contamination by nitrate uranyl with doses up to 400 mg/L for uranium
content in drinking water [A21, F2, R9].
176. Some studies found that uranium affected oocyte quality in vivo with a 50% reduction in the
proportion of healthy oocytes from 20 mg/L [F2] and germinal vesicle oocytes cultured in vitro in thepresence of 1 mM uranyl acetate and observed for 72 hours [A21]. In vivo, these morphological effects
were observed from uranium content in drinking water above 5 mg/L [F2]. A study with similar
approaches (uranium ingestion via drinking water in mice) indicated similar results, with increased
dysmorphism of oocytes in contaminated groups (from 2.5, 5 and 10 mg U/kg per day chronicallyadministered in drinking water for 40 days) in a dose independent manner [K24]. In addition, a study
on mice contaminated in utero by uranium levels in drinking water from 0.5 to 60 µg/L, showed anincrease in uterine weight and a decrease in primary follicles [R9].
177. Reproductive effects following chronic oral uranium exposure were observed in rats exposed
during the first and the second generation [H10]. No effect were observed in F0 rats, but pregnancy rate,normal labour rate, and survival rate were decreased in offspring [H10]. In vitro organ culture system
was used to investigate the effects of uranium on human gonads during the first trimester of gestation(7–12 weeks) [A13]. Uranium at 0.05 mM increased the apoptosis rate, decreasing the germ cell densityof human fetal ovaries. The authors also demonstrated that human fetal germ cells were more sensitive
to uranium than mouse germ cells.

2.

Male reproductive function

178. Some animal studies indicated accumulation of uranium in testes with 0, 5, 10, and 25 mg/kg/dayof uranyl acetate dehydrate before mating and up to 21 days post birth [P7]. A linear dose-dependence
was found in testes of Japanese quail, with a ratio of accumulation similar in testes and in kidney [K25].
A dose-dependence of uranium concentration in testes was measured in rats and their offspring after
chronic oral exposure via food [H10]. The accumulation of uranium was higher in the second
generation.
179. Concerning the effects of DU on sex hormone levels, experimental studies performed in male rats
showed differing results. One experiment performed in male rats indicated an increase in testosterone
and luteinizing hormone levels and a decrease in follicle stimulating hormone level after a four-monthcontaminated food ingestion [H10], while a nine-month contaminated drinking water ingestion did notlead to changes in testosterone and 17β-estradiol levels [G20]. The different uranium doses 4 and
40 mg/kg per day in the studies of Hao et al. [H10], and 2.7 mg/kg per day in the study of Grignard etal. [G20] were used to explain the different levels. However, changes in testosterone levels were not

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UNSCEAR 2016 REPORT

observed in depleted-uranium Gulf War veterans [M19, M20]. A nine-month chronic oral exposure toenriched uranium produced a significant increase in the blood levels of testosterone at 40 mg/L in
drinking water, while no effect was observed with DU [G20].
180. Some reports on rodents highlighted a negative impact of uranium on male reproductive function,
including a decrease in male fertility and in the spermatid number per testis with a few
histopathological effects on the seminiferous tubules and interstitium after chronic exposure [L37,
L41]. Although some abnormal morphological forms and sperm parameters measured were affected byuranium exposure, these changes were independent of the uranium dose levels from 10 to 40 mg/kg per
day corresponding to ~200 to 800 mg/L in drinking water, respectively [L37]. Further, a doseindependent decrease in the pregnancy rate was observed in untreated females mated with male mice
exposed to between 10 and 80 mg/kg per day of uranium [L41].
181. Implantation of DU pellets did not change the concentration, motion and velocity of sperm, andthere is no evidence of detrimental effects of uranium on mating success, suggesting that implantationof up to 20 DU pellets of 1×2 mm (760 mg) in rats did not have an adverse impact on male
reproductive success, sperm concentration, or sperm velocity [A19]. Testicular histopathologicalabnormalities with deformations of seminiferous tubules (marked reduction in the seminiferous tubule
diameter) were observed in mice after high acute exposure [J1]. A seven-day daily intraperitoneal
administration of uranyl nitrate (0.5 mM/kg) induced a marked reduction in the seminiferous tubule
diameter and gametogenic count, with signs of testicular necrosis and exfoliation of germ cells,
including karyolysis and karyorrhexis figures.
182. The contribution of new in vitro models, such as organotypic culture systems, helps theunderstanding of the underlying action mechanism of chemicals. This approach was used as a
toxicological test to evaluate the effects of various compounds, including uranium, on gametogenesis
and steroidogenesis in rat, mouse and human testes [H1]. Some effects on germ cell development(reduction of the number of germ cells) or Leydig cell function were observed for uranium
concentrations above 5 × 10−5 M in human testis cells and above 5 × 10−4 M in rat testis cells [A13].
183. Uranyl fluoride injected in vivo into mouse testes led to an increase in the frequency ofchromosomal aberrations in spermatogonia and primary spermatocytes [H27]. These results wereconfirmed in another study that indicated the highest effects when doses of UO2F2 increased up to
6 mg/kg at 12 days post-exposure [H28]. A study performed with enriched uranium demonstrated
chromosome aberrations in spermatogonia [Z15]. Chromosome fragmentation, translocation,
polyvalence in primary spermatocyte and DNA strand breakage were observed in sperm. Effects ofuranium on human male reproduction studied in Gulf War I veterans did not evidence deleterious
effects on sperm quality, including volume, concentration, total sperm count, and functional parametersof sperm motility [M20, M21, M22].

3.

Effects on embryos and development

184. The effects of uranium on embryotoxicity were studied in rats and mice after acute exposure[D17]. Subcutaneous injections of uranyl acetate dihydrate (0.5, 1 and 2 mg/kg/day) in mice from day
6 to day 15 of gestation induced various effects [B44]. Although it was not dose-related, embryotoxicity also occurred in all uranium-treated groups (significant increases in the number of non-viable
implantations and in the percentage of postimplantation loss). Both the maternal NOAEL and the
NOAEL for embryotoxicity of uranyl acetate dihydrate were below 0.5 mg/kg/day, whereas the
NOAEL for teratogenicity was equal to 0.5 mg/kg/day.

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 407

185. Few animal studies investigated the effects of uranium exposure on development and results wereconflicting, depending on the quantity of administrated uranium and the exposure duration.
Subcutaneous injections of uranyl acetate dihydrate from day 6 to day 15 of gestation induced
malformations detected at 1 and 2 mg/kg/day in mice [B44], while no effect was observed at
0.5 mg/kg/day.
186. Paternain et al. demonstrated that embryo lethality could be observed in mice contaminated at
25 mg/kg/day [P7]. Significant increases in the number of dead young per litter were seen at birth andat day 4 of lactation in the 25 mg/kg/day group. The growth of the offspring was always significantly
lower for the uranium-treated animals. Since no effect was observed at lower doses (5 and
10 mg/kg/day), the results suggest that uranium does not cause adverse effects on fertility, general
reproductive parameters, or offspring survival at the concentrations usually ingested by man.
187. In vitro studies were also conducted on one-cell mouse embryos in culture medium with uranyl
nitrate at concentrations of 26, 52, 104 and 208 µg/mL [K23]. The results obtained showed thatconcentrations from 26 µg U/mL induced the delay of embryo development and the impairment of
blastomere proliferation. A study of acute toxicity was performed in mice receiving 4 mg/kg of DU per
intraperitoneal injections (i.p.) at day 11 of gestation and observations were made 4 days later.
Paradoxically, the authors found an increase in the fetus length and weight [M46].
188. Subcutaneous injections of uranyl acetate (0.415 and 0.830 mg/kg/day) were given to pregnant
rats on days 6 to 15 of gestation [A9]. Maternal toxicity and embryotoxicity were noted at the higher
dose, while fetotoxicity was evidenced at both doses. The fetotoxicity was evidenced by significant
reductions in fetal body weight and increases in the total number of skeletal abnormalities [A9].
189. Some studies indicated developmental toxicity of uranium, including teratogenicity, following anacute subcutaneous administration of 1 or 2 mg/kg/day [B44, D18]. A study by Zhu et al. [Z15] wasperformed with enriched uranium injected in rat testes, causing skeletal abnormalities in fetal rats witha positive correlation to the injected dose.
190. Hereditary effects of uranium were investigated in rats following the implantation of uranium
pellets in muscle (up to 12 DU pellets corresponding to 360 mg) [A20]. This study indicated nochanges in sperm motility and ribcage malformations, suggesting that uranium was not a significant
reproductive or developmental hazard. This study is in accordance with another study that indicatedthat uranium did not cause any adverse effects on fertility, general reproductive parameters, or offspringsurvival at the concentrations usually ingested by humans [P7]. One study performed on miceinvestigated the transmission of genetic damage to offspring of fathers contaminated with uranium via
depleted-uranium-implanted pellets [M45]. The authors demonstrated a dose-dependent increase in
mutation frequency in the offspring. Congenital malformations [A6, A7, S48] or birth defects [A3, B54,
F1] were reported in some human populations, but these effects were not correlated to a quantificationof uranium exposure.

4.

Conclusion

191. Some publications focused on reproduction and development issues after exposure to uranium inanimals (rats and mice). These studies indicated that both male and female reproductive functions(quality of oocytes and sperm parameters, embryo viability, and the development processes) may beaffected by acute and chronic exposure. However, these effects were observed for exposure levels
greater than either typical occupational or environmental levels of exposure.

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UNSCEAR 2016 REPORT

H. Other organs
1.

Skin

192. A study of acute exposure of rabbits was conducted using different administrations of uraniumoxide: subcutaneous deposit of ~30 mg U3O8; cutaneous deposit of 70 mg of U3O8 or 5 µg of uranium
acetate (1.9 kBq of 233U) [W3]. Uranyl nitrate, in ethereal or aqueous solution, produced a superficial
coagulation necrosis within a few hours [O4]. A similar but delayed effect was seen seven–nine days
following the application of powdered hygroscopic uranium pentachloride to the skin. Uranium
tetrachloride in suspension in lanolin caused a moderate local erythema of the skin at the site ofapplication, which disappeared in one–two days.
193. In vitro studies on primary cultures of rat skin keratinocytes and fibroblasts [P15] showed agreater decrease in proliferation rate associated with a greater mortality rate in rat skin keratinocytes
than in rat skin fibroblasts after uranium exposure. This can be explained by the three times higher
ability of keratinocytes to incorporate uranium compared to fibroblasts. This greater capacity of
epidermal cells than dermal cells to incorporate uranium was confirmed in vivo in hairless rats
following topical contamination with uranyl nitrate.
194. The consequences of protracted exposure to uranium were investigated using guinea pigs andrabbits [O4]. Guinea pigs that were repeatedly exposed to uranyl nitrate exhibited a superficial
coagulation necrosis and an inflammation of the epidermis. Their skin was in a constant state of
encrustation and desquamation accompanied by rapid regeneration from beneath. Rabbits that wererepeatedly exposed at multiple dermal sites exhibited an effect similar to that seen after single acuteexposure, but those rabbits that were repeatedly exposed at the same site developed severe dermal
ulcers after five to ten applications of the compound. Another in vivo study in rats revealed that longterm exposure of the dermis to uranium (U3O8 at 0.012 g/d for 30 daily topical applications) led to an
epidermal atrophy which, in turn, resulted in an increased permeability of the skin [U2].

2.

Endocrine system and metabolism

195. Vitamin D is essential for the homeostasis of calcium and phosphorus in the body. A few studies
have determined the effects of acute or chronic uranium contamination on the metabolism of vitamin D
in rats. Contamination by DU via drinking water at 40 mg/L [T11] induced a decrease in the bloodlevels of vitamin D in rats following a nine-month chronic exposure. Moreover, uranium targeted key
transcription factors (PPARα, PPARγ, HNF-1α, HNF-4α, LXRr, RXRα, and VDR) involved in thismetabolism [T11, T12]. However, these molecular changes did not lead to the emergence of diseaseassociated with vitamin D metabolism.
196. Chronic contamination by DU (uranyl nitrate, 2.7 mg/kg/day for 9 months) affected cholesterol
metabolism in the liver and brain [R1, R2], mRNA levels of the enzymes involved in the cholesterol
storage were modified in the liver and brain, and mRNA levels of enzyme involved in the cholesterolsynthesis were modified in the brain. Uranium also affected the proteins involved in the transport ofcholesterol and the regulation of cholesterol homeostasis. Thus, a chronic ingestion of uranium

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 409(40 mg/L in drinking water) caused subtle molecular effects on metabolism in the liver and brain ofrats. However, overall cholesterol levels were unaltered in this study using 40 mg/L uranium indrinking water.
197. Studies of rats exposed to uranium showed both disruptive effects on the reproductive system andestrogenic effects. Chronic contamination with DU (560 Bq/L) in drinking water produced no change inthe blood levels of the two principal steroid hormones—oestradiol and 17-testosterone (synthesized by
the testis), whereas contamination with enriched uranium (1,680 Bq/L; 40 mg/L) produced a significant
increase in the blood levels of testosterone [G20]. Consistent with the absence of hormonal changes
with DU, chronic contamination with DU did not induce a change of gene expression. However, theexpression of enzymes involved in the metabolism of hormones was amplified following a nine-month
exposure to enriched uranium. In addition, enriched uranium increased the gene expression oftranscription factors (RXR, LXR, FXR, SHP, SF-1, DAX-1) that positively regulate steroid
metabolism. These results show a differential effect of depleted and enriched uranium contamination on
testicular steroidogenesis [G20].
198. The human body is frequently exposed to potentially toxic compounds and is able to metabolize
them in order to protect itself. Xenobiotic-metabolizing enzymes, including cytochrome P450
(CYP450), play a central role [G25]. The kidneys and liver can metabolize many drugs or otherxenobiotics. Disturbances of this system have been demonstrated in vitro [M43] and in rats exposed tonephrotoxic [M57] or non-nephrotoxic uranium concentrations [P6, S29]. These studies showed that theexpression and activity of CYP450 can be modified by uranium exposure. The consequences of suchmodifications on xenobiotic metabolism were investigated using acetaminophen [G26]. Rats
contaminated with DU presented slower plasma acetaminophen elimination and also more marked renal
histological changes and an increase in blood markers of liver damage (at 40 mg/L during 9 months).
However, only slight effects of uranium were reported on enzymes of xenobiotic metabolism when
acetaminophen was administrated to rats as a single therapeutic treatment [R25].
199. In conclusion, studies performed on animal models (rat and mouse) chronically exposed toradionuclides showed that the chronic ingestion of uranium resulted in subtle biological effects on
various metabolic systems. These modifications did not lead to the appearance of pathologies, even for
uranium levels in drinking water up to 40 mg/L. The observed biological effects probably resulted froman adaptive response to the internal contamination.

3.

Immune and haematopoietic systems

200. Very few studies have investigated the effects of uranium exposure on the haematological system.
Reduced erythropoiesis might be expected owing to accumulation of uranium in bone in proximity tobone marrow and renal damage might result in a decrease in the number of red blood cells. A study
using rats found that chronic ingestion of uranium at 40 mg/L in drinking water for nine months led tokidney deterioration which may have been responsible for an observed decrease in the red blood cellcount; there was an associated modification of spleen erythropoiesis and levels of molecules involvedin erythrocyte degradation [B19].
201. A study by Giglio et al. was performed to assess the effect of uranium (uranyl nitrate) on the rateof erythropoiesis in rats [G9]. The authors showed that a single injection of uranium at 1 mg/kg induced
a transient depression of the red cell volume between 7 and 14 days. These effects were associated witha decreased Epo production and direct or indirect damage of erythroid progenitor cells.

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UNSCEAR 2016 REPORT

202. An experimental study using mice fed uranium-contaminated food [H11] compared the effect ofconcentrations of DU of 0, 3, 30 and 300 mg/kg feed for a duration of four months. The mostsignificant effects were observed in the 300 mg/kg group while the effects were either minor orindiscernible in the other groups. At high dose, the authors observed decreased immune function,
manifesting as decreased secretion of inflammatory mediators in the peritoneal macrophages, and also
reduced cytotoxicity of the splenic natural killer cells. Moreover, the cellular and humoral immune
functions were abnormal (decreased proliferation of the splenic T cells, proportion of the cluster ofdifferentiation (CD) 3+ cells, ratio of CD4+/CD8+ cells and delayed-type hypersensitivity, and increased
proliferation of the splenic B cells, total serum immunoglobin (Ig) G and IgE, and proportion of splenicmIgM+mIgD+ cells). The authors concluded that chronic intake of high doses of DU (300 mg/kg) had asignificant impact on the immune function, most likely due to an imbalance in T helper Th1 andTh2 cytokines.
203. Studies on rats have addressed the effects of uranium exposure on the mucosal immune system of
the intestine following ingestion of uranium-contaminated drinking water [D26, D27, D28]) and inlungs following inhalation [M54, M56, P6]. Studies performed on the rat intestine found that Peyer’s
patches (aggregated structures of gut-associated lymphoid tissue) were a site of retention of uranium
following chronic ingestion (with uranium content in drinking water of 40 mg/L) after 9 months,
without inducing any biological effects on the function of Peyer’s patches [D26]. However, chronic
ingestion of uranium (nine months at 40 mg/L in drinking water) led, in the long term, to some changes
in the immune cell populations in lamina propria, diffuse gut-associated lymphoid tissue, notably
increase in neutrophil number (+300%) and decrease in macrophage (−50%), and mast cell number(−30%) in rats [D28].
204. Dublineau et al. indicated that immune cell populations in the intestine (neutrophils, mast cells,
macrophages) did not vary after 9 months of uranium exposure in rats, at any exposure level (from0.2 to 120 mg/L in drinking water) [D29]. The authors observed an increase in mRNA levels ofcytokines (IFN-gamma, IL-10, and CCL-2), at uranium levels in drinking water >20 mg/L, and adecreased protein expression of cytokines (IFN-gamma, IL-10, and TNFα). Such effects may be
explained by uranium interaction with proteasome [M6]. Changes in immune cell populations of theintestinal wall were not reported in the short term following acute contamination with a sublethal doseof uranium in rats (204 mg/kg of DU per os) [D27].
205. Immunological changes of long-term uranium exposure were investigated in mice fed withvarious DU doses (0, 3, 30 and 300 mg/kg food) [H11]. A four-month contamination of animals with
300 mg U/kg food led to a decreased secretion of nitric oxide, interleukin (IL)-1β, IL-18, and tumour
necrosis factor (TNF)-α by the peritoneal macrophages, a decreased proliferation of the splenic T cells
and increased proliferation of the splenic B cells associated with an increase in total serum
immunoglobin (Ig) G and IgE, and proportion of splenic mIgM(+)mIgD(+) cells. Such effects were notobserved for the lowest groups (3 and 30 mg U/kg food).
206. Some studies reported effects of uranium exposure on pulmonary immune cells following
inhalation in rats [M54, M56, P6]. An uptake of uranium by pulmonary macrophages was found(phagocytosis), followed by an accumulation of uranium in lysosomes [B20, M61]. This accumulation
was associated with release of pro- and anti-inflammatory cytokines in vivo [M54, M56] or in vitro[G5, W8, Z12]. Inflammatory cells were observed in proximity to aggregates of uranium particulates[L9]. This inflammatory cell response may lead to oedema and bronchial inflammation [B8, M59]. An
induction of apoptosis was observed in vitro in macrophages and T lymphocytes (CD4+) [W8].
207. In the follow-up of United States veterans of the first Gulf War published by McDiarmid et al.
[M25], the authors reported for the first time on the analysis of immune function in Gulf War I veterans

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 411

exposed to DU, investigating the lymphocyte response to stimulation by two T-cell mitogens. The
results on released cytokines (IFN, IL-10 and IL-2) in the high- and low-uranium exposure groups did
not provide evidence of an effect of uranium on the immune cells.
208. Haematological parameters were recorded in some human populations exposed to uranium.
Haematocrit, haemoglobin and RBC content from uranium processing site workers remained within thenormal ranges [S10]. Haemoglobin concentrations, erythrocyte counts, haematocrit values, and mean
corpuscular volumes were determined on a group of miners in a uranium mine [V6]. There were small
differences of mean values from those of controls, such as diminution of haemoglobin concentrations
and erythrocyte counts when uranium exposure time increased, but the mean values remained within
normal limits. A clinical study of Gulf War veterans showed that soldiers exposed to uranium had areduction in haemoglobin and haematocrit levels [S32].
209. In conclusion, it appears that slight modifications of the immune system, either systemic or
mucosal (including intestines and lungs), were induced following uranium exposure only at high doses,
at concentrations usually not incorporated by humans.

4.

Cardiovascular system

210. Few human data are available concerning uranium effects on the cardiovascular system. A study
was performed on people who had used drinking water from drilled wells containing high levels of
uranium (up to 1,500 µg/L) for several years. This study suggested that uranium exposure wasassociated with greater diastolic and systolic blood pressure [K28]. Two other studies were performed
among residents living near the Fernald Feed Materials Production Center, which functioned as auranium processing facility from 1951 to 1989 [P22, W2]. The systolic and diastolic mean bloodpressure levels were higher for population groups living near this uranium plant than those found in thegeneral population, and for most age- and sex-specific groups. However, neither distance nor direction
from the site influenced the blood pressure measurement results, suggesting that these findings are notexposure-related [P22]. Another study showed similar results: women with higher uranium exposure
had elevated systolic blood pressure compared to women with lower exposure, but the changes
observed in diastolic blood pressure or hypertension were not related to exposure level [W2].
211. Concerning experimental studies, there is a lack of literature on the effect of uranium exposure on
the development of cardiovascular diseases. One of the consequences of uranium-induced kidneydisease could be adverse effects on the cardiovascular systems by a number of mechanisms, includingchanges to the renin-angiotensin pathway. Several publications have focused on the relationshipbetween kidney and cardiovascular systems in humans [P1, R21, T2]. Consequently, the absence of
relevant publications precludes demonstration of an obvious link between uranium exposure andinduction of cardiovascular pathologies.

5.

Effects on DNA

212. Uranyl acetate staining of DNA has been used for staining fixed cells for more than forty years[H27]. In vitro studies performed with purified DNA showed the presence of a tight binding site for theuranyl ion (UO22+) in a four-way DNA junction [M53], demonstrating that uranium can be bound to
DNA in vitro. Exposure to uranium may lead to inhibition of DNA-binding proteins [H15, V7]. In vitro
studies demonstrated that uranium can displace transcription factor binding to DNA and that it can bindto serum proteins accumulation in the nucleus of human cells [H15, V7]. Experimental studies have

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UNSCEAR 2016 REPORT

shown accumulation of uranium in the nucleus of human cells (renal, hepatic, neuronal cells) treated invitro [G30, R26] and in kidney cells of rats treated in vivo [H23].
213. Depleted uranium can induce oxidative DNA damage [B3, K2, M40]. These studies suggested
that DU could induce DNA lesions through interaction with cellular oxygen species. These results were
corroborated by an in vivo study that showed that inhalation of DU resulted in a production ofhydroperoxides in lung tissue of rats [M56]. A recent in vitro experimental study reported the formationof uranium–DNA adducts and mutations in mammalian cells after direct exposure to a compound of
DU [S34]. The data suggest that uranium could be chemically genotoxic and mutagenic. The uniquemutation spectrum in hypoxanthine-guanine phosphoribosyl transferase (HPRT) locus elicited byexposure to DU suggests that uranium-generated mutations in ways that are different from spontaneous,
free radical and radiological mechanisms [C33].
214. Another study on rats demonstrated that UO2 repeated pre-exposure by inhalation increased DNA
single-strand breaks induced by a single UO4 inhalation exposure in epithelial nasal cells, bronchoalveolar lavage cells and kidney cells, whereas a single inhalation exposure to UO4 alone had no such
effect [M55]. In vitro studies indicated that uranium-induced DNA single-strand breaks were catalysedin the presence of ultraviolet light or ascorbate [M53, N7, Y2]. In vivo [M54, M55] and in vitro [D5,T6] studies showed that uranium induced DNA double-strand breaks. An in vitro study demonstrated
that DU was a weak clastogen2 and induced aneugenic3 effects while enriched uranium is a more potent
clastogen [D5].
215. In summary, experimental studies performed in vivo in rodents or in vitro in cell cultures
suggested that uranium could be chemically genotoxic and mutagenic through the formation of strand
breaks and covalent uranium-DNA adducts.

6.

Cytogenetic damage

216. Uranyl fluoride injected in vivo into mouse testes led to an increase in the frequency ofchromosomal aberrations in spermatogonia (breaks and polyploids) and primary spermatocytes(fragments, chromatid exchanges, reciprocal translocations) [H27]. An in vitro study performed onhuman bronchial fibroblasts indicated no significant increase in chromosomal damage following
chronic exposure to uranyl acetate and a slight increase following exposure to uranium trioxide [W22].
An increase in the frequency of sister chromatid exchange and in the frequency of chromosomal
aberrations was also observed in cell lines [L36] and in human osteosarcoma cells [M41].
217. A study by Hao et al. using rats showed that chronic oral exposure to DU in food led to anincreased frequency of micronuclei in bone marrow cells at 4 and 40 mg U/kg/day in weaning rats
[H9]. Miller et al. investigated the transmission of genetic damage to offspring of males micecontaminated with uranium via depleted-uranium-implanted pellets. The authors demonstrated a dosedependent increase in mutation frequency in the offspring [M45].
218. Chromosome aberrations in Gulf War I veterans following DU exposure were evaluated in several
studies [A8, B2, M24, N6]. Four biomarkers of genotoxicity (micronuclei, chromosome aberrations,mutant frequencies of HPRT and PigA4) were examined. There were no statistically significant
2

A clastogen is a mutagenic agent that induces disruption of chromosomes.
An aneugenic agent promotes aneuploidy in cells during mitosis or meiosis leading to the presence of abnormal number ofchromosomes in cells.
4
PigA gene mutation assay is used to assess genomic instability through the measurement of fluorescently labelled antibodies tospecific membrane proteins by flow-cytometric methodology.
3

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 413

differences in any outcome measure when results were compared between the low- and the highuranium groups. However, modelling suggests a possible threshold effect for mutant frequencies
occurring in the highest uranium exposed cohort members [M24]. This study demonstrates a relatively
weak genotoxic effect of the DU exposure. In the second study [B2], the results indicate that ongoing
systemic exposure to DU occurring in Gulf War I veterans with DU embedded fragments does not
induce significant increases in micronuclei in peripheral blood lymphocytes compared to micronuclei
frequencies in veterans with normal uranium body burdens.
219. Some studies investigated the possible effects of uranium exposure on chromosomal damage inworkers and reported an increase in chromosomal aberrations in miners [P25, Z3, Z4] whereas otherssuggested no significant effects [B49, K21, L42, M33, M70, W25]. Martin et al. performed a study forworkers from fuel production and fuel enrichment plants to analyse asymmetrical chromosome
aberrations and sister chromatid exchanges [M13]. Both worker groups had higher levels ofchromosome aberrations than the studied controls. This effect appeared not to be linked to externalradiation. Smoking increased the frequency of dicentrics but not the sister chromatid exchanges in theworkers exposed to soluble uranium, suggesting some interaction between the two clastogens.
220. In conclusion, some experimental studies performed on animals or on culture cells at high dosesof uranium demonstrate an increase in the frequency of chromosomal aberrations (fragments, chromatid
exchanges, reciprocal translocations). Human data are inconclusive but consistent with effects at higher
levels of exposure.

7.

Tumorigenic potential

221. Information about the carcinogenesis processes induced by uranium in experimental models andhumans was provided by the WHO [I5]. Publications on experimental models are detailed below.
222. Chronic inhalation of natural uranium ore dust alone created a risk of primary malignant and nonmalignant lung tumour formation in rats [M49]. The risk of the induction of malignant tumours was notdirectly proportional to dose but was directly proportional to dose rate. A study performed on micereported an increase in osteosarcoma and acute myeloid leukaemia with 239Pu and 241Am but nosignificant difference between results for 233U exposed animals and controls [E4]. The authors
considered that the observed differences between the radionuclides were due to differences in
irradiation of peripheral and central regions of the bone marrow, with lowest doses from 233U. No renal
and liver carcinomas were noted [E4].
223. A study by Hahn et al. conducted in vivo showed that intramuscular DU fragments induced soft
tissue sarcomas (fibrous histiocytoma, fibrosarcoma and osteosarcoma) at the site of implantation [H3].
However, no direct dose-dependent relationship could be drawn due to the presence of varying
corrosion products. Intravenous injection of murine haematopoietic cells into depleted-uraniumimplanted mice was followed by the development of leukaemia in 76% of mice implanted with DU
pellets in contrast to 12% of control mice [M44].
224. Miller et al. reported the ability of DU-uranyl chloride (10–250 µM) to transform immortalized
human osteoblastic cells to the tumorigenic phenotype [M39]. The changes in phenotype arecharacterized by anchorage-independent growth, tumour formation in nude mice, expression of highlevels of the k-ras oncogene, reduced production of the Rb tumour-suppressor protein, and elevatedlevels of sister chromatid exchanges per cell.

414

UNSCEAR 2016 REPORT225. Some sparse studies have been performed on DU clean-up workers and on populations fromdepleted-uranium-contaminated regions [K20, M35, M36]. In clean-up workers, the total number of
DNA alterations was higher immediately after decontamination than before decontamination, buttumours did not develop in the group of DU clean-up workers during the investigation period of four
years [M36]. This point is not surprising, since four years would be a short latency period for thedevelopment of any tumour. In summary, some publications reported tumorigenic effects of uranium inanimals (mice or rats) or tumorigenic potential in in vitro models.

I.

Relative biological effectiveness226. Relative biological effectiveness is an empirical value that measures the capacity of a specific
type of ionizing radiation to produce a biological effect in a particular biological system (for instancesurviving fraction of irradiated cells). It is the ratio of biological effectiveness of one type of ionizing
radiation relative to a reference radiation (gamma- or X-rays), given the same amount of absorbed
energy. Thus, in the case of uranium isotopes, these values refer exclusively to radiobiological effectsof alpha particles compared to the reference radiation. The chemical toxicity of uranium may influenceobserved RBE values. RBE values are used in experimental radiobiology and are also used by ICRP to
derive radiation weighting factors for the calculation of effective dose, which allows the summation ofradiation doses for all radiation types for the control of exposure. The value of the weighting factor
used by ICRP for alpha particles is 20 compared with a value of 1 for all low-LET radiation [I15].
While there are no direct determinations of alpha particle RBE for uranium isotopes, and such studies
would require the use of high specific activity isotopes such as 233U to avoid chemical effects, there is
no reason to consider that alpha particles from uranium will have different relative effectiveness fromalpha particles of similar energy emitted by other radionuclides (e.g. 222Rn and 239Pu [M12]).

J. Lethal effects
227. The lesion that can cause animal death after acute contamination is tubular nephritis, which, as
described above, results from the metallotoxic properties of uranium at high doses. Following
intravenous injection of a soluble form of uranium, the lethal dose (LD50) over a period of three weeks
after injection was measured as 0.1 mg/kg for rabbits and 20 mg/kg for mice [B20]. In rats, deathoccurred from 3 to 7 days for doses 8 mg/kg [F17]. Following oral ingestion of uranium acetate, theLD50 after two weeks was more than 100 mg/kg in rats and mice [D17].
228. One publication reported the deliberate ingestion of uranium (15 g of uranium acetate) [P9]. This
ingestion resulted in an acute renal failure, with persistent renal impairment six months after the initial
exposure. However, no death was reported despite the high uranium level ingested. This suggested thatseveral grams of uranium for ingestion intakes are necessary for inducing death in humans [K5].
229. As indicated in the recent review of Pernot et al. [P14], the definition and classification of thedifferent types of biomarkers have varied slightly, depending on the biomedical field considered. A
biomarker has been defined as “any measurement reflecting an interaction between a biological system
and an environmental agent, which may be chemical, physical or biological”. No real bioindicator is
specific for uranium exposure. Concerning uranium exposure, the analysis of uranium content (total

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 415

concentration or radiological activity in 238U, 234U and 235U) in urine [C18, L21, S37] may provide
information on uranium exposure. More indications are given in the uranium measurements section.
230. Since the kidneys are sensitive to damage by uranium, renal molecules were studied to describe
potential health consequences [G27, G28]. However, standard potential bioindicators have not yet
proven to be sufficiently specific and reproducible. The development and validation of bioindicatorsthat link uranium exposure to renal damage would be valuable [G27].
231. Currently, urinary levels of albumin, glucose, β2-microglobulin and N-acetyl-β-Dglucosaminidase activity are used to evaluate possible renal effects of uranium in humans [K26, K28]
or in animals [G27]. The use of osteopontin as a potential bioindicator of uranium effects was recently
investigated in a combined experiment and human study [P30]. A decreased osteopontin level in urine
was found when the concentration of uranium in urine, after acute exposure, exceeded 30 µg/L. Such a
decrease may suggest renal damage induced by uranium exposure. However, no clear relationshipbetween exposure level, duration of exposure and observed renal effect can be drawn from the available
studies on humans. One experimental study performed on rats at 40 mg/L of uranium in drinking water
failed to demonstrate variations in renal osteopontin mRNA level in the kidneys [R27].
232. Zamora et al. [Z6] studied the effects of chronic ingestion of uranium via drinking water onhuman populations. They found statistically significant subtle changes in two of the biologicalindicators measured, namely ratios of glucose and LDH to creatinine excreted (table 13) that did nottranslate to any observed health effects. In addition, LDH excretion decreased with higher exposure.
The excretion of β−2 microglobulin increased but the increase was not statistically significant.
233. Some biochemical parameters were measured in the urine of Gulf War veterans [M20, M22, M23,
M24, M25, S32]. Table 14 summarizes some of the results obtained. None of the levels measured were
statistically significantly different from control values. Glucose concentrations were lower for the highexposed veterans whereas the concentrations of -2 microglobulin and retinol binding protein werehigher. The results of the study indicated that there were a few subtle trends in changes of biochemicalparameter levels of exposure.
234. The use of omic techniques (genomics, proteomics and metabolomics) to screen unknown
biological or toxicological effects is expanding to develop new bioindicators. Omic methodology has
recently been used to screen mRNA, proteins or metabolome involved in the response to uranium
exposure in various cell lines or in animals [G21, M6, P27, P29, T4, T5]. A recent study has
demonstrated the relevance of metabolomics in cases of uranium exposure [G21]. The aim of this study
was to assess the biological changes in rats caused by ingestion of natural uranium in drinking waterover 9 months and to identify potential biomarkers related to such contamination. LC-MS
metabolomics identified urine as an appropriate biofluid for discriminating the experimental groups. Of
the 1,376 features detected in urine, the most discriminant molecules were metabolites involved intryptophan, nicotinate, and nicotinamide metabolic pathways, in particular N-methylnicotinamide.
These results are in accordance with a previous study that showed the role of N-methylnicotinamide inrats with experimental renal failure induced by uranium [S12]. The study of Grison et al. thus
establishes the validity of using metabolomics to address chronic low-dose uranium contamination[G21]. These studies show that while reliable biomarkers of tissue damage are not yet available,
modern techniques have the potential to identify such molecules.

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Table 13. Inter-subject variability in bioindicator data (adapted from [Z6])Low exposurea(<1 μg U/L (n=20))

High exposurea(2–781 μg/L (n=30))

55.9 (20.2–111)

82.4 (32.7–427)

Mass of LDH per unit mass ofcreatinine excreted (U/g)

25 (7.6–66)

17 (0–410)

Mass of -2 microglobulin per unit
mass of creatinine excreted (μg/g)

43 (11–270)

56 (15–340)

Urinary parameter
Mass of glucose excreted (mg/d)aMedian value with range provided in brackets. The group ‘Low exposure’ is considered as the control group.

Table 14. Bioindicator data for Gulf War veterans (adapted from [M23])
Normalrange

Low exposurea(<0.1 μg uranium per gramof creatinine (n=25))

High exposurea(≥0.1μg uranium per gramof creatinine (n=10))

Glucose g/day

0.0–0.5

6.3±4.38

0.12±0.01

-2 microglobulin μg/g creatinine

0–160

59.08±7.48

81.72±13.28

Urine retinol binding protein μg/g
creatinine<610

31.00±4.73

48.11±9.73

Urinary parameteramean ± standard error.

VIII. EPIDEMIOLOGICAL STUDIES
235. Several population groups have been considered for epidemiological studies on the health effects ofuranium due to occupational (uranium mines, milling and processing plants, facilities involved in the
nuclear fuel cycle) or to environmental exposure (elevated uranium levels in drinking water, vicinity ofuranium processing facilities or areas affected by depleted-uranium-munition use). Clear and comparative
epidemiological information is limited. Nonetheless, in appendix A, tables A1–A4 summarize theprincipal studies and characterize their value in assessing uranium risks for occupational exposure, andtables A5 to A7 for environmental exposure.

A. Studies of occupational exposure236. The preparation of the fuel used in nuclear power plants relies on a complex cycle, including thesteps of mining, crushing of the ore and preparation of Yellowcake (uranium mills and processing
plants), conversion of uranium oxide to UF6, enrichment, and fuel manufacturing. Other activities
include research and reprocessing of the nuclear fuel. These activities involve different types of jobs
and different patterns of uranium exposure. Also, historical methods of individual exposure monitoringvary (measurement of ambient exposure, bioassay of excreted uranium, personal film dosimeters, invivo measurements).

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 417

237. Studies of workers involved in the nuclear fuel cycle present a good potential to investigate cancer
and other health effects of internal uranium exposure on the basis of long-term follow-ups, insofar as
they have individual estimates of exposure levels. These studies have been identified as among the most
pertinent ones to quantify a potential exposure-risk relationship [L16].

1.

Uranium miners

238. Miners are exposed to internal contamination due to inhalation of long-lived radionuclides (LLR)
from uranium ore dust, but also to external gamma radiation and radon and its progeny. The main
source of radiological exposure of uranium miners is radon and its decay products. Many
epidemiological studies of uranium miners have been performed that demonstrated an association of
accumulated exposure to radon and its decay products with lung cancer risk [I16, L17, N1, U9, W17].
239. A concerted European effort (Alpha risk project) [T10] has furthered the assessment of healthrisks associated with uranium exposure based on the Czech, French CEA-COGEMA and German
Wismut cohorts. Individual exposure to uranium was estimated from measurement of ambientconcentration (in Bq/m3). The assessment of uranium contamination in miners was based on thereconstruction of individual accumulated ore dust exposure over time. The development of dedicateddosimetric software enabled the estimation of cumulative organ doses due to radon and its progeny,
LLR arising from uranium ore dust, and external gamma ray exposure [M10, M11]. This approachallowed the estimation of the contribution of uranium exposure to the total organ dose, and the initiative
has resulted in a number of cohort and nested case-control studies of uranium exposure and mortality
by cancers, leukaemia or cardiovascular diseases [D23, D31, K15, K17, M50, M51, M52, R4, R5, T13,
V2]. A difficulty with the uranium miner studies is that the lung doses from short-lived radon decay
products are much larger than and often correlated with the LLR exposure from uranium, whichhampers the assessment of exposure effects for lung cancer.

2.

Uranium millers

240. The first steps after uranium extraction are crushing of the ore and preparation of the Yellowcake
by uranium millers or as part of uranium processing. Like the miners, millers are subjected to inhalationof radon and radon decay products, external gamma radiation and inhalation of LLR from uranium ore
dust. One potential constraint of uranium miller studies is that, despite uranium LLR exposure being ahigher percentage of total radiation dose, uranium miller exposure is typically much lower thanuranium miner total exposure. Individual exposure to uranium among millers has been estimated solely
from measurement of ambient concentration (in Bq/m3).
241. Only a few cohort studies have specifically considered health risks among uranium millers[B36, K17, P21, Z1]. These studies reported mortality results, but only the German miller study
estimated individual LLR exposure from uranium ore dust and potential health risk [K18]. They
studied 4,054 millers who had never worked in underground or open pit mines and assessed mortality
from various diseases in relation to their exposure to radon, external gamma ray, LLR and silica dust
exposure [K18].

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3.

Nuclear fuel cycle workers

242. A variety of epidemiological studies enabled the estimation of health risks among workers
involved in the fuel nuclear cycle: in the United States (Fernald; Rocketdyne; Oak Ridge TEC, Y-12
and K-25; Paducah; Linde; Mallinckrodt), in the United Kingdom (UKAEA and AWE composite
cohorts, Sellafield), and in France (CEA; AREVA NC; Eurodif) [C5, Z9, Z10].
243. After milling, the nuclear fuel cycle entails different successive steps, including conversion,enrichment, fuel manufacturing, reprocessing and research. These steps involve diverse radiologicaland chemical exposure to different forms of uranium compounds. However, as these steps are notdistinguished in most of the epidemiological studies, they are considered together in this section.
Individual exposure to uranium has often been estimated solely from measurement of ambientconcentrations. Uranium exposure is preferably estimated from bioassays of individuals’ excreted
uranium or by other personal measurement (in vivo counting). In most studies exposure estimates forworkers could be based only on ambient measurements at job locations or even extrapolations for timeand location with no available ambient measurements.

4.

Results of occupational exposure studies

244. The exposure situations may influence cancer risk in occupational settings, especially for lungcancers. For miners, the main exposures are from radon (internal exposure), external gamma radiation,and exposure to uranium dust. Typically, workers in the nuclear fuel production are subjected to lessradon exposure but additional potentially toxic chemical exposure. In addition, for lung cancer,
confounding factors may occur due to variations in smoking habits, about which most studies containlittle or no information. Exposure levels have been assessed differently in various studies depending on
the occupational exposure setting and the historical exposure information available.
245. Characteristics and results of uranium miner, miller and processor studies are summarized inappendix A, tables A1–A4, with the studies in tables A3 and A4 having more quantitative analyses. The
health outcomes are described below by organ systems. Since well-conducted nested case-control
studies provide risk estimates that in principle are comparable to those from the full cohort, the resultsof both types of studies are provided together without distinction.
246. Studies that have identified uranium-exposed cohorts have varied in the amount of data theypresented to adduce uranium health effects. Studies that contain a subset of uranium-exposed workers
who are not actually identified as exposed provide essentially no information on uranium effects; moststudies in appendix A, table A1 fall into this category. Others have identified uranium workers but haveno individual exposure levels and therefore provide only standardized mortality ratios (SMRs) or otheroverall statistics which yield little valid information on uranium effects because they consider thosewith appreciable exposure and little or no exposure as a composite group. Also, SMRs are subject to“healthy worker bias”; such studies are shown mainly in appendix A, table A2. Analyses by duration of
employment can also be subject to “healthy worker survivor bias” and therefore should be viewed
circumspectly. Analyses that show SMRs or relative risks (RRs), including hazard ratios (HRs) andodds ratios (ORs) by cumulative dose groups are somewhat informative, but difficult to interpret since
the common problem of small numbers within groups leads to inconsistencies in risk estimates acrossgroups; trend analysis summaries are helpful but fail to provide quantitative overall risk estimates(appendix A, table A3). Studies that provide quantitative risk estimates per unit of exposure arepresented in appendix A, table A4.

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 419

247. Studies that explicitly estimate individual uranium exposure and use those data to calculatequantitative dose-response associations with specific health end points are valuable for assessing risk.
Typically, dosimetric estimates are made for detailed types of jobs at particular work locations for
specific periods based on ambient monitoring data, and this job exposure matrix (JEM) is applied toindividual work records. Better individual dosimetry is achieved when individual urinalyses of uranium
excretion or other personalized measurements are available, supplemented by ambient monitoring.
However, some quantitative studies had limited ability to distinguish effects from LLR exposure fromuranium from those associated with external irradiation or radon decay products because those doses
were much larger than the LLR doses to organs. That problem applies especially to lung cancer amongminers, since radon decay products usually contribute much more lung dose than LLR, so that in minerstudies the LLR lung dose was often only 1–2% as great as the radon decay product dose and the twodoses are typically correlated. In that situation, estimates of LLR-associated risk may be inaccurate toan unknown degree.
248. The following overview of associations of LLR exposure from uranium with diseases thought tobe induced by uranium will concentrate on the most informative studies that have presented quantitative
dose-response data. The most recent reports of study cohorts are shown in preference to older reports toavoid redundancy of information, since the most recent data with longer follow-up and sometimes
improved dose information provide better risk estimates.
249. The study results will be considered by organ systems. Essentially, all the informative studies ofoccupational populations exposed to uranium have been conducted since 1980, some with exposure
dating back to the mid-1940s. These studies considered a large range of potential health effects. Amongthe most frequently considered ones are respiratory cancers (lung, laryngeal), urological cancers(kidney, bladder, prostate), digestive cancers (stomach, intestine, liver, pancreas), cancers of brain andcentral nervous system, lympho-haematopoietic malignancies (leukaemia, lymphoma, multiple
myeloma). Non-cancer diseases have also been considered, especially respiratory, circulatory and renal
system diseases. Further details about these studies are presented in appendix A, table A4.

(a) Lung cancer and other respiratory diseases
250. Lung cancer. Studies with quantitative risk estimates for respiratory diseases from LLR exposure
from uranium are presented in table 15. Ten reports were found that had dose-response estimates forlung cancer risk based on individualized estimates of uranium exposure. The study of CEA-COGEMA
miners in France reported a positive risk coefficient for LLR exposure from uranium, though the radonand external-radiation co-exposure make the result difficult to interpret. Other miner studies have notcharacterized and analysed LLR exposure from uranium and lung cancer. The German study ofuranium millers [K18] reported a dose-response analysis of LRR exposure from uranium and lungcancer risk. The association was non-significant and in the negative direction. Two other studies ofuranium millers reported no significant elevation in lung cancer mortality [B36, P21].
251. The remaining eight reports of LLR dose-related lung cancer mortality were based on workers innuclear processing industries. The nature of the workplaces and the types of measurements andprocedures that generated the individual exposure estimates are given in appendix A, table A4, for thevarious studies. The large study of Fernald workers, with 269 lung cancer deaths, showed a positive,
but non-significant, association and most other studies had negative or non-significant coefficients ofrisk. Additional uranium worker cohorts showed no significant elevation in overall lung cancer SMRs
[C36, D33, D34, M26, M27]. However, the relatively small French study of the AREVA NC Pierrelatte
plant showed significant positive dose responses [G31]. This is the only study that conducted separate
analyses for exposure to uranium compounds that differ in solubility (Types: F, M and S). They found420

UNSCEAR 2016 REPORT

the Type M and Type S exposure conferred lung cancer risk (table 15). In a further analysis [G32] that
distinguished exposure to natural uranium from the reprocessed uranium which has a different isotope
composition, they found that risks were especially dose related for reprocessed uranium of the Type M
and Type S. This suggests that insoluble uranium particles with a longer residence time in the lungconfer more risk than the more soluble particles with a shorter residence time. However, theseinteresting results were based less than 55 lung cancer deaths and require confirmation in larger studies.
Table 15. Dose-response studies of uranium exposure and risk of respiratory system diseases
Study references

No. ofdeathsNature ofuranium work

Unit of uranium (LLR)
exposure

Risk estimate per unit
LLR exposurea

LUNG CANCER
France, CEACOGEMA [R5]
(Also [R4, V2])

94

MiningkBqh/m3

ERR=0.32 (95% CI: 0.09, 0.73)

Germany, milling
[K18]

159

Milling

100 kBqh/m3

ERR=−0.61 (95% CI: −1.42, 1.9)

France, AREVA NC
[C6]

48

Processing

Years of Types F, M,
and S uranium
exposure

HR=1.01 (95% CI: 0.96, 1.01) [F]
1.07 (1.01, 1.13) [M]
1.07 (1.01, 1.14) [S]France, AREVA NC
[G32]

53

Processing

Years of exposure to
Types F, M and S
reprocessed uranium

HR=1.07 (95% CI: 0.96, 1.19) [F]
1.13 (1.03, 1.25) [M]
1.13 (1.01, 1.25) [S]

USA, TEC/Y-12/
Mallinckrodt/
Fernald [D35]b

787

Processing

0.5, 2.5, 5, 25, 50,≥250 mGy LLR

USA, Rocketdyne
[B38]

94

Processing

100 mSv LLR

USA, Fernald [S19]269

ProcessingmGy LLR

ERR=0.022 (95% CI: −0.009, 0.06)

USA, Paducah
gaseousdiffusion [C17]129

Processing

μg.y

RR=0.91 (95% CI: 0.5, 1.6) (21–50 μg.y)
0.95 (0.6, 1.6) (51–125 μg.y)
0.51 (0.3, 0.9) (>125 μg.y)

USA, Y-12 [R12]111

Processing

10 mSv LLRFrance, Gaseous
diffusion [Z10]100

Processing

Low, medium, high

OR=1.03, 0.57, 0.85, 0.82, 0.64, 2.05
(95% CI: 0.20, 21), respectivelyc
RR=1.01 (95% CI: 0.89, 1.16)

ERR=–0.77 (95% CI: –2.5, 1.0)
Exposure to natural soluble uranium
compounds (with 90% CI):
Medium: RR=0.92 (CI: 0.54, 1.6)
High: RR=0.74 (CI: 0.42, 1.3)
Enriched uranium (n=23):
Medium RR=1.8 (CI: 0.64, 4.6)
High RR=0.69 (CI: 0.21, 1.9)
Depleted uranium (n=23):
Medium RR=1.2 (CI: 0.33, 3.7)
High RR=1.5 (CI: 0.61, 3.9)

EXTRATHORACIC RESPIRATORY CANCERS
Germany, Wismut
[M51]

554

Miners

10 kBqh/m3

ERR=0.098 (95% CI: –0.11, 0.31)(Laryngeal cancer)

Germany, Wismut
[K17]

234

Miners

100 kBqh/m3

ERR=−0.17 (95% CI: −2.50, 2.16)(All extra-thoracic airway cancers)

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 421

Study references

No. ofdeathsNature ofuranium work

Unit of uranium (LLR)
exposure

Risk estimate per unit
LLR exposurea

NON-MALIGNANT RESPIRATORY DISEASEFrance, CEACOGEMA [R5]37

MinerskBqh/m3

ERR=−0.086 (95% CI: n.e.)dUSA, Y-12 [R12]50

Processing

100 mSv

ERR=−0.085 (90% CI: −0.32, 0.15)

USA, Fernald [S19]e102

Processing

100 μGy LLR

ERR=−0.0062 (95% CI: −0.007, 0.0006)a

Risk estimation metrics: ERR (excess relative risk) for which zero represents no excess or deficit; HR (hazard ratio) and RR(relative risk or rate ratio) are expressed as a multiple of the rate in the baseline (lowest/no exposure) group. (ERR = RR – 1).
bThis study [D35] was partially repeated later on [R12, S19] but included two cohorts that were not reported elsewhere.

c

The risk estimate of 2.05 for the ≥250 mGy group became 0.36 in an analysis with adjustment for smoking.

d

Not estimable from the likelihood profile.

eAnalysis of chronic obstructive pulmonary disease.

252. In summary, there is mixed evidence for the lung carcinogenicity of uranium. The inconsistency
may relate to variations in the amount and types of exposure data available (e.g. extrapolated dosereconstruction, ambient monitoring, uranium urinalyses or other personalized exposure measures) or tovariations in methods for calculating LLR doses from the raw data. The inconsistency can also relate toother radiation or chemical exposure that may not have been adequately accounted for in the LLR
analyses, or to dose-dependent variations in smoking behaviours. On the other hand, since the studies
have low statistical power to detect risks, given the relatively small numbers of lung cancers and fairly
low levels of LLR exposure from uranium for most workers, it is notable that two cohorts with doseresponse data showed a statistically significant association [C6, R5].
253. Other respiratory cancers. A study of all extrathoracic airway cancers, based on a well-definedcohort of the German uranium miners, showed a negative dose-response coefficient for LLR exposure
from uranium [K17]. Another study of German uranium miners had LLR dose-response data forlaryngeal cancer, which yielded a non-significantly positive risk coefficient [M51]. There was no
association of LLR exposure from uranium and laryngeal cancer among French uranium miners [R5].
Other studies have reported overall excess laryngeal or upper airway cancers in worker populations
potentially exposed to uranium [B18, B29, B33, D35, G31], but those are based on small numbers ofcancers, without individual uranium exposure estimates, and with limited information on the risk
factors of smoking and alcohol intake.
254. Non-malignant respiratory disease (NMRD). Four studies with dose-response analyses allreported negative dose-response coefficients for NMRD [B38, R5, R12, S19]. Pinkerton et al. [P21]reported an inverse relationship of NMRD with length of time working in uranium milling. SMRs in a
number of other cohorts with uranium exposure were not significantly elevated [C20, C36, D33, D34,
M26, M27, R14, Z10] so there is no evidence that NMRD is associated with uranium exposure.

(b) Lymphatic and haematopoietic cancer
255. Leukaemia. Since leukaemia shows a large excess relative risk from radiation exposure, it is astrong a priori candidate to investigate regarding uranium exposure effects. Table 16 shows uranium
dose-response related results for leukaemia. A nested case-control study of leukaemia mortality within
the large collection of German Wismut miners [M50] found a positive but non-significant associationof estimated LLR exposure with non-chronic lymphocytic leukaemia (non-CLL) risk. The positive risk
coefficient was driven entirely by the highest dose group (≥20 kBqh/m3, ERR=1.26, 90% CI: 0.7, 2.2,

422

UNSCEAR 2016 REPORTn=14). A smaller miner study in Czechia found a statistically significant dose-response association ofleukaemia mortality with total red bone marrow dose, of which the majority was from LLR exposure[T13]. Three smaller dose-response studies of leukaemia risk in the uranium processing industries did
not support excess risk (table 16). In summary, there is limited dose-response evidence for an
association of uranium exposure with subsequent leukaemia. Atkinson et al. [A29] reported a nonsignificant SMR (SMR=1.10; 95% CI: 0.89, 1.33; n=103) in nuclear workers, particularly among
workers monitored for internal radiation exposure. Others likewise have reported non-significant
overall SMRs in uranium worker cohorts [C36, D30, D33, D34, L44, M26, M27, P24, S5, Z2, Z10], but
did not estimate worker uranium doses or conduct LLR dose-related analyses.
256. Other lympho-haematopoietic malignancies. Lymphoma is a biologically plausible outcome of
inhalation exposure to uranium, since uranium deposited in the lung tends to migrate to the thoracic
lymph nodes. However, there are few studies providing dose-dependent analyses of uranium exposureand lympho-haematopoietic malignancies other than leukaemia (table 16). The cohort of German
uranium millers [K18] showed a non-significant negative risk coefficient for all lympho-haematopoietic
malignancies. In the Rocketdyne cohort there was a non-significant positive dose-response [B38],
though the specifics were not reported. A French study of gaseous diffusion plant workers reported a
non-significant excess risk for all lympho-haematopoietic malignancies other than leukaemia among
those with medium or high exposure to soluble uranium [Z10]. A study in the United States showed for
uranium miners an elevated risk of lympho-haematopoietic malignancies (SMR=1.38, 95% CI: 1.03,
1.82) [S5]. Pinkerton et al [P21] showed a suggestive excess of lymphatic and haematopoietic
malignancies, excluding leukaemia, (SMR=1.44, 95% CI: 0.83, 2.35, n=16) with no dose response.
257. A study of non-Hodgkin’s lymphoma (NHL) at the Paducah gaseous diffusion plant [C17]produced high relative risks, but the fact that there was no dose response suggests that the high values
are likely attributable to a deficit of NHL in the baseline comparison group rather than large excesses in
the other dose groups. NHL mortality was non-significantly elevated for exposure to uranium in the
Fernald cohort [S19]. The Czech uranium miner study [T13] indicated a non-significantly elevated
SMR of 1.4 (95% CI: 0.9, 5.1, n=16) for NHL but did not provide a dose-response analysis. Other
uranium worker cohorts reported null overall SMRs for lymphomas or all lympho-haematopoietic
cancers [B38, C9, C36, D30, D33, D34, G31, L44, M27, P21, P24, S5, Z2].
258. In the study at the Oak Ridge K-25 gaseous diffusion plant by Yiin et al. [Y3] with a doseresponse analysis of multiple myeloma and uranium exposure, a significant association was not found.
A significant overall excess of multiple myeloma was found among United States miners (SMR=1.97,
95% CI: 1.05, 3.37, n=13), and an excess based on two cases (SMR=8.38; 90% CI: 1.44, 26.20) was
reported among French CEA-COGEMA uranium processing workers [B9]. A non-significant excess ofmultiple myeloma was observed among Mallinckrodt uranium processing workers [D33] (SMR=1.30,
95% CI: 0.42, 3.0, n=5). The SMRs for multiple myeloma were not significantly elevated in otheruranium worker groups [C17, Z2].

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 423

Table 16. Dose-response studies of uranium exposure and risk of lympho-haematopoietic
malignancy mortality
Study references

No. of deaths

Germany, Wismut
[M50]

b

Nature ofuranium work

Unit of uranium(LLR) exposure

Risk estimate per unit LLR
exposure a

LEUKAEMIA
218 (non-CLL)

Mining100 kBqh/m3

Czechia [T13]30

Mining

Svc

USA, Paducah[C17]

21 (all types)

Processing#1 (0–20 μg.y)#2 (21–50 μg.y)#3 (51–125 μg.y)#4 (>125 μg.y)

Baseline#2: RR=0.73 (95% CI: 0.2, 3.0)
#3: RR=0.49 (CI: 0.2, 2.3)
#4: RR=0.77 (CI: 0.2, 2.5)

USA, Rocketdyne[B38]

10 (non-CLL)

Processing

100 mSv LLR

RR=1.06 (95% CI: 0.50, 2.23)

USA, Fernald [S19]28 (non-CLL)

Processing

100 μGy LLR

HR=0.18 (95% CI: 0.012, 0.80)

ERR=0.76 (90% CI: –1.26, 2.78)
ERR=2.5 (90% CI: 0.3, 9.3)

OTHER LYMPHO-HAEMATOPOIETIC MALIGNANCIES
Germany [K18]23 (all lymphohaematopoietic)

Millers

Zhivin, Gaseous
diffusion [Z10]28 (all lymphohaematopoietic)

Processing

100 kBqh/m3

ERR= –0.65 (95% CI: –2.78, 1.47)

Medium and highexposure

Natural soluble uranium
compounds (with 90% CI):
Medium RR=1.4 (CI: 0.52, 3.9)
High RR=1.08 (CI: 0.37, 3.3)

USA, Paducah[C17]

26 (non-Hodgkin’s
lymphoma)

Processing#1 (0–20 μg.y)#2 (21–50 μg.y)#3 (51–125 μg.y)#4 (>125 μg.y)

USA, Fernald [S19]12 (non-Hodgkin’s
lymphoma)

Processing

100 μGy

HR=1.2 (95% CI: 0.88, 1.5)

98 (multiple
myeloma)

Processing

10 μSv

OR, 1.04 (95% CI: 1.00, 1.09)

USA, K-25 [Y3]

Baseline#2: RR=9.95 (95% CI: 1.2, 81)#3: RR=8.85 (CI: 1.1, 71)#4: RR=5.74 (CI: 0.7, 45)aRisk estimation metrics: ERR (excess relative risk) for which zero represents no excess or deficit; HR (hazard ratio) and RR(relative risk or rate ratio) are expressed as a multiple of the rate in the baseline (lowest/no exposure) group. (ERR = RR – 1).
b

Non-CLL, all leukaemias except chronic lymphocytic leukaemia.

cAnalysis was for total red bone marrow dose, of which 52–64% was estimated to be due to LLR from uranium dust inhalation.

(c) Digestive system cancer
259. Stomach cancer. Only a few studies have conducted dose-dependent analyses of uranium
exposure and digestive system cancers (table 17). Studies of German uranium millers [K18], theFernald uranium processors [S19], and Rocketdyne workers with internal exposure monitoring [B38]
all showed weakly positive but non-significant dose-response risk estimates for stomach cancer. The
risk of mortality from stomach cancer was increased with the alpha absorbed stomach dose among
German uranium miners (ERR per Gy=37.3; 95% CI: 3.4, 71.1, n=592), but the contribution of LLR to424

UNSCEAR 2016 REPORT

the absorbed stomach dose was less than 1% [K15]. Other uranium worker cohorts reported null overall
SMRs for stomach cancer [B38, C9, C17, D34, L44, M26, R5, S5, Z1, Z10].
260. Intestinal cancer. The study of German uranium millers showed non-significant ERR coefficientsfor both colon and rectal cancers [K18], and both the French uranium miners study [R5] and theRocketdyne study [B38] reported non-significant LRR dose coefficients for colorectal cancers. On theother hand, a dose-response analysis of combined small intestine and colon (but not rectum) cancer at
the United States Fernald uranium processing plant yielded a statistically significant excess risk [S19].
Since this result was based on relatively small numbers, it is in need of confirmation by larger studies.
Other uranium worker cohorts reported null overall SMRs for colon or colorectal cancers [C9, D33,
D34, G31, L44, M26, P21, Z1, Z10] but did not conduct LLR dose-related analyses.
261. Pancreatic cancer. The United States Paducah [C17] and Fernald [S19] studies provided doseresponse analyses of pancreatic cancer after LLR exposure (table 17). Neither study provided anindication of an association with uranium exposure, nor did the studies of French uranium miners [R5]or Rocketdyne workers [B38]. However, the studies had relatively small numbers of pancreatic cancers
and thus limited statistical power. Other reports have indicated non-significant SMRs for pancreaticcancer in uranium processing workers [C36, D33, D35, L44, M26, Z1, Z10] but did not estimate
worker uranium doses or conduct LLR dose-related analyses.
Table 17. Dose-response studies of uranium exposure and risk of digestive system disease mortality
Study references

No. of deaths

Nature ofuranium work

Unit of uranium(LLR) exposure

Risk estimate per unit LLR
exposure a

STOMACH CANCER
Germany, [K18]49

Milling

USA, Fernald [S19]29

Processing

100 kBqh/m3

ERR=1.5 (95% CI: −2.9, 5.9)b100 μGy

ERR=0.041 (95% CI: −0.20, 5.6)

INTESTINAL CANCER
Germany, [K18]22 (colon)

Milling

100 kBqh/m3

ERR=–0.07 (95% CI: –3.3, 3.2)

Germany, [K18]26 (rectum)

Milling

100 kBqh/m3

ERR=0.56 (95% CI: –3.1, 4.2)

USA, Fernald [S19]48 (colon &small intestine)

Processing

100 μGy

ERR=1.5 (95% CI: 0.12, 4.1)

PANCREATIC CANCER

a

USA, Paducah [C17]30

Processing#1 (0–20 μg.y)#2 (21–50 μg.y)#3 (51–125 μg.y)#4 (>125 μg.y)

Baseline#2: RR=1.42 (95% CI: 0.4, 4.7)
#3: RR=0.49 (95% CI: 0.1, 1.9)
#4: RR=0.97 (95% CI: 0.3, 3.0)

USA, Fernald [S19]41

Processing

100 μGy

HR= 0.61 (95% CI: 0.015, 3.5)

Risk estimation metrics: ERR (excess relative risk) for which zero represents no excess or deficit; HR (hazard ratio), RR(relative risk or rate ratio) and OR (odds ratio for case-control studies) are expressed as a multiple of the rate in the baseline(lowest/no exposure) group. (ERR = RR – 1).
b

Estimate is adjusted for radon exposure levels.

262. Liver cancer. Dufey et al. [D31] observed a non-significant increase in liver cancer mortality risk
associated with high-LET absorbed liver dose among German uranium miners (ERR per Gy=48.3; 95%
CI: −32.0, 128.6 (n=159) after adjustment for low-LET dose). However, the contribution of LLR to theabsorbed liver dose was less than 2%, so the association is not informative regarding uranium risk.
Among French uranium miners there was also a non-significant association of LLR exposure and liverANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 425

cancer risk [R5]. Other uranium worker cohorts reported non-significant SMRs for liver cancer [B38,
C9, M26, M27, P24, R5, Z10] but had no uranium dose-response analyses.
263. In summary, there is no persuasive evidence of an association between uranium exposure anddigestive cancers. The effect, if any, is likely small. However, the relatively small numbers of digestivecancer cases and consequent limited statistical power to detect effects make any conclusion uncertain.

(d) Kidney and other urological cancers
264. Kidney cancer. The toxicological data suggest that uranium exposure may be related to urologicalcancers, especially to the kidney because of the potential for both adverse radiological and metal effects
upon that organ. The available epidemiological studies with dose-response results are shown intable 18. Dose-response analyses of kidney cancer in relation to LLR exposure have been conducted forthe cohorts of miners in France and Germany [D23]. Neither cohort showed a significant association
with kidney cancer even though the large German miner cohort had a substantial number of renal
cancers. In the French CEA-COGEMA cohort, the overall SMR was elevated for kidney cancer(SMR=2.0, 95% CI: 1.2, 3.1, n=20), but LLR analyses were not reported [V1].
Table 18. Dose-response studies of uranium exposure and risk of urological system disease mortality
Study references

No. of deaths

Nature ofuranium work

Unit of uranium(LLR) exposure

Risk estimate per unit LLR
exposure a

KIDNEY CANCER
France, CEACOGEMA [D23]11

MiningkBqh/m3

HR=0.89 (95% CI: 0.55, 1.42)

Germany, Wismut
[D23]

174

MiningkBqh/m3

HR=1.009 (95% CI: 0.991, 1.027)

Germany [K18]11

Milling

100 kBqh/m3

USA, Fernald [S19]15

Processing

100 μGy

ERR=7.36 (95% CI: –11, 26)
ERR=0.039 (95% CI: −0.021, 0.55)

OTHER UROLOGICAL DISEASES
Germany [K18]30 (Prostatecancer)Milling

USA, Fernald [S19]19 (chronic
kidney disease)

Processing

100 kBqh/m3

ERR=0.21 (95% CI: –2.8, 2.4)

100 μGy

HR=0.98 (95% CI: 0.78, 1.1)aRisk estimation metrics: ERR (excess relative risk) for which zero represents no excess or deficit; HR (hazard ratio) and RR(relative risk or rate ratio) are expressed as a multiple of the rate in the baseline (lowest/no exposure) group. (ERR = RR – 1).

265. The study of German uranium millers [K18] had a large risk coefficient for kidney cancer but theassociation was non-significant because the number of cases was small (n=11) and consequently theconfidence interval of the estimate was very wide. The dose-response coefficients for the Fernald
uranium processing workers [S19], the Rocketdyne workers [B38] and the French uranium miners [R5]were all non-significant. Overall SMRs for kidney cancer were found to be increased in two uraniumprocessing facilities: Y-12 [C20, L44] and Capenhurst [M26]. But at Capenhurst, the increase waslimited to unexposed workers. The UK Nuclear research workers also experienced elevated mortality
from kidney cancer [B18, C8]. However, none of these studies included an investigation of the relation
with internal exposure to uranium. Other uranium worker studies did not report significantly elevatedoverall SMRs for kidney cancer [C9, C17, L44, S5, Z10] but had no uranium dose-response analyses.

426

UNSCEAR 2016 REPORT266. Bladder cancer. French uranium miner studies reported a non-significant LLR dose-response
association with bladder cancer [R5]. An increase in mortality (an elevated overall SMR) from bladdercancer was reported among Fernald workers [D35]. It was potentially associated with a high exposure
to the cutting fluids used during uranium metal production but not with internal exposure to uranium
compounds [R14]. In a more recent study of the Fernald workers [S19], elevated mortality wasobserved only among salaried female workers; for cancer of the bladder and other urinary organs, theSMR was 5.13 (95% CI: 1.06, 15.0; n=3). However, the small number of cases, no increase in bladdercancer mortality among males (n=21) or hourly paid female workers (n=0), and no dose-response
information argue against a meaningful excess [S19]. Other reports have indicated non-significant risks(usually SMRs) for bladder cancer in uranium processing workers [B38, Z1, Z10], but did not estimate
worker uranium doses or conduct LLR dose-related analyses.
267. Prostate cancer. The recent study of German uranium millers analysed uranium dose and prostatecancer risk and found no evidence of an association [K18], albeit the number of prostate cancer deaths
was small (table 18). The study of United States Rocketdyne workers reported a non-significant doseresponse trend in the negative direction for prostate cancer (n=63), but did not provide a risk coefficient[B38]. The LLR dose-response trend among French uranium miners was likewise non-significant [R5].
268. Radiation-exposed worker cohorts with excess prostate cancer have been investigated for internal
radionuclide exposure in the UK Atomic Energy Authority cohorts [A28, B18, R22] but with only alimited examination of uranium exposure. Rooney et al. [R22] conducted a nested case-control
examination of the relation between prostate cancer and occupational exposure in those cohorts,
including 136 prostate cancer cases and 404 controls and examination of 29 radionuclides. Three cases
and twelve controls had an indication of potential uranium exposure, and zero cases and six controls
had documented uranium exposure (RR=0, 95% CI: 0, 2.55), so no effect on prostate cancer mortality
was observed. Other studies have reported non-significant overall SMRs for prostate cancer among
uranium workers [D33, L44, M26, Z1, Z10] but had no uranium dose-response analyses.
269. In summary, except for the German Wismut cohort, the number of kidney cancers was very small,thus constraining the ability to detect small-to-moderate effects. Since the studies of kidney cancer andchronic kidney disease are uniformly negative, the risk of uranium exposure for kidney cancer is weakor absent. The data do not generally support an increase in bladder cancer associated with uranium
exposure; in fact, the only dose-response report regarding bladder cancer risk had a negative coefficient[B38]. The two studies that have evaluated dose-response data for prostate cancer did not show asignificant association.

(e) Brain and central nervous system cancers
270. Experimental evidence indicates that uranium compounds, particularly more soluble ones, crossthe blood-brain barrier, thereby potentially putting the brain at risk. There have been only a small
number of studies with dose-response analyses of brain and central nervous system (brain/CNS)
tumours (table 19). An analysis of brain/CNS tumours by LLR dose groups was conducted among
Paducah gaseous diffusion uranium enrichment workers [C17], but no trend by dose was seen. A studyof French uranium miners reported a non-significantly positive brain/CNS tumour dose-response risk
for LLR exposure, accompanied by an overall excess (SMR=1.71, 95% CI: 1.00, 2.74) [R5]. Similarly,
three United States studies of uranium processing workers did not find a dose-related excess ofbrain/CNS tumours [B38, C17, C20].
271. Carpenter et al. [C7] investigated the possible association between brain/CNS cancers andexposure to external and internal radiation among Oak Ridge Y-12 workers. The internal dose to theANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 427

lung, calculated as in the study by Checkoway et al. [C20], was used as a surrogate for the internal doseto the brain for the 47 cases and 120 matched controls. Odds ratios (ORs) were non-significantly
elevated for categories of cumulative lung dose: ≥150 to 290 mSv (OR=2.8; 95% CI: 0.7, 11.9; n=5);≥300 to 450 mSv (OR=2.7; 95% CI: 0.8, 9.3; n=5); and among workers with mean annual lung dose>150 mSv (OR=1.7; 95% CI: 0.7, 4.2; n=16). No dose–response trend was observed after adjustment
for possible confounding factors (26 different chemicals, socio-economic status, duration of
employment), using either a 5- or 10-year dose lag. Other studies of uranium worker cohorts indicatedno significant overall excess of brain/CNS tumour deaths [C9, D33, L44, M26, M27, P24, R14, Z10]
but did not estimate worker uranium doses or conduct LLR dose-related analyses.
272. A limiting factor in all of these studies was the very small numbers of tumours, so only quite alarge excess risk would be detectable. Clearly more data are needed to make a better judgment aboutbrain/CNS tumour risk from uranium exposure.
Table 19. Dose-response studies of uranium exposure and risk of brain and central nervous system(CNS) tumour mortality
Study references

No. ofdeathsNature ofuranium work

Unit of uranium(LLR) exposure

Risk estimate per unit LLR
exposure a

BRAIN/CNS TUMOURS
kBqh/m3

ERR=0.28 (95% CI: n.e. <0, 1.87)b

Processing#1 (0–20 μg.y)#2 (21–50 μg.y)#3 (51–125 μg.y)#4 (>125 μg.y)

Baseline#2: RR=0.66 (95% CI: 0.1, 4.2)
#3: RR=1.07 (95% CI: 0.2, 4.8)
#4: RR=0.45 (95% CI: 0.1, 2.2)

Processing#1 (0–9 mSv)
#2 (10–49 mSv)
#3 (≥50 mSv LLR)

Baseline#2: RR=1.10 (95% CI: 0.2, 6.5)
#3: RR=0.45 (95% CI: 0.1, 3.2)

France, CEACOGEMA [R5]
[Also [V2]]17

Mining

USA, Paducah [C17]14USA, Y-12 [C20]14aRisk estimation metrics: ERR (excess relative risk) for which zero represents no excess or deficit; HR (hazard ratio) and RR(relative risk or rate ratio) are expressed as a multiple of the rate in the baseline (lowest/no exposure) group. (ERR = RR – 1).
bn.e. = not estimable from the likelihood profile.

(f) Circulatory diseases
273. Suggestive findings from external radiation studies over the past 15–20 years have prompted
examinations of circulatory disease risks as a potential radiation effect of internal radionuclides. The
available studies of uranium exposure have evaluated the risk for all circulatory system diseases (CSD)
and major subcategories of CSD, namely ischaemic heart diseases (IHD) and cerebrovascular diseases(CeVD) (table 20).
274. Circulatory system diseases. The German Wismut miner cohort is the largest to study uranium
associated CSD risk. They reported a negative CSD risk coefficient for internal LLR exposure on thebasis of 5,417 CSD deaths [K14]. Two reports of CSD in the French CEA-COGEMA miner cohort areof interest (table 20). A report of the entire cohort indicated a non-significantly positive risk coefficient[R5], while another report of the subset of the cohort for whom information on radon, external radiation
exposure, and medical risk factors for CSD also was available showed a nearly significant risk of CSD
from LLR exposure after accounting for the other factors [D24].

428

UNSCEAR 2016 REPORT275. A report of the French AREVA NC uranium processing cohort found that exposure to Type M
and Type S (less soluble) reprocessed uranium conferred statistically significant risk of CSD, as did
Type S natural uranium exposure [G33], but Type F exposure did not. The study found statisticallysignificant associations of CSD with the number of years exposed to Type S exposure to both
reprocessed and natural uranium, and a near-significant association for Type M reprocessed uranium(table 20) [G33]. Less quantitative but supportive findings were that CSD mortality was increased
overall among workers exposed to slowly soluble reprocessed (HR=2.13; 95% CI: 0.96, 4.70) andnatural uranium (HR=1.73; 95% CI: 1.11, 2.69). In the subgroup of smokers, the risk estimates werehigher but with larger CIs (HR=1.91; 95% CI: 0.92, 3.98 for natural uranium and HR=4.78; 95% CI:
1.38, 16.50 for reprocessed uranium). The AREVA NC findings suggested that types of uranium with along residence time in tissues may confer risk of CSD.
276. McGeoghegan et al. [M29] reported an association between mortality from CSD and radiation
exposure in males performing industrial work in the British Nuclear Fuels Limited cohort (BNFL). This
cohort consisted partially of uranium workers (37% of the cohort were workers employed at the
Springfields uranium processing plant and 6.7% at the Capenhurst uranium enrichment plant). Their
analysis of internal exposure to any radionuclide gave a dose-response ERR per Gy of 0.76 (90% CI:
0.37, 1.23; n=2,275), but a dose-response analysis was not available specifically for uranium exposure.
277. Ischaemic heart diseases. The large German Wismut study of uranium miners found a negativerisk coefficient for LLR exposure and the end point of all heart disease (table 20) [K14]. The otherstudies evaluated IHD as a cardiovascular end point. Studies of French uranium miners (CEACOGEMA) reported negative risk coefficients for IHD for the entire cohort [R5] and for the subset
where they could account for radon, external radiation, and medical heart disease risk factors [D24].
The German uranium miller study also reported a non-significantly negative risk coefficient for LLR
dose [K18]. McGeoghegan et al. [M29] reported an association between mortality from IHD andradiation exposure in males performing industrial work at BNFL. Their analysis of internal exposure to
any radionuclide gave a dose-response ERR per Gy of 0.52 (90% CI: 0.09, 1.06; n=1,494), but uranium
exposure was not analysed.
278. Cerebrovascular diseases. The same set of studies also provided risk estimates for CeVD
(table 20). The large German Wismut study of uranium miners found a negative risk coefficient for
LLR exposure and CeVD [K14]. Studies of French uranium miners (CEA-COGEMA) reported nonsignificant positive risk coefficients for CeVD for the entire cohort [R5] and for the subset withadjustment for radon, external radiation, and medical heart disease risk factors [D24]. The German
uranium miller study reported a non-significantly negative risk coefficient for LLR dose [K18]. In the
study of French AREVA NC uranium processing workers, there was a statistically significant risk forType S exposure to reprocessed uranium and a near-significant risk for Type M reprocessed uranium[G33], but little indication of risk from natural uranium exposure. Further, McGeoghegan et al. [M29]also reported an association between mortality from CeVD and radiation exposure in males performing
industrial work in the BNFL cohort. Their analysis of internal exposure to any radionuclide gave adose-response ERR per Gy of 1.47 (90% CI: 0.49, 3.00; n=456).

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 429

Table 20. Dose-response studies of uranium exposure and risk of circulatory system disease mortality
Study references

Nature ofuranium work

No. of deaths

Unit of uranium(LLR) exposure

Risk estimate
per unit LLR exposure a

ALL CIRCULATORY SYSTEM DISEASES
Germany, Wismut
[K14]

5 417

Mining100 kBqh/m3

ERR=−0.2 (95% CI: −0.5, 0.06)

France, CEACOGEMA [R5]185

MiningkBqh/m3

ERR=0.016 (95% CI: −0.06, 0.13)

France, CEACOGEMA [D24]76

MiningkBqh/m3

HR=1.13 (95% CI: 0.97, 1.31)

France, AREVA NC
[G33]

111

Processing

Reprocesseduranium,
absorption Types
M, S

HR (95% CI)
Cumulative exposure duration(per year):
1.09 (1.02, 1.18) [M]
1.11 (1.03, 1.20) [S]
1.04 (1.00, 1.07) for naturaluranium [S]High cumulative exposure:
3.40 (1.47, 7.85) [M]
8.79 (1.21, 28) [S]2.84 (1.38, 5.85) for naturaluranium [S]Cumulative exposure score:
1.14 (1.05, 1.24) [M]
1.17 (1.07, 1.27) [S]
1.07 (1.02, 1.13) for naturaluranium [S]France, Gaseous
diffusion [Z10]281

Processing

Medium and highvs. no exposure

Natural soluble uranium
compounds RR (95% CI):
Medium RR=0.98 (0.71, 1.3)
High RR=1.2 (0.85, 1.6)
Enriched uranium (n=45):
Medium RR=0.96 (0.32, 2.9)
High RR=0.84 (0.28, 2.8)
Depleted uranium (n=45):
Medium RR=0.64 (0.23, 1.7)
High RR=0.84 (0.32, 2.3)

HEART DISEASE
Germany, Wismut[K14]

3 719 (all heartdisease)Mining100 kBqh/m3

ERR=−0.3 (95% CI: −0.6, 0.02)

France, CEACOGEMA [R5]72 (Ischaemic
heart disease)

MiningkBqh/m3

ERR=−0.029 (95% n.e. (<0, 0.14)b

France, CEACOGEMA [D24]26 (Ischaemic
heart disease)

MiningkBqh/m3

HR=0.94 (95% CI: 0.73, 1.20)

Germany [K18]341 (Ischaemic
heart disease)

Milling

100 kBqh/m3

ERR=–0.09 (95% CI: –0.84, 0.65)

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UNSCEAR 2016 REPORT

Study references

No. of deaths

France, AREVA NC[G33]

48 (Ischaemic
heart disease)

Nature ofuranium work

Unit of uranium(LLR) exposure

Risk estimate
per unit LLR exposure a

Processing

Reprocesseduranium,
absorption Types
M, S

HR (95% CI)
Cumulative exposure duration(per year)
1.08 (0.97, 1.21) [M]
1.14 (1.03, 1.26) [S]
1.04 (0.99, 1.10) for naturaluranium [S]High cumulative exposure:2.05 (0.53, 7.85) [M]
4.38 (0.47, 41) [S]2.57 (0.82, 8.07) for naturaluranium [S]Cumulative exposure score:
1.12 (0.99, 1.27) [M]
1.17 (1.03, 1.33) [S]
1.13 (1.05, 1.22) for naturaluranium [S]France, Gaseous
diffusion [Z10]95 (Ischaemic
heart disease)

Processing

Medium and highvs. no exposure

Natural soluble uranium
compounds (with 95% CI):
Medium RR=0.71 (0.39, 1.3)
High RR=0.91 (0.53, 1.5)

CEREBROVASCULAR DISEASE
Germany, Wismut
[K14]

1 297

Mining100 kBqh/m3

ERR=−0.05 (95% CI: −0.5, 0.6)

France, CEACOGEMA [R5]41

MiningkBqh/m3

ERR=0.125 (95% CI: −0.06, 0.50)

France, CEACOGEMA [D24]16cMiningkBqh/m3

HR=1.17 (95% CI: 0.90, 1.53)

Germany [K18]171

Milling

100 kBqh/m3

ERR=–0.17 (95% CI: –1.14, 0.80)

France, AREVA NC
[G33]

31

Processing

Reprocesseduranium,
absorption Types
M, S

HR (95% CI)
Cumulative exposure duration(per year)
1.09 (0.93, 1.27) [M]
1.11 (0.95, 1.29) [S]
1.04 (0.97, 1.11) for naturaluranium [S]High cumulative exposure:5.71 (1.48, 22) [M]
3.26 (0.97, 11.0) for naturaluranium [S]Cumulative exposure score:
1.13 (0.97, 1.31) [M]
1.16 (1.00, 1.35) [S]
1.01 (0.92, 1.12) for naturaluranium [S]ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 431

Study references

No. of deaths

France, Gaseous
diffusion [Z10]77

Nature ofuranium work

Unit of uranium(LLR) exposure

Risk estimate
per unit LLR exposure a

Processing

Medium and highNatural soluble uranium
compounds (with 95% CI):
Medium RR=1.2 (0.66, 2.3)
High RR=1.07 (CI: 0.6, 1.9)aRisk estimation metrics: ERR (excess relative risk) for which zero represents no excess or deficit; HR (hazard ratio) and RR(relative risk or rate ratio) are expressed as a multiple of the rate in the baseline (lowest/no exposure) group. (ERR = RR – 1).
bn.e. = not estimable from the likelihood profile.

c

A case-control subset of workers for whom information was available on radon and external gamma exposure and medical risk

factors. Risk estimates were adjusted for those factors.

279. Studies of CSD/IHD/CeVD end points typically had greater statistical power than those for mostcancer diseases because the numbers of CSD-related deaths were much larger than for most types ofcancer. Other considerations arise; one might expect lesser statistical power to detect CSD because theconsensus has been that the risk coefficients derived for external radiation of the circulatory system areseveral times smaller than for cancer induction. On the other hand, damage to the kidney, which isthought to be a primary target organ for uranium, affects the risk of heart disease, probably through therenin-angiotensin pathway. Results regarding CSD end points are also difficult to interpret because of
the numerous medical and lifestyle factors that affect cardiovascular risk. The results of the French
processing workers (AREVA NC) suggesting that uranium compounds with low solubility may induce
CSD more than soluble compounds are somewhat puzzling. While less soluble uranium compounds arethought to confer more risk to the lung because of longer residence times, it is believed that moresoluble compounds confer larger doses to most other organs than insoluble compounds do because ofdifferences in biokinetics. If so, the results of the French processing workers (AREVA NC) are contrary
to what one would expect for CSD, IHD and CeVD.

(g) Conclusion
280. From the occupational exposure studies, a weak association of lung cancer risk with uranium
exposure is concluded. However, currently available results are not consistent enough to demonstrate a
causal association with uranium exposure. The results for leukaemia, other lympho-haematopoietic
malignancies, digestive system cancers, kidney and other urological cancers and brain/CNS tumours
did not provide clear evidence of uranium exposure-related risks. The results for non-malignant
diseases—respiratory, cardiovascular and kidney diseases—also showed no relationship with uranium
exposure. A number of studies without dose-response analyses for LLR exposure have provided nulloverall risk estimates for every health end point considered; while the negative SMRs are not very
specific, at least they suggest the risks are likely to be small.

5.

Limitations of occupational exposure studies

281. For uranium miners studies, recent developments enabled the calculation of organ doses and,
therefore, estimation of the contribution of LLR from uranium ore. Studies published up to now havedemonstrated a very small contribution of uranium to miner dose. Organ doses appear to be dominatedby radon and radon decay products for lungs and by external gamma exposure for other organs.
Demonstrating a potential risk associated to uranium, therefore, appears difficult.

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UNSCEAR 2016 REPORT

282. Studies of uranium millers are limited in size, with the exception of the German Wismut miller
study, which allowed assessment of individual radiation doses from uranium exposure [K18].
Undertaking analysis of combined cohorts after further development of organ dose calculation should
improve results in the future.
283. Most studies of workers in the nuclear fuel cycle are limited by the difficulties in estimating thedoses due to uranium internal contamination. Recent studies are the most meaningful since they havebeen based on more accurate exposure assessment and for some of them internal organ-specificabsorbed doses were estimated by implementing the latest updates of the ICRP models and dosimetric
tools. Nevertheless, improvement of dosimetric estimation is still needed to provide pertinent estimatesof potential risks associated with uranium contamination. Segmentation between the different steps ofthe nuclear fuel cycle should enable improved assumptions regarding the solubility of the uranium
compounds. The combined use of individual urinalysis dosimetry and of job exposure matrices may
also allow improved characterization of radiation doses from uranium exposure and the quantificationof related risks.
284. Natural uranium is not very radioactive (238U decays very slowly) and its chemical properties areoften such that any inhaled or ingested uranium is excreted rather quickly from the human body. Thus,
studies of exposure to enriched or reprocessed uranium may be more informative. Some publications
indicated that—even when the uranium doses are known—external exposure can dominate [L24]. In
addition, most of these studies had major limitations (poor statistical power, no or imprecise estimatesof doses, insufficient accounting for other exposure influences). Moreover, other exposure is mostly nottaken into account, such as exposure to chemicals, heat and noise, which may also contribute to theincrease of certain diseases.
285. Uranium worker data have often been limited to studies of male Caucasians. Quantitative
generalization to women or other population groups is therefore uncertain. No occupational studies
have attempted to examine genetic, epigenetic or metabolic susceptibility factors for uranium relateddiseases. Worker studies also provide no information about children, who may be more susceptible tothe effects of uranium exposure than adults.
286. Continued follow-up of the principal uranium worker cohorts that have individualized worker
exposure data will be valuable. Many members of the cohorts were relatively young at their most recent
follow-up. Because mortality rates for many cancers increase as a power of age attained, these cohorts
will become increasingly informative with future follow-up, providing greater ability to detect smaller
effects and generate more precise risk estimates.
287. Since most studies have a limited number of uranium workers and relatively slight exposure formost of the workers, it is unlikely that epidemiological studies of individual nuclear facilities will havesufficient statistical power for a reasonable prospect to detect risks. Consequently, international pooled
studies with high-quality, harmonized individual exposure estimates are likely to be necessary to assessuranium risks with high precision. However, studies of workers have advantages over those on
environmental exposure of the public or of special groups (e.g. military personnel deployed in regionswith potential DU exposure). The advantages particularly centre around having measurements toestimate individual uranium exposure levels, along with other occupational radiological exposure sothey can potentially distinguish LLR exposure from other exposures. In addition, some of the studies
had data on various chemical exposure at workplaces.

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 433

B. Studies of Gulf War veterans288. Depleted-uranium munition and armour were extensively used by the United States military in thefirst Gulf War in Iraq and Kuwait (Desert Storm) and again in the Balkans military action. Military
personnel were exposed to DU via inhalation or wounds, notably due to friendly-fire incidents burning
depleted-uranium-containing tanks and ammunition and clean-up operations [B26].
289. Several authors reported investigations on the Gulf war and the Balkan veterans, especially on
United States, United Kingdom, Canadian, Danish and Dutch veterans [B28, M1, M2, S36, U11],
which were reviewed by others [I20, L5, S6]. For example, the Institute of Medicine [I20] discussed
extensively the results of about 25 studies on health outcomes following exposure to natural uranium
and DU. The review integrated malignant (lung cancer, leukaemia, Hodgkin’s and non-Hodgkin’s
lymphoma, bone cancer, renal cancer, bladder cancer, brain and other nervous system cancers, stomach
cancer, prostatic cancer, testicular cancer) and non-malignant (renal disease, respiratory disease,neurological effects, reproductive and developmental effects, cardiovascular effects, genotoxicity,
haematological effects, immunological effects, and skeletal effects) pathologies. They concluded that
there was insufficient evidence to determine whether an association exists between exposure to uranium
and the health outcomes cited above. However, the two following major limitations of the veteranstudies were identified: (a) short period of follow-up and (b) poor assessment of uranium exposure.
290. The Royal Society comprehensively reviewed the use of DU, especially on the battlefield [R28].
The Royal Society concluded in its report that doses from DU were unlikely to be high, even in themost unfavourable (battlefield) conditions, so that lung cancer risks were unlikely to be more thandoubled. The report indicated a potential non-radiological risk associated with exposure to DU, in
particular with its nephrotoxicity. A summary of studies of the health status of veterans with potential
or known exposure to DU is given in appendix A, table A5.
291. A recent study by Strand et al. aimed to investigate cancer incidence and also all-cause mortality
in a cohort of Norwegian military present in Kosovo between 1999 and 2011 [S42]. Cancer incidence
and mortality were studied from 1999 to 2011 and compared to national rates. The authors found no
excess incidence of cancer except an elevated SIR for melanoma of the skin in men. All-cause mortality
was half the expected rate (SMR=0.49; 95% CI: 0.35, 0.67).
292. A biennial health surveillance programme established for the United States Gulf War veterans has
shown continuously elevated DU concentrations in urine among those with embedded fragments forover 20 years [M25]. No differences have been seen between the high- and low-exposure groups withregard to haematology, clinical chemistries, neuroendocrine parameters, bone metabolism,
neurocognitive function, immune function, pulmonary function or nodules. Regarding renal function
and injury, no high vs. low exposure differences were found for 16 clinical indicators of renal function,
six urine markers for kidney injury, or four urine measures of low molecular weight proteins, except for
two sensitive biomarkers of proximal tubule function that suggested subtle renal injury [M25].
293. In a study by Hines et al., some self-reported respiratory symptoms, mean pulmonary function
values and prevalence of low-dose chest computed tomography abnormalities were compared in two
populations of Gulf War veterans (high body burden group vs. low body burden) [H18]. The authors
found no significant differences between the two groups, suggesting that DU levels inhaled during the1991 fire incidents probably do not cause long-term adverse pulmonary health effects [H18].
294. In conclusion, several studies on the health pathologies among veterans with potential or no
exposure to DU were published. Up till now, no clinically significant pathology related to DU has beenfound in the veteran’s cohorts. The diversity of these studies in terms of topics has limited their

434

UNSCEAR 2016 REPORT

reproducibility, except for the biennial examinations of a small group of United States veterans withretained DU shrapnel in whom comprehensive examinations have consistently found no clinically
adverse effects.

C. Studies of environmental exposure
1.

Living around uranium processing facilities

295. Numerous, mostly ecological, studies have been carried out to assess whether long-term residence
in the vicinity of nuclear fuel cycle facilities or nuclear power plants affects the health of the residents.
To focus on uranium effects, only studies carried out in population groups living around uranium
processing facilities (after uranium mining and prior to electricity production) were examined. Elevenpublished studies were identified over the past ten years, which are presented in appendix A, table A6.
296. Because of potential bias, inability to check the validity of ecological results, and the lack ofsufficient measurements of ambient uranium exposure levels, no firm conclusion could be drawn from
ecological studies. Further, caution is required in interpreting ecological studies in general as causal
inference is not warranted because of numerous limitations in their study design. The major limitationof ecological studies is the potential of ecological associations to misrepresent, sometimes greatly, thebiological effect at individual level. Thus, an association observed between variables on an aggregate
level does not necessarily mean that the same association will exist at individual level [G19].
297. Lane et al. reported a review of 13 epidemiological studies conducted in Port Hope, Canada in thepast 30 years, including residents and workers [L7]. These studies included environmental
measurements of the radiological and non-radiological contaminants, the estimation of the multipathway of exposure and also the health risks to the population, using environmental monitoring data or
dose reconstruction methods based on a variety of approaches. The authors concluded that, taken
together, the findings of these studies conducted on the Port Hope community indicated that observed
adverse health effects were unlikely to be the result of exposure to environmental contaminants fromradium and uranium processing. Other studies shown in appendix A, table A6 are also consonant withthat conclusion.

2.

Living in an environment affected by depleted uranium munition use

298. Some epidemiological studies attempted to determine if the health of populations living incountries or regions involved in the recent conflicts (i.e. Iraq, Kuwait, Bosnia and Herzegovina,
Kosovo, Serbia and Montenegro) was affected by the use of DU in shrapnel or tanks. Iraq is the moststudied country for investigation on possible effects of DU. Several publications aimed at describing
the incidence and types of congenital malformations [A6, A7, S48] or birth defects [A3, B54, F1].
However, these publications failed to demonstrate a link between the increase in these pathologies andthe environmental exposure to DU, notably due to the absence of evaluation of the exposure levels.
299. Few studies investigated cancer incidence in these populations. Al-Hashimi and Wang used intheir study three sub-periods (1980–1990, 1991–2000, and 2001–2010), corresponding to the three Iraq
wars, the Iran–Iraq war (1980–1988), the Gulf War I in 1991 and the Gulf War II in 2003 [A4]. The
authors reported increases in the total number of cancer cases. However, the in-depth analysis indicatedANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 435

a decrease in incidence rates in most cancer types when they were analysed statistically, considering
population growth in the Ninawa province in the northern part of Iraq.
300. Another study aimed to describe changes in haematological malignancies (leukaemia andHodgkin’s lymphoma) in Croatian counties potentially exposed to DU in comparison to the pre-war
period [L3]. This study did not find a significant difference in the incidence of these haematologicalmalignancies.
301. In parallel to these health studies, measurements were made of daily urinary uranium excretion inGerman peacekeeping personnel (n=1,228) and unexposed subjects coming from the South of Germany(n=113) to assess potential intakes of DU [O1]. A daily urinary excretion of uranium of
13.92.2 ng/day (3 to 23 ng/day) measured for German peacekeeping personnel was similar to that of
unexposed subjects (12.82.6 ng/day).

3.

Drinking water with elevated uranium levels

302. Possible health effects after long-term ingestion of uranium via drinking water was reviewed by
Guseva Canu et al. [G32]. The description and main results for selected studies of the possible impactof elevated levels of uranium in drinking water are summarized in appendix A, table A7. This table
notes potential uranium effects ascribable to its dual radiological and chemical toxicity. However, somestudies related the effects to chemical toxicity only [K26, K27, K28, M9, S8, Z6], while other studies
related to potential radiation effects [A32, K29, S7].
303. As shown in appendix A, table A7, most of the studies focused on the nephrotoxicity of uranium
using cross-sectional study designs. In total, five studies were carried out: in Canada [M9, Z6, Z8]; inFinland [K26]; and in Sweden [S8]. The uranium concentrations in water were fairly similar in all thestudies, with median concentrations in the range 20–30 µg/L among the exposed groups. All thesestudies, except the one from Mao et al. [M9], found no glomerular effect of chronic ingestion of
uranium. Among people drinking water from private drilled wells, uranium exposure caused damage tothe proximal tubule, shown by nephron reabsorption alteration [K26, S8, Z6] or tubular cytotoxicity[S8, Z1] was observed in four of the studies. Several biomarkers were measured in these studies (e.g.
creatinine for glomerular filtration function at the early stage of renal injury) but none was specific for
injury caused by uranium. Kurttio et al. [K29] carried out a case–cohort study in Finland of bladder andkidney cancer after long-term consumption of private well water containing uranium and its decay
products. No association between the prevalence of these cancers and the uranium concentration in well
water was found [K29].
304. Lymphatic and haematopoietic malignancies were considered in three studies [A32, S7, W23].
Seiler [S7] investigated whether 16 children with leukaemia in the City of Fallon, Nevada, UnitedStates, had higher levels of naturally occurring radioactive material in their well water compared toother inhabitants of the town. Water samples were collected in 2001 for the measurement of uranium,
radon and gross alpha concentrations, and leukaemia cases were identified for 1997–2000. To resolve
this potential time sequence problem, the authors also retrieved the 1989 citywide water analyses. The
natural origin of the uranium present was confirmed by the calculation of the isotopic ratio. No
difference was indicated in uranium concentration in the water drunk by the children compared to otherinhabitants.

436

UNSCEAR 2016 REPORT305. Witmans et al. [W23] compared the uranium concentrations in water between non-Hodgkin’s
lymphoma cases and their matched controls selected from the Saskatchewan (Canada) cancer registry.
The cases had been exposed to significantly higher uranium concentrations in drinking water than thecontrols. However, uranium was one of 63 inorganic constituents tested in the study.
306. A case–cohort study by Auvinen et al. [A32] of Finnish adults that enrolled 35 cases of leukaemia
also reported a negative result regarding exposure to naturally occurring uranium (and its decay
products) in drinking water and leukaemia [A32]. The statistical power of the study was limited and nodata on potential confounding factors were available. The risk of stomach cancer from exposure tonaturally occurring radionuclides in drinking water was investigated. However, no association was
found in this study.
307. Clinical studies in Nova Scotia, Canada performed on 324 persons exposed to variable amounts ofnaturally occurring uranium in drinking water (up to 0.7 mg/L) found no relationship with overt renal
disease. Though there was a trend towards increasing excretion of urinary -2 microglobulin withincreasing concentration of uranium in well water, this was not seen in the group with the highest
uranium well-water concentrations. This group had significantly reduced its consumption of well waterby the time the measurements were made, supporting the hypothesis that the suspected tubular defectmight well be rapidly reversible [M64, M65].
308. A pilot study by Mao et al. of three communities in Saskatchewan with mean uranium levels
ranging from 0.71 (control) to 19.6 g/L found a statistically significant association (p=0.03) between
increasing but normal levels of urine albumin and the uranium exposure [M9]. Another Canadian study
on two groups of subjects with chronic exposure to uranium in drinking water, the first group exposed
to <1 g/L and the other exposed to 2–781 g/L found no correlation with alkaline phosphatize and-2 microglobulin in urine. The authors concluded that the uranium concentrations observed in the
study affected the kidney function at the proximal tubule [Z6].
309. Another study by Zamora et al. [Z8] on chronic ingestion of uranium in drinking water
demonstrated subtle changes in two of the indicators measured that were statistically significant—namely, glucose and LDH excretion concentrations. However, this did not result in any observable
health effects (see also table 13). In addition, the change in LDH excretion was rather beneficial andwas seen only in males. Thus, these changes are not nephrotoxic effects.
310. In conclusion, epidemiological studies of public uranium exposure to drinking water indicate thatchemical toxicity of uranium may occur mainly in the kidneys and, in high concentrations, uranium
may affect the kidney function. However, the functional alterations found in the kidneys were small andwithin normal limits, so the clinical significance of the findings may be minimal. The available
literature focused on lymphatic and haematopoietic tissue malignancies is limited to three studies,
which do not support a causal association between uranium exposure and those malignancies.

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 437

IX. RESEARCH NEEDS
311. The estimation of organ doses from incorporated uranium isotopes depends on the availability of
reliable biokinetic data and the construction of physiologically realistic biokinetic models. In general,
good human and animal data are available for the construction of models. However, limited information
is available on the age-dependence of organ retention and excretion rates, including information on the
cross-placental transfer of uranium. In addition, more information is required on the distribution of
uranium within tissues and cells, for example in CNS tissue and lungs.
312. Dosimetric models in general apply the same assumptions regarding source and target
distributions within tissues to all internal emitters. For example, uranium isotopes and otherradionuclides deposited in bone are assumed to accumulate on internal bone surfaces and/or in bonevolume and target cells for cancer induction are assumed to reside along bone surfaces (bone cancer)
and throughout red bone marrow (leukaemia). The validity of such assumptions requires further
investigation, with consideration of the inhomogeneity of uranium distribution within tissues and cells.
313. Toxicological studies of uranium exposure are required to distinguish the chemical andradiological components of damage caused to cells and tissues, including short-term damage to organfunction and longer-term effects including cancer. Comparisons of radionuclide toxicity and RBE
determinations would assist in quantifying the potential health effects of uranium isotopes. Studies of
the age-dependence of chemical and radiological toxicity would be valuable.
314. Future epidemiological studies require careful consideration of the acquisition of dosimetric datato assess individual organ and tissue doses for cohort subjects. A high priority would be a consortium
effort by investigators to develop pooled data on uranium risks. The result could be considerable gains
in the statistical power and precision of risk estimates that would potentially provide the best overallanswers achievable as to health effects from uranium exposure. Pilot studies to quantify the magnitudeof uncertainties in exposure assessment would ideally be part of this effort, so that sound estimates of
dosimetric uncertainties could be incorporated into the risk modelling [L13, S21].
315. Concerning (molecular) epidemiology, setting up prospective follow-up or case control studies inselected subgroups, including collection of information on biomarkers, has the potential to provide
more specific dose–response curves for defined subsets of cohorts and thereby improve knowledge ofhealth effects in humans, including cancer and non-cancer diseases. High-throughput technologies(especially the -omics) would be relevant to apply to this field. However, proposed biomarkers willneed to be rigorously evaluated as to their ability to improve exposure and risk assessment.
316. Mixed exposure should be taken into account when studying effects, such as other radionuclides(e.g. 239Pu, 222 Rn), other chemical carcinogens (e.g. solvents, smoking, dust, silica, asbestos) and also
the physical forms of uranium (solubility), e.g. through further development of exposure matrices in
epidemiological research, and through animal studies.
317. Understanding the molecular mechanism of action of uranium on cells in culture and animalmodels, both as a metal and as a radionuclide, would be important in (a) facilitating the identification of
bioindicators; (b) identifying portions of the molecular response that are attributable to the radiation
response, the heavy metal response or both; and (c) defining the possible development of mitigators,(few mitigators for uranium exposure).

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UNSCEAR 2016 REPORT

X. GENERAL CONCLUSION
318. This annex provides a detailed review of sources and levels of uranium in the environment,
exposure of the public and workers to uranium, biological effects of uranium, and epidemiologicalstudies of nuclear workers and the public exposed to uranium.
319. Uranium is a naturally occurring radionuclide and is ubiquitously distributed in the environment.
In daily life, people are exposed to uranium originating mainly from drinking water and foodstuffs.
Average uranium levels in water vary between countries and within countries, with typical values ofaround 2 µg/L (~25 mBq/L of 238U) in groundwater and 1 µg/L (~12.4 mBq/L of 238U) in public watersupplies. Some drinking water samples (<3%) may exceed the national or international guidelines set toprevent kidney toxicity. Concerning foodstuffs, potatoes, meat, fresh fish and bakery products are themain sources of uranium ingestion. The total daily intake from water and food consumption is around
1.5 μg/d (18.6 mBq/d of 238U).
320. The main routes of entry of uranium into the body are inhalation and ingestion. The absorption toblood in each case is highly dependent on the chemical form (speciation) of the intake. For example,
human data show that the absorption of ingested uranium is a few per cent of intake for soluble forms
in water compared with substantially less than 1% for insoluble oxides. Human and animal data havebeen used to model the behaviour of uranium absorbed to blood, showing that the main site of retention
is the skeleton, with lower amounts in soft tissues and rapid urinary excretion of a large proportion. The
ICRP models make appropriate use of the available data.
321. Uranium is both a radioelement and a metal, and biological effects may result from the combined
effects of the chemical element or species and the radiation. The radiological and chemicalconsequences of internal exposure to uranium depend partly on the route of intake (principally
inhalation or ingestion), and the chemical form of the intake. Some effects are likely to be related to thechemical toxicity of uranium species, namely the renal effects, whereas others are rather related toradiological toxicity of uranium, including tumorigenic effects such as soft tissue sarcomas in rats andosteosarcoma in mice. In general, chemical effects are observed with short lag-times after exposure
whereas radiological effects such as carcinogenesis have long lag-times.
322. Considering the chemical effects of uranium species, the kidneys are the most sensitive target
organ. At higher levels, chemical effects of uranium are also observed in bones, indicating that uranium
can induce effects on bone metabolism such as the impairment of bone growth and formation. Chemical
effects of uranium have also been observed, in rodent studies, in liver, gonads, central nervous system,and the immune system. These experimental studies indicate that uranium induces biological effects inthese organs, but the changes do not lead to the appearance of observable pathologies. While effects inthese tissues may be seen at higher doses, damage to kidneys (and skeleton) is likely to be critical.
Concerning the central nervous system, animal studies suggest that high doses of uranium may havesome negative effects on the behaviour of animals. With the exception of kidney damage, animalstudies showing toxicological effects have used concentrations of uranium substantially above those towhich humans are exposed. No clinically significant pathologies have been found in the veteran cohortspotentially exposed to DU. Moreover, the biennial examinations of a group of United States veteranswith retained DU shrapnel have found no clinically meaningful adverse effects.
323. Epidemiological studies of uranium miners and millers have included estimates of doses, showing
the small contribution of uranium to overall doses and the dominant contributions of radon and radon
decay products to lung dose and external gamma radiation for other organs. Most studies of nuclear
workers are limited by difficulties in estimating radiation doses due to uranium. A weak association oflung cancer risk with uranium exposure is suggested but the currently available results are notANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 439

consistent enough to demonstrate a causal association. Results for other malignancies and nonmalignant disease were also negative. The Committee concluded that epidemiological studies of public
exposure to uranium in drinking water have reported small functional alterations in the kidneys, within
normal limits and hence of minimal clinical significance.

XI. ACKNOWLEDGEMENTS
The Committee wishes to acknowledge with gratitude the help of the experts E. Ansoborlo (France),
Y. Gueguen (France), I. Guseva-Canu (France), D. Laurier (France), P. Lestaevel (France), F. Paquet(France), E. Rage (France), J. Harrison (United Kingdom) and R. Shore (United States), who weredirectly involved in preparing the evaluation, and S. Baatout (Belgium) for critically reviewing themanuscript. In particular, the Committee would like to thank I. Dublineau (consultant), who mainly
elaborated this scientific annex based on a first draft prepared by K. Bundy (Canada), D. Chambers(Canada), R. Lane (Canada) and B. Thériault (Canada). The views expressed in this scientific annex
remain those of the Committee and do not necessarily represent the views of individual experts, theUnited Nations or its Member States.

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 441

APPENDIX A. TABLES SUMMARIZING URANIUM LEVELS IN
WATER, STUDIES OF OCCUPATIONAL AND ENVIRONMENTALEXPOSURE TO URANIUM
Tables A1 to A4 present uranium levels in water.
Tables A5 to A8 present four groups of occupational studies that provide increasing levels of
information on uranium-specific risks.
Tables A9 to A11 present studies of risks from various sources of potential military and environmental
exposure to uranium.

Table A1.

Overview of uranium content in groundwater worldwide

The numbers in italics correspond to calculated data obtained from the mass activity of natural uranium of 25,400 Bq/g and from the relative proportion of 238U (12.4 mBq/μg) in natural uranium of 48.3%; * meanvalue (not median value); n.i: information not included in the study

Continent
EuropeSouth America

Location

Sample number

Total uranium (μg/L)

238

Reference

U (mBq/L)

Finlandn.i.

396

0.15 (<0.01–10)

1.84 (0.12–123)[T22]

FinlandSouth

13

285 (5.6–3 410)

3 504 (143–41 924)

[P28]

Finland

South

288

1.5 (0.3–800)

19 (4–9 990)

[V4]

Finland

South

194

28 (0.001–1 920)

344 (0.01–23 605)

[K26]

Finland

South

167

2 (<0.01–1 770)

24.6 (<0.12–21 761)

[M67]

SwedenVärmland153

6.7 (<0.20–470)

82.4 (<246–5 778)

[S8]

SwedenSouth

328

14.2 (<2–425)

177 (<27–5,293)[I22]

NorwaySouth476

2.5 (−750)

31 (−9 300)

[F12]PolandSwieradow51

1.26 (0.2–2.4)

15.55 (2.4–29.4)

[K10]

Spain

Catalonia

37

4.88 (<0.41–56.1)

60 (<5–690)[O8]

Switzerland

n.i.

5 548

0.77 (0.05–92.02)

9.5 (0.61–1 131)

[S33]

FranceVals les Bainsn.i.

1.85 (0.55–3.6)

22.7 (6.8–44.3)[M34]

GreeceNorth

10

2.02 (0.15–7.66)

25.06 (1.82–95.3)[S1]

Canadan.i.

n.i.

0.31 (0.16–6.23)

3.8 (2.0–76)[L12]CanadaKitigan Zibi, Quebec

32

39.25 (0.4–845.33)

470 (4.9–10 392)

[Z8]

CanadaSoutheastern Manitoba

287

10 (<0.02–2 020)

124 (<0.25–25 048)

[B21]

CanadaNova Scotia

20

0.39 (0.06–41.08)

4.8 (0.8–505)

[K19]

USA

Connecticut

11

16.3 (0.21–1 166)

200 (2.6–14 335)

[M5]

USA

Connecticut

35

157 (1.8–7 780)

1 930 (22.1–95 649)

[O5]

USA

Cities with the largest
average concentration

55 4331.04 (0.03–1 945)

12.95 (0.37–24 124)

[C34]

Brazil

Sao Paulo et Santa Catarina

78

0.28 (0.008–15.0)

3.4 (0.1–184.3)[B43]

Braziln.i.

358

1.2* (<0.01–7.5)

14.8 (0.12–92.2)[G14]

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 443

North America

Country

Country

Asia

Table A2.

Sample number

Total uranium (μg/L)

238

Reference

U (mBq/L)

Moroccon.i.

15

4.5 (0.37–25.1)

55 (4.5–309)[H4]

Ethiopia

Rift Valley138

0.55 (0.005–48)

6.8 (0.06–590)

[R10]

Egypt

Eastern desert

12

5.21 (1.19–519.4)

64.1 (14.6–6 386)

[D2]

Ghana

North and coast

195

0.114 (<0.001–1.99)

1.4 (<0.01–24.4)[R23, Z16]

Fujian Province

China

110 and 552

0.54 (0.03–13.4)

6.6 (0.4–164)

[Z16]

Iran

Caspian Sea

27

2.2 (0.24–5.4)

27 (3–66)[J5]

Jordann.i.

n.i.

1.3 (0.5–6.7)

16.0 (6.6–82.4)[V3]

Bangladesh

West

67

2.5 (0.2–10)

30.7 (2.5–25.8)

[F14]

Japan

Niigata

230.001 (0.0005–0.03)

0.018 (0.005–0.383)

[T14]

India

Hisar

38

33.9* (5.3–113.5)

417 (65.2–1 395)

[G3]

India

Kula area

15

0.83 (0.26–2.56)

10.2 (3.2–31.5)

[S27]

India

Punjab25

22 (2.65–74.98)

271 (32.6–922)[K22]

Overview of uranium content in surface water worldwide

The numbers in italics correspond to calculated data obtained from the mass activity of natural uranium of 25,400 Bq/g and from the relative proportion of 238U (12.4 mBq/μg) in natural uranium of 48.3%;
n.i: information not included in the study

Continent
EuropeCountry

Location

Sample number

Total uranium (μg/L)

238

U (mBq/L)

Reference

Finland

Southern Finland

184

0.18 (0.08–34)

2 (1.0–420)[V4]

Finlandn.i.

152

0.099 (<0.01–0.92)

1.2 (<0.12–11.3)[T22]

North America

USA

Cities with the largest average
concentration

34.5610.45 (0.03–1 737)5.55 (0.37–21 534)

[C34]South America

Argentina

n.i.

92

1.9 (<0.01–50)

23.4 (<0.12–615)

[B41]

Asia

India

Upper Siwaliks and Punjab

34

3.84 (1.08–19.68)

47.2 (13.3–242)

[S23]

Iran

Ardabil22

4.2 (2.1–13.6)

51.6 (25.8–167)[H2]

UNSCEAR 2016 REPORT

AfricaLocation

444

Continent

Table A3.

Overview of uranium content in public water supplies worldwide

The numbers in italics correspond to calculated data obtained from the mass activity of natural uranium of 25,400 Bq/g and from the relative proportion of 238U (12.4 mBq/μg) in natural uranium of 48.3%; * meanvalue (not median); n.i: information not included in the study

Continent
EuropeCountry

Location

Sample number

Total uranium (μg/L)

238

U (mBq/L)

Reference

Southern Finland951

1.25 (<0.01–1 770)

15.4 (<0.12–21 761)

[M67]

Germany

Bavaria461

0.9 (<0.01–39)

11.1 (<0.12–479)[R3]

Germany

National scale564

3.2 (<0.7–320)

39.3 (8.6–3 934)

[B23]

Germany

National scale

164

0.073 (0.00115–9.0)

0.90 (0.001–111)[B24]

Germany

National scale36

2.0 (<0.16–60.2)

24 (<2–740)[G2]

Austrian.i.

41

0.91 (0.06–79.2)

11.2 (0.72–975)[G6]

Austria

Waldviertel

48

1.1 (0.1–57.5)

14.05 (0.7–707)[W5]

GreeceAksios, Kalikratia

23

3.46 (0.061–10.02)

42.5 (0.75–123)[K7]

PolandCentre26

0.39 (0.03–1.94)

4.8 (0.4–23.9)

[P20]

Spain

Biscay40.07 (0.003–0.24)

0.8* (0.04–2.9)

[H17]

Italy

Rome

9

1.46 (0.02–8.37)

18 (0.3–103)[J2]

Norway, Sweden, Finland, Iceland

Several places22

0.107 (0.0049–56.2)

1.32 (0.06–691)[F13]North America

USA

Illinois, Minnesota, Texas, Wisconsin

24

0.25 (0.1–13.2)

3.1 (1.23–162)[L50]

South America

Argentina

12 provinces

145

0.4 (<0,01–21)

4.9 (<0.12–258)[B41]

AfricaMoroccoNational scale

6

0.75 (0.20–1.28)

9.25 (2.5–15.7)

[H4]

Asia

Islamic Rep. of IranArdabil3

7.6 (4.7–11.7)

93.4 (57.8–144)

[H2]

India

Punjab

3.84 (1.11–19.68)

47.2 (13.6–242)

[S22]

India

Punjab

10.4 (1.24–45.42)

260 (30–1 150)

[R8]

India

Punjab29 (2.5–313)

357 (30.7–3 848)

[S11]

India

Himachal Pradesh

46

1.34 (0.56–10.11)30 (10–260)

[R8]

India

Western Haryana

23

17.03 (6.37–43.31)

209.4 (78.3–532)[K1]

Australia

West

23

0.19 (0–1.16)

2.3* (0–14.3)[W7]

Australia

West

173

0.06* (<0.001–1.40)

0.74 (<0.01–17.2)[C1]

Oceania45

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 445

Finland

446

Overview of uranium content in bottled mineral water worldwide

The numbers in italics correspond to calculated data obtained from the mass activity of natural uranium of 25,400 Bq/g and from the relative proportion of 238U (12.4 mBq/μg) in natural uranium of 48.3%; * meanvalue (not median); n.i: information not included in the study

Continent
EuropeCountry

Sample number

Total uranium (μg/L)

238

U (mBq/L)

Reference

Germany908

0.17 (<0.0005–16.0)

2.09 (<0.006–197)[B24]

Germany

21

0.41 (<0.08–11.4)

5 (<1–140)

[G2]

Spain

32

0.48 (0.04–5.8)

5.9 (0.5–70.9)[D14]

Slovenia

11

0.42 (0.09–4.3)

5.2 (1.1–53)

[B13]

Italy

21

1.38 (0.20–9.92)

17 (2.5–122)

[R30]

Italy

51

0.73 (<0.01–7.2)

8.97* (<0.17–89)[D11]

Poland220.59 (0.06–0.86)

7.26 (0.75–10.54)[C19]

Croatia120.53 (0.17–1.19)

6.55 (2.1–14.6)[R29]

France, Portugal, Spain

14

3.58 (1.79–40.7)

44 (22–500)[M14]

France106

0.2 (<0.10–19)

2.25 (<1.3–230)[A27]

28 countries

132

0.23 (0.0002–27.5)

2.8 (0.002–338)[K11]

Austria100.16 (0.012–5.4)

2.0 (0.15–66.4)[W4]

Norway, Sweden, Finland, Icelandn.i.

0.102 (0.0055–32.4)

1.25 (0.07–698)[F13]South America

Argentina

62

1.9 (0.04–11)

23.4 (0.5–135)[B40]

AfricaTunisia10

1.01 (0.13–2.14)

12.36 (1.56–26.36)[G7]

Morocco100.54 (0.34–0.70)

6.6 (4.2–8.6)[M48]

Kuwait23

0.22 (0.05–2.04)

2.74 (0.63–25.07)

[A11]

Asia

UNSCEAR 2016 REPORT

Table A4.

Table A5.

Studies of workforces that include workers (potentially) exposed to uranium though not explicitly identified

Study references

Summary of study

Summary of findings relating to uranium

Relevance for this report

Atkinson et al.
[A29]

UKAEA employees, 1946–1996. 10,249 deaths. Dosimetry:
external radiation doses. Neutron and tritium doses included
when available. Internal doses noted but not quantitative

All cancer: external radiation exposure trend tests in
those monitored for any internal exposure. Trends
not significant for all cancer, stomach, colon, liver,pancreas, lung, bladder, kidney, brain

Not possible to derive uranium-specific risks,because uranium workers not analysed
separately, and uranium-specific doses not used

Prostate cancer: dose-response trend before 1980, but
not 1980–1997 among those with internalmonitoring, but uranium not examined
Further analysis of an extended UKAEA dataset [A29], conducted bytime period to examine internal exposure for pre/post 1980 andprostate cancer. Excess associated with work with heavy-water
reactors. Radionuclides of concern: 3H, 59Fe, 51CR, 60Co, 65Zn. Casecontrol prostate cancer substudy conducted

Prostate cancer: exposure levels at heavy-water
reactors fairly constant over time, but no indicationof elevated risk after 1980. So earlier excess withinternal exposure probably not meaningful

Not possible to derive uranium-specific risks,because uranium workers not analysed
separately, and uranium-specific doses not used

Carpenter et al.
[C8]Cancer mortality 1946–1988 among 75,006 UKAEA, AWE and
BNFL employees. Uranium exposure not assessed

All cancer: analyses of external radiation plus tritium
exposure

Not possible to derive uranium-specific risks,because uranium workers not analysed
separately, and uranium-specific doses not used

Carpenter et al.
[C9]Cancer mortality 1946–1988 among 40,761 UKAEA, AWE or BNFL
employees who had radionuclide monitoring

All cancer: separate analyses conducted for tritium,
plutonium and other radionuclides. Insufficient detailabout other radionuclides, so no uranium analyses

Not possible to derive uranium-specific risks,because uranium workers not analysed
separately, and uranium-specific doses not used

Cragle et al.
[C36]

Savannah River Plant (USA) conducted uranium processing,nuclear fuel fabrication and processing. Follow-up to 1980,
9,860 white male employees. 85% of exposure to externalradiation; exposure to numerous internal radionuclides

Analyses by time of first employment and years of
employment. Suggestion of elevated leukaemia riskin small subgroup of early workers, but uranium
exposure not reported

Not possible to derive uranium-specific risks,because uranium workers not analysed
separately, and uranium-specific doses not used

Fraser et al. [F10]

Cancer mortality and morbidity in UKAEA cohort of 39,718 during
1946–1986. Internal exposure noted for tritium, plutonium orother unspecified radionuclides

Cancer analyses conducted for external exposure,tritium, plutonium and “monitored for any
radionuclide”

Not possible to derive uranium-specific risks,because uranium workers not analysed
separately, and uranium-specific doses not used

Loomis and Wolf
[L44]

Cancer mortality (1947–1990) among 6,591 white males at USA Oak
Ridge Y-12 nuclear material production plant. Plant converted UF6
to UF4 to uranium metal which was fabricated and milled. Other
exposure: beryllium, solvents, machine oils, mercury, lead

No measurements or estimates of uranium exposure,
so no relevant analyses

Not possible to derive uranium-specific risks,because uranium workers not analysed
separately, and uranium-specific doses not used

McGeoghegan
and Binks[M28]

Mortality and cancer morbidity in 2,209 radiation workers at UK
Chapelcross plant, 1955–1995. Main activity: operation of
450 MW Magnox gas cooled reactors

No measurements reported of internal exposure: only
external radiation analysed

Not possible to derive uranium-specific risks,because uranium workers not analysed
separately, and uranium-specific doses not used

McGeoghegan
et al. [M29]38,779 radiation and 15,040 non-radiation male workers at UK
BNFL
facilities(Sellafield,
Springfields,
Capenhurst,
Chapelcross). Some workers had exposure to uranium,
plutonium, tritium and other radionuclides

Investigated circulatory and other non-cancer
diseases. No measurements of internal exposure,analysed external radiation exposure only

Not possible to derive uranium-specific risks,because uranium workers not analysed
separately, and uranium-specific doses not used

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 447

Atkinson et al.
[A30]

448

Abbreviations: LLR, long lived radionuclide exposure, primarily from uranium

Study references
Baysson et al.
[B9]Boice et al. [B36]

Summary of study

Summary of findings relating to uranium

Metallurgy department workers (N=356) of French CEA,
1950–1968 were studied for excess cancers (1950–1990),
since workers believed there was a cancer cluster.
Department research primarily on uranium metallurgy.
Reconstructed external radiation doses; internal
radionuclide and chemical exposure noted. Radionuclides:
thorium, natural uranium/enriched uranium, some
activation and fission products

255 handled radionuclides, principally
uranium, mean exposure duration 11 years

naturalMortality in uranium miners and millers, Grants, New Mexico,
1979–2005: 1,735 underground uranium miners and 904
non-mining uranium millers. No measurements available
on either radon or uranium exposure levels

Increased mortality in underground miners: lungcancer, non-malignant respiratory disease, livercirrhosis. No significant excess among non-mining
millers.Among uranium millers:

All cancer: suggestion of risk (p=0.13) per year
handling radionuclides, but stronger trend forhandling chemicals
Multiple myeloma: suggestive excess, but only 2 cases(0.2 expected, SMR=8.4, 90% CI: 1.4, 26). No evidenceof cancer cluster

Relevance for this report
Study small, low statistical power. Analysis only for
exposure to any radionuclide. Potential confoundingby chemical exposure. Multiple myeloma results:
small numbers, maybe a chance finding frommultiple comparisons. Not possible to deriveuranium-specific risks, because uranium doses notused

One of few studies of uranium millers. Study suggests
uranium exposure effects are small or absent but notpossible to derive uranium-specific risks, because
uranium doses not used

Total cancer: SMR=0.89, n=65
Lung cancer: SMR=0.85, n=21
Cerebrovascular disease: SMR=1.06, n=14Heart disease: SMR=0.84, n=73
Non-malignant respiratory disease: SMR=1.07, n=25
No suggestive excesses of kidney cancer, liver cancer
or lymphoma, but small numbers
Dupree-Ellis etal. [D33]

Mortality (1942–1993) investigated among 2,514 white maleworkers at Mallinckrodt (USA) uranium processing plant.
Mean cumulative total dose, 47.8 mSv. For ~11 years plantalso processed pitchblende, which increased externalradiation exposure

All cancer: SMR=1.05 (95% CI: 0.93, 1.07). Some
evidence of excess kidney cancer (ERR per Sv=10.5,
90% CI: 0.6, 57; n=10) in relation to external radiation
exposure

Only total dose analysed, mainly external radiation.
Not possible to derive uranium-specific risks,because uranium doses not used

Kreuzer et al.
[K16]

Circulatory system disease (CSD) mortality (1946–2008,
n=9,039 CSD deaths) in 58,982 male German Wismut
uranium miners. External radiation estimated using a job
exposure matrix

Circulatory disease: ERR per Sv for external gamma
radiation: −0.13 for CSD, −0.03 for ischaemic heartdisease, and 0.44 (95% CI: −0.16, 0.44) for
cerebrovascular disease

Analysis for external radiation exposure only. Not
possible to derive uranium-specific risks, because
uranium doses not used

UNSCEAR 2016 REPORT

Table A6. Studies of groups of workers identified and investigated as uranium workers but not monitored specifically for potential exposure to uranium, so uraniumspecific doses were not available

Summary of study

Summary of findings relating to uranium

Relevance for this report

Lane et al. [L8]17,660 Canadian Eldorado uranium workers (Beaverlodge andPort Radium miners, Port Hope uranium refinery/processing).
Radon decay product exposure, mortality (1950–1999) andcancer incidence (1969–1999)

Significant associations of radon decay productexposure and lung cancer mortality (n=618) in eachsubcohort. No associations for any other cancers. No
estimates of LLR risk

Analysis for radon decay products only. Not possible
to derive uranium-specific risks, because uranium
doses not used

McGeoghegan
and Binks[M27]

Mortality and cancer incidence at UK Springfields uranium
production plant, 1946–1995. Main activities, uranium fuelfabrication and UF6 production. 13,960 radiation workers

No measurements of internal exposure; all analyses forexternal radiation exposure

No analyses of internal exposure. Not possible toderive uranium-specific risks, because uranium doses
not used

McGeoghegan
and Binks[M26]

Mortality and cancer incidence studied at UK Capenhurst
plant, 1946–1995. Main activities, uranium enrichment for
military or power plant purposes. 12,540 employees

No measurements of internal exposure; all analyses forexternal radiation exposure

No analyses of internal exposure. Not possible toderive uranium-specific risks, because uranium doses
not used

McGeoghegan
et al. [M30]407 workers involved in 1957 Windscale uranium pile fire.
Mortality and cancer incidence 1957–2007. Estimatedplutonium, but not uranium, doses

No measurements of internal exposure; only externalradiation analyses

No analyses of internal exposure. Not possible toderive uranium-specific risks, because uranium doses
not used

Pinkerton et al.
[P21]

Mortality of 1,484 men employed in 7 uranium mills inColorado Plateau, USA (1940–1998). Mortality (SMRs)
examined by duration of employment and time since first
employment

No individual estimates of radiation exposure weremadeNo data available on internal exposure. Not possible toderive uranium-specific risks, because uranium doses
not usedRooney et al.
[R22]

Nested case-control study in five UKAEA facilities of incident
and fatal prostate cancer and exposure to radionuclides.
136 prostate cancer cases diagnosed 1946–1986 and 404
matched controls. 28 (21%) prostate cancer cases and 46(11%) controls had potential exposure to 3H, 51Cr, 59Fe, 60Co
and/or 65Zn

Prostate cancer: risk increased with duration andconcentrations of exposure to the targetedradionuclides (possible exposure RR: 2.36, 95% CI:
1.26, 4.43; documented exposure RR: 5.32, 95% CI:
1.87, 17.2)

Negative association with uranium exposure (yes/no).
Not possible to derive uranium-specific risks,because uranium doses not used

Indicated prostate cancer not associated with uranium
exposure; but uranium exposure was rare. Found0 cases and 6 controls with documented uranium
exposure (RR=0, 95% CI: 0, 2.55)

Vacquier et al.
[V1]

French CEA-COGEMA uranium miners (1946–1990) followed
up through 1999, mean 30.1 years. Radon exposure
estimates from ambient monitoring and worker job,
location and year. LLR exposure not estimatedElevated SMRs for total cancer (SMR=1.19), lung cancer(SMR=1.43) and kidney cancer (SMR=2.00), but notleukaemia. Significant radon dose response for lungcancer, but not kidney cancer or leukaemia. No
analyses with regard to LLR exposure

Not possible to derive uranium-specific risks, because
uranium doses not used

Walsh et al. [W6]

Radiation dose and prostate cancer mortality examined in
the 1970–1990 subset (55,435 miners) of German Wismut
uranium mining cohort. Follow-up, 1970–2003, n=263
prostate cancer deaths. Only gamma dose analysed

External gamma dose response: ERR per Gy=−1.18(95% CI: –2.4, 0.02)

Study analysed external dose and prostate cancer
mortality. Not possible to derive uranium-specificrisks, because uranium doses not used

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 449

Study references

Studies of workers monitored for potential exposure to uranium, with occupational dose records, but uranium-specific doses not explicitly analysed
Summary of study

Summary of findings relating to uranium

Relevance for this report

Boice et al. andRitz et al.[B38,R16]46,970 employees, Rocketdyne, USA (1948–1999); 5 801
had radiation exposure, 2,232 monitored for internal
radionuclides. Mortality follow-up 1948–2008 (mean33.9 years). Activities: operating research nuclear
reactors, fabricating nuclear fuel, disassembling anddecontaminating reactor facilities, decladding spent
nuclear fuel and storing nuclear material. Intakes of 14
radionuclides calculated for 16 organs using ICRP
biokinetic models; >30,000 urinalyses. Most significant
internal exposue was from enriched uranium, especiallyfor lung and kidney. A few workers received high lungdoses (~0.3 Sv) but 87% of workers had committed
equivalent dose to all tissues well below 10 mSv

For those monitored for internal exposure, no SMR excesses
seen for any cause: all cancers (except leukaemia);
leukaemia; cancers of lung, kidney, stomach, liver,prostate, brain; heart disease and cerebrovascular disease

Sufficient data to estimate internal radionuclide
exposure on basis of urinalyses. Uranium was largest
contributor to internal dose but other radionuclides
also present. Study limitations: relatively low career
doses, incomplete information on smoking. Study does
not suggest any strong uranium risk but has
uncertainties regarding uranium doses and small
numbers of deaths

Dufey et al.
[D30]

Leukaemia nested case-control study in cohort of 58,987
German Wismut male uranium miners, 1946–2003; 128
leukaemia cases (40 CLL and 88 non-CLL) andunspecified number of controls. Mining performed
1946–1989. Cohort mean dose 48.8 mGy, of which
external gamma contributed 40.9 mGy

Leukaemia: analyses for total dose to red bone marrow.
For a 2-years lag, linear ERR per Gy was 1.39 (90% CI:−0.77, 3.56) for all leukaemia and 2.08 (−0.84, 4.99) for
non-CLL. Suggestion of increased non-CLL only for thehighest dose group: for 0.4, 5.0, 25.6 and ≥103.7 mGy
RRs were 0.53, 0.89, 0.67 and 1.25 (90% CI: 0.69, 2.20),
respectively

Strengths included large cohort, long and high quality
follow-up. However, no assessment of LLR uranium
exposure, so study uninformative regarding uranium
effects

Dufey et al.
[D31]

Liver cancer mortality (n=159) in cohort of 58,987 male
German Wismut uranium miners, 1946–2003. Mining
1946–1989. Average liver dose, 47.9 mGy low-LET and
2.4 mGy high-LET irradiation; mean high-LET liver dose:
2.1 mGy from radon/progeny and 0.8 from LLR. Arsenic
measurements available

Liver cancer: the analysis by high-LET dose categories
did not reveal any statistically significant elevations inrisk, and dose-response analysis, adjusting for low-LET
dose, age and calendar years, yielded ERR per Gy=48.3(95% CI: –32, 129). Examined confounding factors
including arsenic exposure and alcoholism

Analysis adjusted for low-LET radiation exposure, but did
not account for radon decay product exposure.
Contribution of LLR to total absorbed liver dose was<2%, so study provides little information regarding
uranium risk

Dupree et al.
[D34]

995 white male employees (1943–1949) of Linde, USAuranium processing company followed up 1943–1979.
Doses reconstructed from ambient monitoring data,
surface contamination, urinalysis and film badges.
Exposure mainly to uranium with low solubility. Job
exposure was categorized as <10, 10–100 and>100 mSv/y of internal exposure (which was greaterthan external exposure levels)

Elevated SMRs found for laryngeal cancer (SMR=4.47,
95% CI: 1.4, 10.4, n=5), arteriosclerotic heart disease(SMR=1.19, CI: 1.01, 1.39, n=159) and non-malignant
respiratory diseases (SMR=1.52, CI: 1.04, 2.14, n=32). No
excess risk seen for lung, colorectal or lymphohaematopoietic malignancies. No analyses by uranium
exposure levels

No quantitative analyses by uranium exposure levels, so
uninformative regarding uranium risk. Given the
exposure information developed, cohort has somepotential to contribute to future uranium risk
assessment

Uranium doses, RR at 100 mSv:
All cancer except leukaemia: 0.98 (95% CI: 0.82, 1.17, n=266)
Lung cancer: 1.01 (CI: 0.89, 1.16, n=94)
Non-CLL leukaemia: 1.06 (CI: 0.50, 2.23, n=10)
Other trends for internal (mainly uranium) exposure: nonsignificant increasing trends, cancers of stomach, kidney,
brain/CNS, lymphomas; Non-significant decreasing trends,
cancers of colorectum, pancreas, prostate, bladder, andnon-malignant respiratory disease

UNSCEAR 2016 REPORT

Study references

450

Table A7.

Study references

Summary of study

Summary of findings relating to uranium

Guseva Canu etal. [G31]

Cancer mortality (1968–2005) examined in 2,709 maleworkers at French AREVA NC Pierrelatte uranium
enrichment and conversion plant. 15 former uranium
miners excluded. Uranium assessments: individual
dosimetry badges; faecal/urine bioassays and in vivo
measurements performed but not available

Kreuzer et al.
[K15]

Stomach cancer mortality (1946–2003; n=592) in 58,677
male German Wismut uranium miners and exposure toexternal radiation, alpha radiation, fine dust with silica
and arsenic. Mean estimated LLR exposure,
4.1 kBqh/m3 (<0.05 mGy), was substantially correlatedwith arsenic exposure

Stomach cancer: for alpha irradiation, ERR per Gy=22.5(95% CI: −27, 72) with statistical adjustment for otherexposure variables. RR in highest alpha dose category(10–26 mGy) not significant: 1.59 (CI: 0.69, 2.49)

Not analysed for LLR exposure. Only <1% of the alphadose to the stomach due to LLR, so uninformative
regarding a uranium risk

Mohner et al.
[M52]

Nested case-control study of leukaemia mortality (1953–
1998) among ~360,000 male German Wismut uranium
miners. 377 leukaemia deaths and 980 controls
matched on age. Cumulative red bone marrow (RBM)
doses from external radiation, radon decay products,
LLR and occupational medical diagnostic radiation(including 17,578 X-ray examinations), using a detailed
job exposure matrix. Mean cumulative LLR RBM dose
estimated to be <0.05 mGy

Added analysis of medical X-ray exposure tooccupational radiation sources. Analyses wereconducted only for total occupational RBM radiation
exposure, both internal and external and for medical
X-rays. Case-control analyses for LLR already reported
in [M50]

Report is uninformative for assessing uranium risk since
analyses were of all radiation exposure combined

Polednak andFrome [P24]18,869 white males worked at the Oak Ridge TEC
uranium conversion and enrichment plant (operated
1943–1947) but not at the Y-12 plant which succeeded
it. Workers in some departments (e.g. chemical dept.)
exposed to high ambient uranium dust. In 1945
average levels of uranium in air in various departments
ranged from 25 to 300 μg/m3. Among 226 men withurine samples, 72% had >0.01 μg/ml and 33% >0.05.
Mortality 1943–1977

Elevation in lung cancer (SMR=1.22, 95% CI: 1.10, 1.36)
but not higher among those working in areas withmore uranium dust or those with longer employment.
Mortality not elevated for stomach cancer (SMR=0.73),
kidney cancer (SMR=0.75), bone cancer (SMR=0.90) or
leukaemia (SMR=0.92)

Individual measurements of uranium exposure levels
available for only an undefined subsample of workers,
so dose-response analyses not conducted. Smoking
information not available

All cancer: the SMR=0.70 (95% CI: 0.60, 0.81, n=193)
Lymphoma and rectal cancer: non-significant increases
in rectal cancer (SMR=1.48, n=10) and non-Hodgkin’s
lymphoma (SMR=1.32, n=8). Trend analyses by timefirst employment and length of employment notsignificant for all cancer, lung cancer, upper aerodigestive tract cancer, and lympho-haematopoitic
malignancy

Relevance for this reportMortality in relation to internal radiation dose notreported, so study uninformative regarding uranium
risk

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 451

Summary of findings relating to uranium

Relevance for this report

Rage et al. [R4]Lung cancer mortality (1956–1999) studied among 3,377
French uranium miners hired ≥1955 when LLR andgamma ray measurements became available. Among
2,745 with exposure to uranium, mean was
1.63 kBqh/m3 (maximum 10.36). LLR contributed only
1.3% of total alpha-particle lung dose. Annual lungdose due to LLR significantly correlated with doses
from low-LET radiation (r=0.49), radon gas (r=0.53), andradon decay products (r=0.50)

Lung cancer: significant risk lung cancer mortality
associated with total absorbed lung dose (ERR per
Gy=2.94, 95% CI: 0.80, 7.53, n=66) and the alphaparticle absorbed dose (ERR per Gy=4.48, CI: 1.27, 10.9).
Assuming RBE=20 alpha-particles, ERR per Gy for total
weighted lung dose was 0.22 (CI: 0.06, 0.53). LLR ERR
was 5.0 (CI: 1.2, 12.3) per 10 mGy

Statistical analysis of LLR provided only weak
information; since LLR were correlated with and a small
percentage of total exposure the LLR risk estimate may
be inaccurate. No information on smoking habits.
Therefore study provides little information regarding
uranium risk

Zablotska et al.
[Z1]

Mortality (1950–1999) and cancer incidence (1969–1999)
of Port Hope, Canada radium and uranium process
workers. 2,472 (87% males) worked only with uranium.
Gamma was predominant radiation exposure, so
analyses were of gamma and radon decay products(RDP), not of LLR. Urinalysis for uranium begun in early
1960s; alpha counting of urine samples for workers
exposed to enriched uranium conducted on a limited
basis, so not used in dose assessment

No significant elevations in various cancer SMRs. No
excess cancer incidence seen for a number of cancer
types or all cancer. Dose-response analyses reported
for RDP and external gamma. In uranium workers, lungcancer RDP risk estimate non-significantly elevated.
Other malignancies and circulatory diseases: no
significant dose related elevations in risk for either RDP
or external exposure

Study of uranium workers was negative, other than aweak association of RDP exposure and lung cancer
incidence (but not mortality). Had no LLR exposure
estimates, so analyses of uranium effects could not bepresented. Substantial uncertainties: limited or no
exposure information for early workers, lack ofsmoking information. Study uninformative regarding
uranium exposure risk

UNSCEAR 2016 REPORT

Summary of study

452

Study references

Table A8. Studies of workers monitored for potential exposure to uranium for whom uranium-specific doses have been used in analyses so that uranium risks can be
explicitly examined
Abbreviations: LLR, long lived radionuclides; n.e., not estimable

Summary of study

Summary of findings relating to uranium

Relevance for this report

Carpenter et al.
[C7]

Case-control study of brain/central nervous system (CNS)
cancer deaths in workers (1943–1977) at 2 nuclear
facilities at Oak Ridge (ORNL and TEC/Y-12):
enrichment of 235U and conversion to UF4 (TEC, 1943–
1947); fabrication and testing of components fornuclear weapons (Y-12); nuclear energy technology
R&D (ORNL). 72 male and 17 female brain/CNS cancer
deaths (1943–1979). 4 matched controls per case. Work
locations/years rated by industrial hygienist for levelsof 26 agents, including uranium compounds

CNS cancer: 63% of brain/CNS tumours were malignant
glial tumours. Ever exposed to uranium, odds ratio(OR)=1.06 (95% CI: 0.5, 2.3, p=0.88) with no exposurelag, or 0.94 with a 10 years lag. Lagged levels of gradeduranium exposure (grades 1, 2 and 3, with 0=no
exposure as referent) had non-significant ORs of 0.88,
1.01 and 0.70, respectively. Analysis by duration ofheavier uranium exposure (grades 2–3, 10 years lag)showed: OR=0.86 for 1–3 years; 0.79, 3–10 years; 0.99,
10–20 years; 1.63, >20 years (n=3), not significant

Only semi-quantitative imputation of amount of
exposure. No elevated risk was apparent, but thereliability of dose categories unclear. Limited
information regarding uranium risk for brain/CNS
tumours

Chan et al. (andsupplement)[C17]

Mortality among 6,759 workers at Paducah Gaseous
Diffusion Plant, USA studied for 1952–2003. Workers
had potential exposure to external and internalradiation,uranium,
severalothermetals,trichloroethylene and other chemicals. Urinalyses ofuranium used to characterize the cumulative dose ofinternally deposited radionuclides as μg.years

RRs for different uranium exposure quartiles compared
to exposure quartile (<21 μg.years) are provided intable 15 (lung cancer); table 17 (leukaemia and nonHodgkin’s lymphoma); table 18 (pancreatic cancer);
table 20 (brain/CNS cancer)

Provides grouped quantitative information about thehealth effects of uranium exposure, though LLR linear
dose-response risk coefficients were not given.
Possible confounding by chemicals and smoking.
Provides semi-quantitative information regarding
uranium risk for several cancer sites

Checkoway et al.
[C20]

Mortality (1947–1979) investigated among 6,781 whitemale workers at Y-12 uranium fabrication plant (OakRidge, USA). 3,490 monitored for internal exposure.
Internal dosimetry: urine analyses begun in 1950, fully
implemented by 1953, and in vivo measurements
added in 1961. Internal lung doses calculated using
metabolic models. For monitored workers, mean lungdose 82.1 mSv. Mean external dose, 9.6 mSv. Other
exposure: beryllium, solvents, machine oils, mercury,lead. 45 lung cancer deaths in those monitored for
uranium exposure

Lung cancer: (n=45) Analysis with 10-years dose lag foralpha irradiation, compared to 0–<10 mSv group,

Provides some information on lung cancer among thosewith measured uranium exposure; showed little
apparent risk. Small numbers of lung and other cancers
limit the quantitative estimates. Smoking information
not available. Provides semi-quantitative information
regarding uranium risk for lung and brain/CNS cancers

0–49 mSv, RR=0.93 (95% CI: 0.41, 2.12)
50–99, RR=0.66 (CI: 0.23, 1.90)≥100, RR=1.12 (CI: 0.47, 2.65, n=11)
Brain/CNS cancer: (n=14) with no lag,
10–49 mSv, RR=1.10 (0.19, 6.5)
≥50 mSv, RR=0.45 (0.06, 3.2)
Other cancers: no bone cancers observed. Trend
analyses for kidney cancer (n=6) or other a priori
cancers not reported for uranium monitored cohort

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 453

Study referencesSummary of findings relating to uranium

Relevance for this report

A case-control study of circulatory system disease (CSD,
n=442) mortality, particularly ischaemic heart disease(IHD, n=167) and cerebrovascular disease (CeVD,
n=105), and 237 matched controls, nested among
5,086 French CEA-COGEMA uranium miners who werefirst employed after 1955, followed up through 2007,
mean 35.4 years. Individual exposure estimated fromambient
monitoring,
1959–1982
anddosereconstruction for 1956–1958. Since 1983, individual
LLR exposure estimated with film dosimeters. Mean
cumulative LLR was 1.2 kBqh/m3 (max=7.6)

Circulatory disease: statistically significant association ofradon exposure with both CSD (hazard ratio(HR)=1.11/100 WLM) and CeVD (HR=1.25/100 WLM) risk.
Records contained information on a number of medical
CSD risk factors for a subset of cases and controls. After
adjusting for radon and external gamma exposure andfor empirically the main medical risk factors, found LLR
HRs per kBqh/m3 of:

Had a substantial set of individual measurements of LLR
exposure. After adjusting for radon and externalradiation, detected no significant LLR risk for CSD, IHD
or CeVD. Confounding by medical CSD risk factors
proved to be small. Uncertainties: exposure prior to1983 estimated from ambient measurements or withno measurements may have had appreciablemeasurement error. Relevant for uranium risk
assessment of CSD

CSD: 1.13 (95% CI: 0.97, 1.31, n=76)
IHD: 0.94 (CI: 0.73, 1.20, n=26)
CeVD: 1.17 (CI: 0.90, 1.53, n=16)
LLR risks in entire nested case-control sample, but without
being able to adjust for medical risk factors, nearly identical

Drubay et al.
[D23]

Kidney cancer mortality in 3,377 French uranium miners
1956–2007, and 58,986 German uranium miners 1946–
2007. For respective French and German cohorts,
median durations of follow-up were 30.0 and 34.8
years; respective median kidney doses were 26.7 mSv(range 0, 498) and 34.4 mSv (range 0, 2 905)

Kidney cancer: SMRs were 1.49 (95% CI: 0.73, 2.67, n=11)
for French cohort and 0.91 (CI: 0.77, 1.06, n=174) for
German cohort. A 10-years lagged LLR dose-response
analysis showed hazard ratio (HR) per kBqh/m3 with
10-years lagged cumulative dose of 0.89 (CI: 0.55, 1.42)
for the French and 1.009 (CI: 0.991, 1.027) for the GermancohortLack of association with estimated LLR exposuresuggests kidney cancer effect is likely small.
Uncertainties: no information about smoking; Frenchcohort was small; workers also had radon and external
gamma exposure

Dupree et al.
[D35]

Nested case-control study of lung cancer mortality (787cases with 787 matched controls) in cohorts at 4 USA
uranium processing facilites: TEC (operated 1943–1947
only) and Y-12 (1947–1982), Mallinckrodt (MCW, 1942–1966) and Fernald (FMPC, 1947–1982). Maximum
follow-up period, 1943–1983. Primary radiation hazard
was from airborne dust of mainly insoluble naturaluranium compounds. For FMPC, MCW and TEC,
ambient uranium monitoring to estimate internalradiation doses. Y-12 also had whole body counting
and urinalysis. Conversion to doses assumed Type S
uranium exposure. Smoking status available for 48% ofcases and 39% of controls

Lung cancer: analysis of internal radiation (primarily fromambient uranium), using 10-years lagged dose with
<0.5 mGy as baseline, showed no increased risk forworkers exposed below 250 mGy. Lung cancer odds
ratios (ORs) for 0.5, 2.5, 5, 25, 50, >250 mGy were 1.03,
0.57, 0.85, 0.82, 0.64 and 2.05 (95% CI: 0.20, 21),respectively for LLR. However, for the 166 case-control
pairs with smoking data, no elevation in risk (odds ratioof 0.36 in the highest internal dose group)

Analysis had large number of lung cancer deaths, afraction had urinalyses and/or whole body counting in
addition to ambient monitoring data. Suggests that
effect of uranium exposure must be small. Estimate of
ERR per Gy not reported, but the grouped dose data do
not suggest a statistically significant elevated lungcancer risk. Limitations: uncertain dose estimates forearly workers, concomitant exposure to radon andexternal radiation, and limited smoking information.
Provides semi-quantitative information regarding
uranium risk for lung cancer

UNSCEAR 2016 REPORT

Drubay et al.
[D24]

Summary of study

454

Study referencesSummary of study

Summary of findings relating to uranium

Relevance for this report

Guseva Canu etal. [G31]

Lung cancer mortality in 2,709 workers at the French
AREVA NC Pierrelatte uranium reprocessing plant
during 1960–2005. The plant enriched uranium via
gaseous diffusion caused uranium chemical
conversion. Uranium was only radioactive material
used at the plant. Semi-quantitative JEM tocharacterize uranium exposure (duration and intensity
specific for each job and calendar year, on a 4 point
ordinal scale). Smoking data on 6% of cohort

Lung cancer: (n=48) for durations of exposure >1 year to
Type F, M and S uranium compounds found hazard
ratios (HRs) of 1.05 (95% CI: 0.43, 2.52), 2.61 (0.87, 7.8),
2.58 (0.76, 8.8), respectively. For duration of exposure
as a continuous variable, HRs of 1.01 (0.96, 1.01), 1.07
(1.01, 1.13) and 1.07 (1.01, 1.14) per year exposure,respectively

Suggestions of elevated lung cancer risk after exposure
to slowly soluble (Types M and S) uranium compounds.
However, substantial uncertainties: small number oflung cancer cases, only an ordinal scale of exposure
intensity,
limitedsmoking
information,noindividualized information on chemical exposure.
Provides semi-quantitative information on uranium risk
for lung cancer, with analyses by uranium solubility,
but uncertainty due to small numbers

Guseva Canu etal. [G32]

Quantified uranium exposure using a job exposure
matrix for 2,897 workers at French AREVA NC
Pierrelatte uranium reprocessing plant. Classified
exposure by natural vs. reprocessed uranium, and bysolubility Types F, M and S. To model cumulative
exposure for individuals, estimated duration andintensity of exposure to Types F, M and S

Lung cancer: (n=53) For natural uranium exposure, nosignificant associations for Type F, M or S. For
reprocessed uranium, significant hazard ratios (HRs) inthe highest cumulative exposure group of 4.35 (95% CI:
1.25, 15) for Type M and 10.5 (CI: 2.3, 48) for Type S.
Dose-response analysis for reprocessed uranium
exposure duration gave HRs of 1.13 (1.03, 1.25) and1.13 (1.01, 1.25) for Types M and S, respectively.
Analysis in subgroup of 345 workers with smoking
information suggested no confounding by smoking

This exploration of chemical types and radioactivity levelof various forms of uranium suggests that naturaluranium has little or no lung carcinogenic effect.
However, the less soluble forms of reprocesseduranium dust, with their greater radioactivity andrelatively long residence time, may induce lung cancer
Limitations: study size was small. Only limited smoking
data available. Results require confirmation in larger
independent study

Lympho-haematopoietic malignancies: (n=21) Found
association with insoluble reprocessed uranium, but on
the basis of only 3 exposed cases
Guseva Canu etal. [G33]

Mortality from ischaemic heart disease (IHD, n=48),
cerebrovascular disease (CeVD, n=31) and total
circulatory system diseases (CSD, n=111) after chronic
exposure to uranium among 2,897 workers at the
French AREVA NC Pierrelatte uranium processing plant(1960–2006). Cumulative exposure to various uranium
compounds was classified by isotopic composition andsolubility type and quantified for individual job
histories via a job-exposure matrix (natural vs.
reprocessed uranium (RPU), and absorption Types F, M
and S)

Circulatory disease: CSD mortality was increased among
workers exposed to Type S RPU (HR=2.13, 95% CI: 0.96,4.70) and Type S natural uranium (HR=1.73, CI: 1.11,
2.69). Additional information on risk by duration andintensity of exposure for CSD, IHD and CeVD is given in
the text of table 20
For the subset of workers with available smoking data
they found nominally higher CSD HRs for RPU Type M
and S exposure among smokers than non-smokers, but
numbers of cases were small

Job exposure matrix was carefully done, but provides only
an approximate quantitation of uranium exposure.
Concerns with study are: small study size, limited
smoking data. Possible confounding: heat andtrichloroethylene exposure were correlated withuranium exposure. The analyses of Types M and S RPU
exposure did not adjust for the common exposure to
Type F and natural uranium. Provides semi-quantitative
information regarding uranium risk for CSD, including
analyses by uranium solubility and isotopic composition,
but with uncertainty due to small numbers

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 455

Study references

Kreuzer et al.
[K17]

Kreuzer et al.
[K18]

Mohner et al.
[M50]

Summary of findings relating to uranium

Relevance for this report

Circulatory system disease (CSD) mortality (1946–1998),
German Wismut male uranium miner cohort in relation
to external radiation, radon and LLR exposure.
Exposure estimated via a job exposure matrix JEM foreach radiation type,1946–1989. 5,417 deaths from CSD(1946–1998), including 3,719 from heart disease and1,297 from cerebrovascular disease. Mean LLR
exposure, 3.5 kBqh/m3 (maximum 132)

Circulatory disease: analyses for cumulative LLR
exposure lagged by 5 years. The LLR risk estimates (ERR
per 100 kBqh/m3) were

The largest systematically defined cohort of uranium
workers available. Uncertainties: no confirmation of
JEM by urinalyses, correlation of LLR exposure withexternal radiation and radon exposure not accounted
for, lack of dosimetric models to estimate LLR exposureof the heart and major arteries

Mortality from cancer of the extra-thoracic airways
among 58,690 male German Wismut uranium miners(1946–2008) in relation to radon and cumulative LLR
exposure. LLR exposure estimates derived by a job
exposure matrix based on ambient measurements as
described in Kreuzer et al. [K14]
Mortality in 4,054 male German uranium millers (1946–
2008) who had never worked as uranium miners, so
radon exposure was low, mean 8 WLM. Estimatedexposure to radon, external gamma radiation, LLR andsilica. Exposure estimates derived via a job exposure
matrix of intensity (from ambient monitoring) by
location, job, calendar year. Mean LLR: 3.9 kBqh/m3.
Preliminary organ dose calculations for alpha-emitting
LLR averaged 3 mGy for lung, and 1 mGy for liver andred bone marrow
Nested case-control study of leukaemia mortality (1953–
1989) among ~360,000 German Wismut male uranium
miners: 377 leukaemia deaths and 980 controls
matched on age. Job exposure matrix JEM by location,
job and year used to estimate red bone marrow (RBM)
exposure. JEM for radon and decay products (RDP),
external radiation and LLR exposure estimates for >500
different workplaces, 750 job titles, 44 calendar years.
Mean cumulative RBM dose was 23.6 mGy; only 2%
from inhalation of LLR

CSD: −0.2 (95% CI: −0.5, 0.06)
Heart disease: −0.3 (CI: −0.6, 0.02)
Cerebrovascular disease: −0.05 (CI: −0.5, 0.6)
In no case did the highest dose category show significantly
elevated risk. Ischaemic heart disease (n=2,690) also did
not show a statistically significant elevation
Cancer of extra-thoracic airways (n=234): non-significant
increase with radon exposure: ERR/100WLM=0.036,
95% CI: −0.009, 0.08
No increase with LLR exposure: ERR per 100 kBqh/m3=−0.17, 95% CI: −2.50, 2.16 (adjusted for radon
exposure)
All cancer: LRR ERR per 100 kBqh/m3 = −0.43 (95% CI:−1.31, 0.44, n=457), adjusted for radon exposure
Lung cancer: LRR ERR=−0.61 (CI: −1.42, 1.9, n=159), (notadjusted for radon)

Estimate of risk for uranium exposure had potential
confounding by external radiation levels, arsenic andsilica dust exposure and smoking habit. Quantitativerisk estimate is relevant for uranium risk assessment:suggests little risk for extra-thoracic airways

Well conducted study, suggesting little/no association of
uranium exposure with various health outcomes. Study
size was small and no smoking information available.
Quantitative risk estimates are relevant for uranium risk
assessment

Additional LLR risk coefficients given in text of table 16(lympho-haematopoietic), table 17 (colon and rectal),
table 18 (kidney and prostate), table 20 (circulatory)

Leukaemia: non-chronic lymphocytic leukaemia (nonCLL) risk not associated with RDP, but showedsuggestive association with LLR exposure. ERR per
100 kBqh/m3 for LLR was 1.04 (90% CI: –0.64, 2.73,
n=377) for all leukaemia, 0.76 (CI: –1.26, 2.78, n=218)
for non-CLL and 1.35 (CI: –1.54, 4.24, n=159) for CLL.
Suggestion that the highest/longest LRR doses may
increase risk: for ≥20 kBqh/m3, (OR=1.26, 90% CI: 0.71,
2.22) for non-CLL. For acute myelogenous leukaemia,
the LRR ERR per 100 kBqh/m3 was 0.83 (CI: –1.9, 3.6)

Provides evidence that LLR exposure has little
association with leukaemia risk. Limitations: prior to1955 little data on exposure levels so dose
uncertainties.
Mortality
may
have
been
underascertained because inadequate identifiers inearly years limited mortality linkage. Underlying cohortand numbers in it rather loosely defined, though it is
the largest uranium worker cohort. Quantitative risk
estimates are relevant for uranium risk assessment

UNSCEAR 2016 REPORT

Kreuzer et al.
[K14]

Summary of study

456

Study references

Study references

Summary of study

Summary of findings relating to uranium

Relevance for this report

Mohner et al.
[M51]

Nested case-control study of laryngeal cancer among~360,000 German Wismut male uranium miners ever
employed, 1950–1989. Tumour registry follow-up,
1961–1989. Two matched controls per case. Crude
information on smoking habits available for many
workers and anecdotal information on alcohol
consumption from medical records. Included 554
laryngeal cancer cases and 929 controls

Laryngeal cancer: elevated risk in highest cumulative LLR
exposure category (≥10 kBqh/m3), OR=1.63 (95% CI:
1.03, 2.59, n=56), adjusted for smoking and alcohol
intake. For continuous LLR cumulative exposure:
ERR=0.098 (CI: –0.11, 0.31) per 10 kBqh/m3, unadjusted,
or ERR=0.156 (CI: –0.11, 0.41) adjusted for smoking andalcohol intake

The same limitations for this study as for Mohner et al.
[M50]. Follow-up successful for only 72.8% of potential
controls. Quantitative risk estimate for laryngeal cancer
is relevant for uranium risk assessment

Rage et al. [R5]5,086 uranium miners employed by CEA-COGEMA in
France; followed up 1946–2007 (mean 32.8 years).
Cohort included 3,377 miners first employed after
1955, for whom radon, LLR and external γ-ray exposure
was recorded. Assessment of LLR exposure based on
ambient measurements 1959–1982 and individual
measurements thereafter. Doses retrospectively
reconstructed for the period 1956–1958 [R4]. Post-1955
workers had mean of 1.64 kBqh/m3 of LLR (range 0.01–
10.4)

Internal LLR dose-response analyses were conducted,
doses lagged 5 years. LRR results expressed as
ERR/kBqh/m3:
All cancer, 0.022 (95% CI: −0.049, 0.12, n=315);

LLR exposure related to various mortality end points.
Limitations: LLR correlated with and a small percentage
of total radiation exposure; smoking information
unavailable; lung cancer analyses did not adjust forsilica exposure. Quantitative risk estimates are relevant
for uranium risk assessment

Nested case-control study of lung cancer among 3,864
Y-12 (Oak Ridge, USA) workers hired 1947–1974. Y-12
was a nuclear material fabrication plant. Internal
exposure primarily LLR from ambient uranium dust.
Individual monitoring for external radiation exposure
began in 1948 and became plant-wide in 1961.
Urinalysis monitoring increased in coverage throughthe 1950s and in vivo monitoring begun in 1961. Mean
external lung dose (10.1 mSv) was fourfold lower than
mean cumulative internal dose (44.7 mSv). Other
exposure: beryllium, solvents, machine oils, mercury,
lead

Nested case-control analyses were conducted, withmatched controls. Exposure lagged by 5 years

All cancers except lung, −0.065 (n.e. 0.019, n=221);
Other LLR risk coefficients given in text of table 15(respiratory), 19 (kidney), 20 (brain/CNS) and 21(circulatory). In summary, only lung cancer showed asignificant positive association with LLR exposure

Lung cancer: LLR dose-response negative (ERR per
100 mSv=–0.077, 90% CI: –0.23, 0.07)
Smoking-related diseases other than lung cancer: LRR(ERR per 100 mSv=(–0.089) was negative, as was nonmalignant respiratory disease (–0.085), but with wide
confidence intervals

Provides evidence that uranium risk for lung cancer islikely small. Strength: measured exposure (both urine
assays and in vivo monitoring) and the fact that radon
exposure did not overshadow the LLR exposure of thelung. LLR exposure had substantial uncertainties, inpart because 58% of exposure person-years had
imputed rather than measured doses. Limitations: dose
uncertainties, inadequate information on smoking andworkplace chemicals, statistical power limited.
Quantitative risk estimate for lung cancer is relevant for
uranium risk assessment

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 457

Richardson andWing [R12]Lung, 0.32 (0.09, 0.73, n=94);

Summary of findings relating to uranium

Relevance for this report

Silver et al. [S19]

Cohort of 6,409 uranium workers at Fernald (USA)employed (1951–1985), and followed through 2004
(mean follow-up 37 years). Used urine uranium
concentration data (>250,000 urine samples) from1952 forward to estimate exposure to internallydeposited uranium compounds. Mean cumulative
doses to the lung for hourly and salaried workers were1,552 μGy and 388 μGy for LLR, respectively. Mean LLR
cumulative organ dose ranged from 1.1 mGy (lung) to
6.7 mGy (pancreas)

Analyses took into account pay code, birth year,
trichloroethylene exposure, radon and externalradiation. Overall: hourly males showed excess lungcancer (SMR=1.25, 95% CI: 1.09, 1.42, n=297). LRR ERRs
calculated for Caucasian males per organ-specific
100 μGy:
Intestinal cancer (small intestine and colon, notrectum): had a significant elevation in the highest dosegroup (>36 μGy, ERR=1.7, CI: 0.17, 5.7) and a significant
dose response (ERR 100 per μGy=1.5, CI: 0.12, 4.1,n=48). Other dose-response estimates at 100 μGy forinternal doses were null. Additional LLR risk coefficients
given in text of table 15 (lung, respiratory), table 16(leukaemia, lymphoma), table 17 (stomach, pancreas),table 18 (kidney)

This study has longer follow-up and better exposure
assessment than previous ones. Uranium internal
doses estimated for several different organs and linear
ERR estimates adjusted for other radiation exposure.
Sole positive finding related to intestinal cancer, which
is not a very high a priori suspect, so requires
confirmation. Limitations: no smoking data, limited
data on exposure to chemicals and other hazardous
substances, limited statistical power. Quantitative risk
estimates are relevant for uranium risk assessment

9,973 Czech uranium miners studied for leukaemia andnon-Hodgkin’s lymphoma risk. Two cohorts: S, 4,348
exposed 1948–1963; N, 5,625 exposed 1968–1986.
Though had limited exposure measurements, derived
location-job-year estimates of dose rates for hewers
and then proportionately scaled for other jobs todevelop a job exposure matrix JEM. Estimated 52–64%
of the red bone marrow (RBM) dose was from LLR.
Mean LLR RBM dose of 160 mSv for cohort S and
37 mSv for cohort N

Leukaemia: using 2 years dose lag, for 1−19 years since
1st exposure leukaemia SMR=1.0 (95% CI: 0.4, 2.1, n=7).
For >19 years since 1st exposure SMR=1.8 (CI: 1.2, 2.7,n=23). For total follow-up period, SMR=1.5 (CI: 1.1, 2.2,
n=30, mean RBM LLR dose 145 mSv). Due to small
numbers, did not separate out non-CLL leukaemias.
Leukaemia risk slope for total RBM dose (external,radon progeny and LLR), ERR per Sv=2.5 (90% CI: 0.3,
9.3)

Dose uncertainties were probably large, especially forearlier years (that contributed the highest exposure)and the number of malignancies was small. Analyses of
LLR exposure were by average SMRs and not dose
responses. Study suggestive of leukaemia risk fromuranium exposure, but study limitations weaken the
conclusions(Prior reports:[D33, R14,
R15])

Tomasek andMalatova [T13]

Vacquier et al.
[V2]French cohort of 3,377 uranium miners first employed
1955–1990 when exposure to external radiation, radon
and uranium dust (LLR) could be estimated. 3,240 had
internal exposure. Follow-up through 1999, a mean of
26.5 years; mean LLR exposure, 1,632 kBqh/m3. LLR
exposure estimates: reconstructed before 1959;
ambient measurements 1959–1982; since 1983
personal film dosimeters. Since external radiation, LLR
and radon exposure instances were correlated,
determined which were associated with cancer risks

Non-Hodgkin’s lymphoma (NHL): dose response notsignificant (p=0.16), though a nominal overall excess(SMR=1.5, CI: 0.9, 2.2)
LLR exposure was correlated r=0.52 with cumulative
radon exposure, 0.47 with external exposure.
Combined cumulative radiation exposure showedsignificant dose response only for lung cancer. Linear
ERR risk coefficients per (kBqh.m-3) for LLR exposure,lagged 5 years:
All cancer: 0.001 (95% CI: –0.08, 0.11);
Lung cancer: 0.25 (CI: 0.02, 0.70);
Brain/CNS cancer: 0.17 (CI: n.e., 2.0)

LLR dose-responses could not be calculated withadjustment for radon or external exposure. Substantial
correlation of LLR with radon and gamma exposure, so
LLR risk estimates not well defined but suggestive
small association with lung cancer. Limitations:
uncertainties in exposure assessments, correlated
exposure, lack of smoking data. Quantitative risk
estimates are relevant for uranium risk assessment

UNSCEAR 2016 REPORT

Summary of study

458

Study references

Study references

Summary of findings relating to uranium

Relevance for this report

Yiin et al. [Y3]

Nested case-control study of multiple myeloma among
47,941 workers at the K-25 Oak Ridge, USA gaseous
diffusion uranium enrichment plant (operated 1945–1985). Five matched controls per case. Exposure tosoluble and insoluble uranium compounds. Individual
uranium dose estimates based on ambient
measurements and urinalysis. Formed groups
according to the strength of the dosimetry: grouping I,
multiple urinanalyses and extensive ambient
measurements; groupings II also included those withfewer measurements; grouping III, all cases andcontrols. External radiation doses and medical
radiation exposure also compiled

Multiple myeloma: 98 multiple myeloma death cases
and 490 controls. Analyses adjusted for birth cohort,
external and medical irradiation, mercury, nickel andtrichloroethylene exposure. For those with the best
estimated uranium doses to the bone marrow(Group I), the odds ratio (OR) at 10 μGy was 1.04 (95%
CI: 1.00, 1.09). For total case-control group (Group III),
OR was identical: 1.04 (CI: 1.00, 1.09). Indicated a weakassociation of bone marrow dose from uranium withmultiple myeloma risk

High quality study. Included detailed uranium exposure
assessment and analyses adjusted for external andmedical radiation doses and prevalent chemicals. Weakassociation found between uranium bone marrowdose and multiple myeloma risk; requires confirmation
by other studies. Limitations: less measurement data
were available for workers in the earlier days.
Quantitative multiple myeloma risk estimate is relevant
for uranium risk assessment

Zhivin et al.
[Z10]Studied 4,688 French gaseous diffusion uranium
enrichment workers (AREVA NC, CEA and Eurodif) with
exposure to mainly soluble uranium compounds(Type F). Used plant-specific job exposure matrices
JEMs to estimate cumulative exposure. The AREVA NC
job exposure matrices showed 64% sensitivity and 80%
specificity in validation against bioassay data. Medianfollow-up, 30.2 years, 1968–2008. Had estimates ofpotential
confounding
exposuresituations:trichloroethylene, heat, noise

From job exposure matrices, grouped workers into no,low, medium, high exposure for analysis. Analysed
external radiation, and natural, enriched and DU.
Analysed all cancer, lung cancer, lymphohaematopoietic malignancies, and circulatory diseases.
Results by exposure group presented in text of table 15(lung), table 16 (lympho-haematopoietic), table 20(circulatory)

Study developed quantitative estimates of uranium
exposure, but analysed only low, medium and highgrouped uranium exposure. Valuable because it
considered mainly highly soluble uranium andcompared natural, enriched and depleted isotopic
forms.
Providessemi-quantitative
information
regarding uranium risk for CSD, including analyses ofisotopic composition

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 459

Summary of study

460

Table A9.

Studies of groups with potential military uranium exposure

Bogers et al.
[B28]

Summary of study

Summary of findings related to uranium

Relevance for this report

Following lay-press reports of alleged excess leukaemia
among Dutch Balkan veterans, study examined cancer
incidence, comparing 18,175 Balkan-deployed military
male personnel with 135,355 non-Balkan deployed
military males and with general population rates.
Maximum follow-up was for nearly 15 years. Some
differences between the Balkan and non-Balkan
cohorts’ military status, e.g. conscripted soldiers (15%vs. 81%, respectively)

All cancer: total cancer incidence 17% lower among
Balkan than non-Balkan personnel (hazard ratio (HR)0.83, 95% CI: 0.69, 1.00)

No information about exposure to DU available, so study
non-informative about uranium effects

Miscellaneous cancers: rates of digestive, respiratory,urogenital and haematological cancers nonsignificantly lower in the Balkan vs. non-Balkan group
Leukaemia: HR could not be calculated for leukaemia
because of the small number of cases (Balkan n=5)

Hines et al. [H18]37 US Gulf War I veterans who had inhalation exposureto (and sometimes retained fragments of) DU fromfriendly fire incidents were examined. Compared thosewith high vs. low body burdens of DU, as measured byurine assay. DU remaining in the body is 40% lessradioactive but chemically similar to natural uranium.
Low and high exposure groups similar in age, race, BMI
and smoking

Non-malignant respiratory disease: no differences
between low and high exposure groups for any of a listof pulmonary symptoms or for history of steroid
prescriptions. No significant differences regarding DU
exposure levels in pulmonary function parameters or
chest CT findings

Study strengths: had clinical examinations, spirometrictesting, symptom reporting, and smoking information.
DU findings do not indicate any non-malignant
pulmonary effects. Provided limited information
regarding uranium effects due to small sample size andlikelihood that exposure levels were low

Labar et al. [L3]

Ecological study to examine childhood haematologicalcancers in Croatian counties with DU (10 counties),
chemical plant damage (2 counties) or "population
mixing" (4 counties). Compared disease rates forchildren ages 0–14 before (1986–1990), during (1991–
1995) or after (1996–1999) the Croatian war in thosecounties

Childhood haematological malignancies: in the 10
counties with DU exposure, no significant increases
during or after the War were found for lymphatic
leukaemia, myeloid leukaemia, Hodgkin’s lymphoma,
or non-Hodgkin’s lymphoma

No evidence for an effect of DU on childhood
haematological malignancies. Exposure very low, andecological data are susceptible to various unidentified
biases, so study provides no meaningful information
on uranium risk

Macfarlane et al.
[M2]

A 13 year follow-up was conducted of 51,753 UK
veterans deployed in the Gulf War and 50,808 otherveterans matched for age-group, sex, rank, service andlevel of fitness, who were not deployed to the Gulf.
57% responded to a questionnaire about deployment
experiences and morbidity

7% of those with questionnaires responded they hadreceived DU exposure, among whom there were 9
disease-related deaths. The DU exposed vs. unexposed
yielded a RR=1.00 (95% CI: 0.99, 4.04) after adjustment
for age, sex, smoking and alcohol intake

DU exposure was based on unverified self-reports andthe risk estimate was for overall disease deaths, not apriori causes. The small number of deaths (n=9) among
those reportedly exposed to DU is a very weak finding

McDiarmid et al.
[M22]

US Gulf War veterans with DU exposure were followedup (1991–2005) for clinical and laboratory end points.
On basis of repeated urine uranium measurements,
10 were designated as high DU exposure and 24 as low
exposure. Exposure resulted from inhalation, woundcontamination and/or embedded fragments (for~30%)

The extensive clinical examination did not show any
differences between the high and low exposed group.
Other high/low exposure comparisons: no difference inurine retinol binding protein, a biomarker of renal
proximal tubule function. No differences found on:other renal measures, a neurocognitive test battery,neuroendocrine parameters, semen parameters, or
HPRT mutations. A borderline increase in chromosome
aberrations in the high exposure group

Among the several dozen parameters measured, only
chromosome aberrations showed a (suggestive)
difference. The multiple comparisons and small sample
size limited the statistical power and meaningfulnessof the comparisons

UNSCEAR 2016 REPORT

Study referencesSummary of study

Summary of findings related to uranium

Relevance for this report

McDiarmid et al.
[M23]

35 US Gulf War veterans with DU exposure in 1991 wereagain evaluated in 2007 with numerous clinical andlaboratory measures

Only two parameters showed marginal differences
between the high- and low exposure groups: β2
microglobulin (81.7 vs. 69 μg/g creatinine, respectively;
p=0.11) and retinol binding protein (48.1 vs. 31 μg/g
creatinine; p=0.07). No differences were seen in rates ofchromosome aberrations or HPRT mutations

Among the several dozen parameters measured, only
two showed a suggestive difference. The multiple
comparisons and small sample size limited thestatistical power and meaningfulness of thecomparisons

McDiarmid et al.
[M25]

37 US Gulf War veterans with DU exposure in 1991 wereagain evaluated 20 years after exposure (2011) for
numerous clinical and laboratory end points. Reportfocused on acute renal toxicity and included three newsensitive markers of kidney tubular injury

No differences between high- and low-exposure groups
for: haematology, clinical chemistries, neuroendocrine
parameters, bone metabolism, neurocognitive
function, immune function, pulmonary function or
nodules. Regarding renal function and injury, no highvs. low exposure differences were found for 16 clinicalindicators of renal function, 6 urine markers for kidney
injury, or 4 urine measures of low molecular weightproteins, although a re-analysis using a different
definition of high exposure showed elevations in two
kidney injury markers

Two sensitive markers of kidney tubular injury suggested
subtle renal injury, but this found only after the maincategories of high vs. low exposure showed nodifferences. Multiple comparisons and small sample
size limit the inferences that can be drawn from thestudy

Strand et al.
[S42]Cancer risk and all cause mortality studied among 6,076
Norwegian military UN peacekeepers in Kosovo, 1999–
2011. No information available on DU exposure. Mean
follow-up, 10.2 years; 4.4% women

69 cancer cases observed (SIR=1.04, 95% CI: 0.81, 1.33).
Suggestion of elevation in melanoma (SIR=1.90, CI:
0.95, 3.4, n=11). No elevation in stomach, liver, lung,
prostate, kidney, bladder, brain cancers or lymphohaematopoietic malignancies

With no data on DU exposure, study is uninformative
regarding uranium risk

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 461

Study references

462

Study references

Summary of study

Summary of findings related to uranium

Relevance for this report

Boice et al. [B31]

Study evaluated cancer incidence (1993–1997) amongresidents near Apollo (began operations 1957) andParks uranium-plutonium processing plants in
Pennsylvania, USA. Study population included about
17,000 individuals in 8 nearby municipalities

All cancer: found 581 incident cancers when 574 were
expected (SIR=1.01, 95% CI: 0.93, 1.10)

Earlier investigation had obtained soil measurements ofuranium, plutonium and other isotopes and air
measurements of gamma radiation; all levels well
below Nuclear Regulatory Commission releaseguidances. Those measurements too sparse to be used
to directly assess uranium effects associated with
exposure of individuals. Limited environmental
measurements and negligibly low exposure levels
mean study is not informative as to uranium effects

Boice et al. [B30]

Boice et al. [B32]

People in two counties proximal to the Apollo and Parks,
Pennsylvania, USA former uranium/plutonium material
processing plants were concerned regarding possible
elevated rates of cancer, especially childhood
leukaemia. Study compared cancer mortality rates inthose two counties (population ~443,000) with six
other counties (population ~864,000) matched on age,
race, urbanization and socioeconomic factors.
Comparisons made before, during and after operationsof the uranium-plutonium plants

Cancer mortality rates investigated in Karnes County,
Texas, USA, a county with uranium mining and milling
activities from 1959 to early 1990s, with 3 mills and >40
mines. No uranium enrichment activities. Karnes cancer
mortality rates before, during, after that period (1950–
2001) compared with four match control counties.
1,223 cancer deaths observed in Karnes County (1,392
expected) and 3,857 in control counties. Texas
Department of Health monitored Karnes radiationlevels, found no elevations in radioactive material
in/near homes

A priori tumour sites: for tumour sites with potentiallygreater exposure they found: lung (SIR=0.88), kidney(SIR=1.05), non-Hodgkin’s lymphoma (SIR=1.10), liver(SIR=0.61) and bone (2 observed, 1.19 expected), noneof which were statistically significant elevations. Also
thyroid and female breast cancer rates were notelevated, nor was leukaemia
All cancer: during 1950–1995, 39,287 cancer deaths
occurred in the proximal counties and 77,382 in thecontrol counties. Compared to control counties, RRs in
proximal counties for all cancer deaths before (1950–
1965), during (1965–1980) and after (1980–1995) werevirtually identical—0.95, 0.95 and 0.98, respectively—
indicating no effect of potential uranium/plutonium
exposureStrengths: the mortality ascertainment was high, sample
size was large. Most likely uranium-related cancer types
examinedA priori tumour sites: for childhood leukaemia (total n=119
proximal county cases and n=272 control cases) before,
during, after RRs=1.02, 0.81, 0.57, respectively. Lungcancer (RR=0.85, 0.99, 0.95), bone (RR=0.96, 1.00, 1.01),
liver (RR=0.98, 1.07, 1.01) and kidney (RR=1.00, 1.08, 1.02)
not significantly elevated in the proximal counties

Conclusion: because of low exposure levels andecological nature of the study, does not adequately
address the health risks of uranium

All cancer: Karnes County RRs of 1.0 in 1950–1964(before/beginning of mining-milling), 0.9 in 1965–1979(early operations), 1.1 in 1980-1989 (later operations
and latency period) and 1.0 in 1990–2001 (few/no
operations)

Strengths: mortality ascertainment was high. A priori
cancer types specifically examined

A priori tumour sites: for prime exposure periods (1965–
1979 and 1980–1989), Karnes county RRs were 1.0, 1.2,
respectively, for lung cancer, 0.8, 0.9 for kidney cancer,
1.0, 0.8 for liver cancer, and 1.3 (n=20) and 1.7 (n=17)
for leukaemia. no RRs significantly elevated. Childhood
cancer mortality 1965-2001, non-significant RR of 1.3(n=8 cases)

Relevance: because of low, unknown exposure levels
and ecological nature of the study, does notadequately address the health risks of uranium

Weaknesses: no individual or even county-level
estimates of uranium exposure levels. Proximal county
areas were rather broad, further diluting possible
exposure, though nearly all inhabitants lived within
20 miles of a processing plant

Weaknesses: limited uranium exposure measurements
available to use in analysis. Cancer mortality
misdiagnoses, especially for liver cancer

UNSCEAR 2016 REPORT

Table A10. Studies of general population groups with potential environmental exposure to uranium by inhalation

Summary of study

Summary of findings related to uranium

Relevance for this report

Boice et al. [B34]

Mortality evaluated during 1978–2004 for 1936–1984
residents of Uravan, Colorado USA, a uranium mill
town. Mining and milling activities during mid-1930s to
1984. The mean follow-up time since first Uravan
residence, 38.1 years

All cancer: no significant elevation in overall cancer
mortality or cancers of lung, kidney, breast; leukaemia;
non-malignant respiratory, renal or liver disease
among females or the 622 uranium mill workers, but
excess lung cancer found among undergrounduranium miners. Had no quantitative information on
exposure levels of mill workers

Study uninformative regarding uranium effects because
no uranium exposure data and has low statistical
power

Boice et al. [B35]

Comparison of 1950–2000 mortality in Montrose County,
Colorado, USA (Uravan and other mining/milling
operations) with five comparison counties

All cancer: no difference in total cancer

Since no information on who was exposed to uraniumand exposure levels, study uninformative regarding
uranium risk

Cancer incidence (1982–2004) and mortality (1950–
2004) in Grants, New Mexico, USA residents: Grants
mining during early 1950s to 1990; milling operations
1958–1990

Lung cancer: found increased mortality from lung cancer
among men (SMR=1.57, 95% CI: 1.21, 1.99)

Boice et al. [B37]

A priori tumour sites: montrose elevation of lung cancer inmales (RR=1.19, 95% CI: 1.06, 1.33), thought to be due tounderground miner radon exposure and heavy smoking.
No excess of breast, kidney, bone, liver or childhood
cancer, leukaemia, non-Hodgkin’s lymphoma, renal
disease or non-malignant respiratory disease

Stomach cancer: stomach cancer mortality among
women was high (SMR=1.30, 95% CI: 1.03, 1.63), butelevated mainly in the early years before milling
operations began

Lung cancer excess was likely due to miner radon
exposure and smoking. Uranium exposed individuals
not identified, so it is uninformative

Chen et al. [C24]

Ecological study of cancer incidence in Port Hope,
Ontario, Canada residents, 1992–2007. In 1981–1982,
air uranium concentrations averaged 0.02 μg m-3
leading to a committed effective dose of 0.16 mSv, but
by 1988–1989 were reduced to 0.00105 μg m-3. Larger
doses received from gamma and radon exposure.
Population ~16,500

Standardized incidence ratio (SIR) for leukaemia, 0.86
(95% CI: 0.60, 1.21), with no elevation of childhood
leukaemia rate SIRs were <1.0 for a number of othercancer sites. A significant elevation of lung cancer
incidence, perhaps related to smoking habits

Provides some information on average air uranium
levels, but analyses were ecological and not specific foruranium exposure. Study useful insofar as it rules outlarge uranium effects

Report of theConsejo deSeguridad
Nuclear [C37]

Ecological study of cancer mortality in municipalities
near seven nuclear power plants and five fuel cycle
facilities (chemical conversion of uranium concentrate)in Spain. Cancer mortality (1975–2003) ofmunicipalities within 30 km of facilities compared tosimilar municipalities 50–100 km distantWith reconstructed external doses, reported increasing
dose-response trends for kidney cancer around nuclear
power plants and for lung and bone cancers around
fuel cycle facilities, but had not estimates of uranium
exposure

Report is uninformative regarding uranium exposure
effects, as it is an ecological study and only estimatedexternal radiation exposure

Lopez-Abente etal. [L46]

Examined solid cancer mortality (1975–1993) in 283
towns in Spain within 30 km of one of four nuclear
power plants or four nuclear fuel facilities, compared to
275 towns 50–100 km away, matched on various
sociodemographic variables

They concluded that lung cancer and kidney cancer
mortality rates were higher in the 30 km area, but othertypes of cancer were not

Various inconsistencies in the results depending on how
the analyses were performed. They chose analyses thatshowed positive effects. No information presented onuranium exposure levels, so study is uninformative
regarding uranium risk

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 463

Study referencesSummary of findings related to uranium

Relevance for this report

Lopez-Abente etal. [L45]

Examined lympho-haematopoieticmalignancies (LHMs)
in 489 towns within 30 km of Spain’s seven nuclear
power plants and five nuclear fuel facilities (“exposed”),
compared to 477 towns 50–100 km away(“unexposed”). Exposed towns reported 610
leukaemias, 198 lymphomas and 122 multiple
myeloma deaths during 1975–1993

No excess LHM found in towns near nuclear power
plants. Reported excess leukaemia mortality near two
nuclear fuel facilities and excess myeloma mortality
near one nuclear fuel facility. No exposed town showedexcess leukaemia for the under age 25 group. Analysesof all nuclear fuel facilities combined did not yieldstatistically significant excess for any end point(leukaemia, leukaemia <25 years, myeloma, Hodgkin’s
lymphoma, non-Hodgkin’s lymphoma)

Selecting a few “significant” results from a large numberof statistical tests is questionable. The results for all
nuclear fuel facilities combined, or all nuclear power
plants, do not indicate elevated risks for lymphohaematological malignancies

Pinney et al.
[P22]

Examined prevalence ratios of diseases among 8,496
residing within 2 miles (3.2 km) from Fernald, USAuranium plant, or within 5 miles (8 km) and ingroundwater runoff direction, or with well/cistern.
Medical conditions obtained by questionnaire andscreening examination. Prevalences were compared to
NHIS/NHANES data (national standardized surveys)

Reported a number of elevated prevalences of kidney
and bladder diseases/conditions compared to NHIS
data, but screening questions or coding sometimes
differed between the two datasets. Found nodifferences for diabetes, thyroid diseases or respiratory
diseases. Several clinical laboratory variables showedsmall but significant differences between those within
2 miles or more distant, and a different set of variables
were significant for those using wells/cisterns

Distance from the plant and/or possible exposure toplant runoff used as surrogates for uranium doses.
Actual measured exposure levels very low. Since
perceived residential risks from Fernald were current in
the population, results based on self reports may havebeen biased. Inconsistencies among comparisons oflaboratory findings create uncertainty in theinterpretation

Table A11. Summary of literature review on health effects of human exposure to uranium through ingestion of surface or groundwater
Study

Study design

Country

Mao et al. [M9]Cross-sectional

CanadaChemical toxicity
urinary system

Zamora et al.
[Z6]

Cross-sectional

CanadaKurttio et al.
[K26]

Cross-sectional

FinlandKurttio et al.
[K28]

Cross-sectional

Finland

EffectEffect measurement

Relevance for this reportofComparison of biomarker levels in urine (microalbuminuria)
and serum (creatinine)

Positive association between uranium cumulative
exposure index and albumin level

Chemical toxicity
urinary systemofComparison of biomarker (glucose, creatinine, protein, beta2microglobulin, alkaline phosphatase, γ-glutamyl transferase,lactate dehydrogenase, N-acetyl-β-D-glucosaminidase) levels

Alkaline phosphatase, beta2-microglobulin levels
correlated with uranium level in water

Chemical toxicity
urinary systemofComparison of biomarker (calcium, phosphate, glucose,
albumin, creatinine, beta2-microglobulin) levels

Significantly increased calcium, fractional excretion

Chemical toxicity
urinary systemofComparison of renal damage indicators (glucose, creatinine,alkaline phosphatase, γ-glutamyl transferase, lactate
dehydrogenase, N-acetyl-β-D-glucosaminidase, calcium,
phosphate, cystatin C, glutathione-S-transferase) in urine

No statistically significant association between
uranium concentrations in urine and any of therenal damage indicators, except glucose excretionin urine and diastolic blood pressure

No association between uranium exposure andother parameters

UNSCEAR 2016 REPORT

Summary of study

464

Study references

StudyStudy design

Country

EffectEffect measurement

Relevance for this report

Selden et al.
[S8]Cross-sectional

SwedenChemical toxicity
urinary systemofComparison of biomarker (albumin, beta2-microglobulin,
protein HC, kappa and lambda chains, N-acetyl-β-Dglucosaminidase) levels in urine

Significant increase in urinary excretion of β-2
microglobuline, kappa and lambda chains, and HC
protein with medium to high uranium
concentrations
inurine.
Dose–response
relationships observed after exclusion of subjects
with diabetes

Kurttio et al.
[K29]

Case-cohort

FinlandUrinary system as targetof radio-toxicity

Comparison of risk of bladder cancer by uranium level andradiation doseNo excess of bladder cancer with increased level ofuranium or radiation dose

Kurttio et al.
[K27]

Cross-sectional

FinlandBone as target
chemical toxicity

Correlation between uranium exposure and biomarkers
associated with bone (osteocalcin, aminoterminal
propeptide of type I procollagen, serum type I collagencarboxy-terminal telopeptide)

Marginal positive association of uranium
concentrations in drinking water with serum type I
collagen and carboxy-terminal telopeptide only inmen (p=0.05)

of

No significant association between uranium
exposure and bone turnover indicators in women
Seiler [S7]Ecological

USA

Comparison of uranium concentration in wells used by casefamilies and other wells

No significant difference between uranium
concentrations in wells used by families ofleukaemia cases (median=3.4 μg/L) and theuranium concentrations in other wells (1.6 μg/L)
No differences in concentrations of gross α activity
or of Rn (617 vs. 563 pCi/L)

Auvinen et al.
[A32]

Case-cohort

Finland

Lympho-haematopoietic
systemComparison of risk of leukaemia according to uranium level

No excess of leukaemia according to uranium levelof drinking water

Witmans et al.
[W23]

Case-control

CanadaLympho-haematopoietic
systemComparison of Uw and Thw exposure between cases andcontrols

Cases had higher uranium concentrations indrinking water than controls (p=0.001)
No significance difference in Th w (p=0.22)

Auvinen et al.
[A32]

Case-cohort

Finland

Digestive system

Stomach cancer risk according to uranium level

No excess of stomach cancer by uranium level

ANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 465

Lympho-haematopoietic
systemANNEX D: BIOLOGICAL EFFECTS OF SELECTED INTERNAL EMITTERS—URANIUM 467

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Abou-Donia, M.B., A.M. Dechkovskaia, L.B. Goldstein et al. Uranyl acetate-induced
sensorimotor deficit and increased nitric oxide generation in the central nervous system in rats.
Pharmacol Biochem Behav 72(4): 881-890 (2002).

A2

Adams, N. and N.L. Spoor. Kidney and bone retention functions in the human metabolism of
uranium. Phys Med Biol 19(4): 460-471 (1974).

A3

Al-Hadithi, T.S., J.K. Al-Diwan, A.M. Saleh et al. Birth defects in Iraq and the plausibility ofenvironmental exposure: A review. Confl Health 6(1): 3 (2012).

A4

Al-Hashimi, M.M. and X. Wang. Comparing the cancer in Ninawa during three periods (19801990, 1991-2000, 2001-2010) using Poisson regression. J Res Med Sci 18(12): 1026-39(2013).

A5

Al-Masri, M.S., A. Al-Hamwi, Y. Amin et al. Radionuclide transfer from feed to camel milk.
J Environ Radioact 132: 8-14 (2014).

A6

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unscear.org

This publication contains:
REPORT OF THE UNITED NATIONS SCIENTIFIC COMMITTEE ON THE EFFECTS
OF ATOMIC RADIATION TO THE GENERAL ASSEMBLY
Scientific Annexes
Annex A. Methodology for estimating public exposures due to radioactive discharges
Annex B. Radiation exposures from electricity generation
Annex C. Biological effects of selected internal emitters—Tritium
Annex D. Biological effects of selected internal emitters—Uranium
Supplementary materials (only available online at www.unscear.org)

E VA L U AT I N G R A D I AT I O N S C I E N C E F O R I N F O R M E D D E C I S I O N - M A K I N G
In 1955 the United Nations General Assembly established the Scientific Committee on the Effectsof Atomic Radiation (UNSCEAR) in response to concerns about the effects of ionizing radiation
on human health and the environment. At that time fallout from atmospheric nuclear weapons
tests was reaching people through air, water and food. UNSCEAR was to collect and evaluate
information on the levels and effects of ionizing radiation. Its first reports laid the scientific grounds
on which the Partial Test Ban Treaty prohibiting atmospheric nuclear weapons testing was
negotiated in 1963.
Over the decades, UNSCEAR has evolved to become the world authority on the global leveland effects of atomic radiation. UNSCEAR’s independent and objective evaluation of the science
are to provide for—but not address—informed policymaking and decision-making related toradiation risks and protection.

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