EFFECTS OF IONIZING RADIATION
United Nations Scientific Committee on the
Effects of Atomic Radiation
UNSCEAR 2006
Report to the General Assemblywith Scientific Annexes

VOLUME II
Scientific Annexes C, D and E

UNITED NATIONSNew York, 2009

NOTE
The report of the Committee without its annexes appears as Official Records of the General
Assembly, Sixty-first Session, Supplement No. 46 and corrigendum (A/61/46 and Corr. 1). The
report reproduced here includes the corrections of the corrigendum.
The designation employed and the presentation of material in this publication do not imply theexpression of any opinion whatsoever on the part of the Secretariat of the United Nations concerning the legal status of any country, territory, city or area, or of its authorities, or concerning
the delimitation of its frontiers or boundaries.
The country names used in this document are, in most cases, those that were in use at the time 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.09.IX.5
ISBN 978-92-1-142270-2

Corrigendum to Sales No. E.09.IX.5

May 2011

Effects of Ionizing Radiation: United Nations ScientificCommittee on the Effects of Atomic Radiation
2006 Report to the General Assembly, with Scientific
Annexes—Volume II
Annex E (“Sources-to-effects assessment for radon in homes and workplaces”),
paragraph 460
The ninth sentence should read
For residential exposure to 150 Bq/m3 , the authors estimated a combined ORof 1.1 (95% CI: 1.0, 1.3).

V.11-82963 (E)*1182963*

ANNEX E
Sources-to-effects assessment for radon in homes and workplaces
Contents
Page
INTRODUCTION. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 201
I.	

SOURCES AND LEVELS OF RADON EXPOSURES. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
A.	 Outdoors. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
B.	 Indoors . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
C.	 Workplaces. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
D.	 Measurements of radon and radon decay products. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .

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II.	 DOSIMETRY . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
A.	 Dosimetric models. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
B.	 Dosimetry of radon and thoron . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
C.	 Dose conversion factors . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
D.	 Uncertainties in dose conversion factors. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
E.	 Exposures in homes and in workplaces other than uranium mines . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .

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III.	 EXPERIMENTAL STUDIES . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 221
A.	 Animal experiments. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 221
B.	 Biomarkers . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 223
IV.	 EPIDEMIOLOGICAL STUDIES OF MINERS . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
A.	 United States: Colorado Plateau miners . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
1.	 Introduction. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
2.	 Radon and radon decay products. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
3.	 Exposure estimation. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
4.	 Other hard rock mining exposures. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
5.	 Epidemiological analyses. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
6.	 Evaluation . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .

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B.	 Canada: Ontario uranium miners. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
1.	 Introduction. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
2.	 Radon and radon decay products. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
3.	 Exposure estimation. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
4.	 Epidemiological analyses. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
5.	 Evaluation . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .

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C.	 Czechoslovak miners. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
1.	 Introduction. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
2.	 Radon and radon decay products. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
3.	 Exposure estimation. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
4.	 Epidemiological analyses. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
5.	 Evaluation . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .

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D.	 Swedish iron ore miners. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 247
1.	 Introduction. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 247
2.	 Radon and radon decay products. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 247
3.	 Exposure estimation. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 248
4.	 Epidemiological studies. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 249
5.	 Evaluation . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 249
E.	 Canada: Beaverlodge, Saskatchewan, miners. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 250
1.	 Introduction. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 250
2.	 Radon and radon decay products. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 251
3.	 Exposure estimation. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 254
4.	 Epidemiological studies. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 258
5.	 Evaluation . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 261
F.	 Germany: Wismut miners. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 261
1.	 Introduction. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 261
2.	 Radon and radon decay products. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 261
3.	 Epidemiological studies. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 262
4.	 Evaluation . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 265
G.	 Canada: Port Radium miner study. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 266
1.	 Introduction. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 266
2.	 Radon and radon decay products. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 267
3.	 Exposure estimation. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 268
4.	 Epidemiological studies. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 268
5.	 Evaluation . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 268
H.	 French uranium miners . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 269
1.	 Introduction. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 269
2.	 Radon and radon decay products. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 269
3.	 Exposure estimation. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 269
4.	 Epidemiological studies. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 270
5.	 Evaluation . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 271
I.	 Canada: Newfoundland fluorspar miners. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 272
1.	 Introduction. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 272
2.	 Radon and radon decay products. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 273
3.	 Exposure estimation. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 273
4.	 Epidemiological studies. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 274
5.	 Evaluation . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 276
J.	 Chinese miners. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 276
K.	 Australia: Radium Hill uranium miners. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 277
L.	 Overall evaluation of miner studies . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 277
V.	 EPIDEMIOLOGICAL STUDIES OF RESIDENTIAL EXPOSURES . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
A.	 Introduction. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
B.	 Case–control studies of residential radon. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
C.	 Ecological studies of residential radon. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
D.	 Overall evaluation of residential radon studies. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .

281
281
281
285
287

VI.	 EFFECTS OF RADON ON organs and tissues other than THE lung. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
A.	 Dosimetric considerations. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
B.	 Epidemiology of cancers other than lung. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .
C.	 Effects other than cancer . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .

295
295
297
298

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VII.	

Page
Implications for risk assessment. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 301
A.	 General studies of exposure uncertainty . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 301
B.	 Biologically based models . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 303
C.	 Risk projection. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 305

VIII.	 Overall Conclusions. .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 309
APPENDIX . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 311
REFERENCES . .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  .  . 313

199

INTRODUCTION
1. For many years, the Committee presented its evaluations of information and data on the “Sources” and “Effects”
of ionizing radiation separately. These were divided into
two volumes in the UNSCEAR 2000 Report [U2]. During
its 50th session, in April 2001, the Committee discussed thefeasibility of preparing documents in an integrated “Sourcesto-Effects” approach. This document provides a “test case”
for future UNSCEAR “Sources-to-Effects” assessments.

UNSCEAR 2000 Report [U2]. The Committee is presently
updating its assessments of these sources.
5. Historically, a wide variety of quantities and units wereused to assess radon exposure. Appendix A provides a shortsummary of the historical units used in this report and therelevant conversion factors. To maintain the integrity of thehistorical data, the original units used in the papers cited inthis annex are preserved. Most historical, and indeed ­current,
measurements of radon in mines are in units of working
level (WL) or working level month (WLM). The former is a­measure of the potential alpha energy concentration in the air
and the latter is a measure of exposure for an assumed working month of 170 hours. The unit of WLM is the traditional
unit used to report exposure to RDPs in studies of ­miners.
The modern quantity for expressing concentration of RDPs
is the equilibrium-equivalent concentration (EEC), which
represents the concentration of 222Rn in equilibrium withits decay products that would have the same potential alphaenergy. To convert from an exposure in WLM, it is necessary
to multiply by 6.4 × 105 (see appendix A) to obtain Bq h m–3(EEC). Where appropriate, the RDP measurement expressed
as EEC is provided in brackets or is otherwise discussed.

2. Radon is an inert noble gas. Its most common isotope,
and the one that is commonly known as radon, is 222Rn, whicharises in the radioactive decay chain of uranium-238. Uranium
occurs naturally in varying levels in all rocks and soils. Somefraction of the radon produced in rocks and soils escapes tothe air; therefore radon is present in the atmosphere. Thus,
simply by breathing, people everywhere are exposed to radiation from radon itself and also from short-lived radon decay
products (RDPs).1 Moreover, radon is ­soluble in water, andgroundwater that passes through uranium-bearing soils androcks contains radon. When radon-rich groundwater is usedas drinking water, people are exposed both through water
consumption and by radon being released from the water tothe air and being inhaled.
3. Thoron (220Rn) is an isotope of radon, and therefore
also an inert noble gas, which arises from the decay chainof thorium-232 (232Th). Thorium is a common element in theearth’s crust and therefore, like radon, thoron is found in air
at varying concentrations. Since radon and thoron are members of different decay chains, the ratio between radon andthoron (or between the decay products of radon and thoron)
will depend in part on the ratio of uranium to thorium inlocal soils, rocks or building materials. As discussed in the
UNSCEAR 2000 Report [U2], the radioactive half-lives ofradon and thoron and their respective decay products arealso very important in determining the exposures of people
in workplaces and homes. Since thoron has a much shorter
half-life (t½ = 55 s) than radon (t½ = 3.82 days), the distanceit can travel before undergoing radioactive decay is verymuch shorter than the distance radon can travel in the same
medium, and therefore its expression in the environment isquite different from that of radon.

6. For many years, radon was recognized as constituting
a hazard to underground miners. However, while it was also
recognized that domestic exposure to radon might carry arisk, there was no direct evidence of this until recently. The
risk from residential exposure to radon and its decay ­products
is of great interest in many countries; thus ­methods to evaluate exposure to radon and thoron and their decay products
as well as the subsequent risks from exposure are of greatinterest.
7. In the past, exposures to thoron and its decay products
were often ignored. As will become evident from the discussion in section I, it has become increasingly clear that the
exposure to thoron and its decay products cannot be ignoredin some environments (both workplace and residential) as it
contributes to the risks otherwise assigned solely to exposure
to radon and its decay products (e.g. [S41]). In some epidemiological studies, no distinction is made between exposure
to radon and its decay products and exposure to thoron andits decay products. Measurement techniques for discriminating between them do, however, exist (e.g. [C21, T12, T14,
T15, T16, T17, Z3, Z4]).

4. Sources of radon and thoron and of potential exposures
to workers and the public are briefly discussed in section I.
A more comprehensive discussion of sources of exposuresto radon and thoron in the workplace is provided in the

8. The use of recent measurements to estimate exposures
received many years ago, for example in the uranium minerstudies and in the studies of residential radon exposures,

1

The term “radon” is used generically in this report to indicate both radon and its
decay products, the latter in fact contributing most of the dose to lung tissue.
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UNSCEAR 2006 REPORT: VOLUME II

carries particular difficulties. In the case of mining, ore
grade, mining methods, ventilation practices and otherfactors have changed over time. Residential radon studies
experience similar problems in estimating past exposures,
because of, for example, changes in heating and ventilation
practices over time. It is thus important to recognize factorsthat have a substantial impact on exposure estimates and toassess the potential magnitude of these impacts. For example, the entry of radon into structures is an important consideration, is well studied and is described in many reports,
including ­references [U2, U5, U6].
9. To understand how radon exposure estimates can betransferred from one miner population to another, or fromconditions in mines to conditions in homes, it is importantto understand the differences in dosimetry of exposures inmines compared with homes. The 1991 National ResearchCouncil companion report [N10] to the report of the BEIR IV
Committee [C19] provides a comprehensive discussion ofthese differences, as do BEIR VI [C20] and the UNSCEAR
2000 Report [U2]. Section II of this annex provides a concise
overview of current issues in radon and thoron dosimetry.
10. An understanding of the mechanisms for the carcinogenicity of radon and its decay products and of how radon
interacts with other agents is important. Much information isavailable on these topics (e.g. [C20, N11, U2]). Informationfrom animal experiments and experiments at the cellular andsubcellular levels relevant to understanding the mechanismsof radon carcinogenicity is discussed briefly in section III.
11. Until recently, the main basis for estimating risks fromresidential exposure to radon and its decay products wasprovided by epidemiological studies of underground minersthat extrapolated the results down to the levels of exposure
seen in homes. Studies of historical miners require retrospective estimation of exposure conditions many years in thepast. Often there are few or even no actual measured exposure data from the early years of mining; the results fromsuch studies are less certain because of this and other factors. Extrapolation from risks estimated in miner studies toresidential exposure conditions involves additional assumptions. For example, such extrapolations must consider the
impact of the relatively short exposures at high exposure
rates seen in mines compared with the longer exposures atlower exposure rates in homes. The different dosimetry of
exposures in mines and in homes, as well as other factors,
must be considered in selecting the most appropriate exposure– (or dose–) response model for extrapolation down toresidential levels.
12. Consequently, direct estimates of residential risk fromthe exposure of the general population are of great interest,and numerous studies of risk from residential exposure toradon were made using case–control studies. The resultsof these studies demonstrate an excess risk at the levels ofradon seen in homes and suggest a pattern of increasing risk
with increasing exposure that is generally consistent with theexperience of epidemiological studies of miners. Section IV
discusses the epidemiological studies of miners exposed toradon and section V discusses the epidemiological studies ofresidential exposure to radon.
13. While an increased risk of lung cancer associated with
exposure to radon and its decay products is well establishedfrom epidemiological studies of underground miners andmore recently from residential radon studies, the potential
risks to tissues and organs other than the lung are also ofinterest and are the subject of section VI.
14. There is great interest in predicting future risks to­people who are exposed to radon and its decay products
either in the workplace or in the home. It is necessary tounderstand the limitations of risk projections and how suchprojections may be affected by consideration of exposures
to other agents (cigarette smoke being the most important)in the workplace or in the home. The characteristics andlimitations of existing models, including biologically based­models, and a recommended approach to risk projection are
the subjects of section VII.
15. The reliability of estimates of radiation exposures is
one important factor in assessing the risk of ­cancer following
radiation exposure. The sources of uncertainty in epidemiological studies and how these uncertainties affect the dose–
response analysis [N5] was discussed in the UNSCEAR
2000 Report [U2]. Uncertainties arise in estimating exposures of miners at work, historically reported in WLM, andof people at home, reported in Bq h m–3.
16. The sources and characteristics of uncertainty in the
exposure of miners are of considerable interest and constitute a major focus of this report. The report of the BEIR VI
Committee [C20] discusses this subject at length, as does themost recent radon report of the National Council on Radiation Protection and Units (NCRP) [N11], which provides themost comprehensive examination to date of the undergroundminer data. There are large difficulties in developing reliable estimates of underground radon exposure for epidemiological studies of miners, especially for the pre-1960 miners[C20, L10]. Studies of miners employed more recently haverelatively reliable exposure information (e.g. [H35, S12,S14]). Uncertainties in assessment of exposure are similarly
important considerations in residential radon studies (e.g.
[D15, D17, L8]).
17. Finally, section VIII provides an overall summary of
the main observations from this annex.

I.

SOURCES AND LEVELS OF RADON EXPOSURES

18. The majority of the dose to the lung arises from exposure to the short-lived decay products of radon and thoron.
Concentrations of the potential alpha energy of these shortlived decay products are estimated by considering the stateof equilibrium between the parent nuclides and their respective decay products. In practice, an equilibrium factor Feq isused to characterize the state of equilibrium. The equilibrium factor Feq is defined as the ratio of the actual potentialalpha energy concentration (PAEC) to the PAEC that would
prevail if all the decay products in each series were in equilibrium with the parent radon or thoron, as the case may
be. However, as discussed in reference [U2], it is simpler
to evaluate this factor in terms of an equilibrium-equivalentradon or thoron concentration.

thoron in Japan. At night and in the early morning hours,
atmospheric (temperature) inversion conditions are often
found; these tend to trap the radon closer to the ground. Thismeans that outdoor radon concentrations can vary diurnallyby a factor of as much as 10. Seasonal variations, related
to the effects of precipitation or to changes in prevailingwinds, also exist [U2]. An evaluation of exposure to outdoor
concentrations of radon in Iowa and Minnesota [S40] in the
United States concluded that outdoor exposure to radon insome areas can be a substantial fraction of an individual’s
exposure to radiation and moreover is highly variable acrossthe population. Outdoor levels of radon provide a baseline
for indoor levels of radon. In tropical climates, indoor andoutdoor concentrations are essentially the same because of
rapid exchange between indoor and outdoor air [C43].

19. The Committee customarily reports concentrationsof radon or thoron (and also the lung dose) in terms of­equilibrium-equivalent concentration (EEC). This is definedas the equivalent concentration of the decay products inequilibrium with the parent gas that yields the same potentialalpha energy per unit volume as the existing mixture.

22. The UNSCEAR 2000 Report [U2] suggests that typical outdoor levels of radon and thoron gas are each of theorder of 10 Bq/m3. There is, however, a wide range of longterm average concentrations of radon, from approximately
1 Bq/m3 to more than 100 Bq/m3, with lower levels typical of isolated small islands or coastal regions and higher
levels typical of sites with high radon exhalation over large
surrounding areas. Although data are relatively sparse forthoron, considerable variability from place to place would be
expected because of thoron’s short half-life, which amplifies
the effect of local variations in exhalation rate. Thoron decay
products were measured continuously outdoors in suburban
New Jersey, United States, for a 2-year period [H20]. The
average outdoor concentration of 212Pb was 0.09 Bq/m3 andthe variability over seasons was a factor of 2. Bismuth-212
was not detectable. The average outdoor concentration ofthoron gas was 15 Bq/m3, yielding a value of 0.005 for theequilibrium factor outdoors [C21].

EEC(222Rn) = 0.105 (218Po) + 0.516 (214Pb) + 0.379 (214Bi)
EEC(220Rn) = 0.91 (212Pb) + 0.087 (212Bi)
where 218Po, 214Pb, etc., and EECs are in Bq/m3. Older publications often report activity in curies (Ci). The Appendix
presents relevant conversion factors.
20. Radon and thoron are ubiquitous in the air at groundlevel and are significant contributors to the average dose fromnatural background sources of radiation [U2]. In homes, inunderground mines and in other situations where radon (andthoron) may be present and where ventilation may be limited, the levels of these radionuclides and their decay products can accumulate to high levels. Soils and rocks are often
the main sources of radon. In unsaturated soils or rocks,
radon moves with air through pores and fractures. In saturated zones, radon moves with groundwater to undergroundopenings, such as mines and caves, and to buildings [N9].
A.

B.

Indoors

23. In buildings with high radon levels, the main mechanism
for entry of radon is pressure-driven flow of soil gas throughcracks in the floor. This arises because the air inside buildings
is normally at a slightly lower pressure than the air outdoors.
This underpressure is the consequence of the air inside thebuilding being warmer than that outside. In temperate zones
especially, this causes a convective flow (“chimney effect”),
which, together with the effect of the wind blowing over chimneys and other openings (“Venturi effect”), draws soil gas andhence radon into the building. However, in addition to pressure
differences, other factors, including relative humidity and soil
moisture, can also influence radon levels in buildings [S66].

Outdoors

21. Concentrations of radon in the outdoor environment are
affected not only by the magnitude of the release rate fromthe ground to the atmosphere but also by atmospheric mixing phenomena. Solar heating during the daytime induces
turbulence, so radon is more readily transported upwards andaway from the ground. Doi and Kobayashi [D18] provide
information on the vertical distribution of outdoor radon and
203

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UNSCEAR 2006 REPORT: VOLUME II

24. While most building materials produce some radon,certain materials can act as significant sources of indoor
radon. Such materials have a combination of elevated levelsof 226Ra (the radioactive parent of radon) and a porosity thatallows the radon gas to escape. Examples are lightweight
concrete with alum shale, phosphogypsum and Italian tuff.
25. Groundwater, particularly in granitic areas, can havehigh levels of radon. Workplaces such as laundries and restaurant kitchens can have high radon levels from the use ofsuch water. Because municipal water supplies are often fromrain catchment surface reservoirs, radon levels in public
water supplies are normally not high, and any problems arenormally limited to wells in geological formations containing naturally elevated levels of uranium. In the United States,
groundwater supplies were reported to contain radon at ­levelsof as high as 106 Bq/m3 or more [N9]. In Germany, treatment
and distribution workplaces for groundwater ­supplies also
were found to contain elevated radon concentrations in air,with up to several hundred thousand becquerels per cubic
metre [U2].
26. There is a considerable amount of data available onradon concentrations in indoor air, and new information is
becoming increasingly available on thoron concentrations
indoors. Substantial compilations of radon measurements
appeared in the UNSCEAR 2000, 1993 and 1988 Reports[U2, U5, U6]. On the basis of current data, the UNSCEAR
2000 Report [U2] deduced values of 40 and 30 Bq/m3 for thearithmetic and geometric means of indoor radon gas concentrations worldwide, with a geometric standard deviationof 2.3. On the basis of the information collected for thepresent report, these values are still appropriate.
27. Published literature provides additional insight into thesources and levels of radon and thoron. Examples include
publications on the assessment of radon in drinking water[N9, T22] and numerous papers on residential radon levels
in Iowa, United States [F3, S40], Italy [B15, F4, P6], Poland[Z9], the former Yugoslavia [Z2], China [W13, W16], theIsle of Man [G6] and the Republic of Korea [C33], among
others. The influence of groundwater on radon concentrations in drinking water and subsequently on residential radonlevels in air from the use of potable water indoors was also
investigated in radon-prone areas of Japan [K27, I6, N17].
28. A great deal of information is available from radonmeasurements in homes in the United Kingdom. A national
survey carried out in the early 1980s provided results of
gamma ray dose rate and radon concentration measurements in more than 2,000 dwellings selected systematicallyaccording to postal codes [W14]. The mean radon levelfrom the survey, after adjustment for dwelling types, was20.5 Bq/m3. The same study also reported data from studies
carried out in regions of the United Kingdom with elevateduranium mineralization. The regional surveys showed average radon levels of up to 300 Bq/m3 for areas of south-west
England, which are about 15 times the national average.
In addition, simultaneous measurements of thoron decay

product levels in high-radon regions showed a mean thoron
decay product level of about 0.6 Bq/m3, though the national
average was estimated to be 0.3 Bq/m3 [W14]. Subsequently,
more than 400,000 measurements were made throughout theUnited Kingdom aimed at identifying dwellings with elevatedradon levels; this information is available for England andWales [G10], Scotland [G11] and Northern Ireland [G12].
29. A nationwide indoor radon survey carried out since1982 in France has updated estimates of the population
exposure [B38]. Indoor radon measurements were performed using passive dosimeters left in place for 2 months
in the main room. A questionnaire was completed regarding housing characteristics. In total, the survey included
12,261 radon measurements distributed over the wholecountry. Corrections for seasonal variations [B36] andhousing charac­teristics were applied. The crude average ofindoor radon concentrations was 89 Bq/m3, and the average corrected for season and housing characteristics was
83 Bq/m3 (the range over French districts was 19 to297 Bq/m3). Weighting by district population density yieldeda national average of about 63 Bq/m3.
30. Bochicchio et al. [B40] report a national radon surveyof some 5,631 dwellings distributed in 232 towns acrossItaly. The authors report a national population-weighted
average of 70 Bq/m3 and a geometric mean and geometricstandard deviation of 52 Bq/m3 and 2.1, respectively. The
authors also report seasonal differences. On a national scale,over the 21 regions studied, the authors report a winter radon
to summer radon ratio with a geometric mean and standard
deviation of 1.23 and 1.71, respectively.
31. European Union (EU) efforts to develop a radon map
of European countries are under way as part of an EU project
to develop a European atlas of natural radioactivity [D20].
32. A major survey of radon concentrations in United Stateshomes measured radon levels in homes in 125 counties in
50 states with the results shown in figure I (after reference[M1]). Approximately 6.1% of the homes surveyed exceeded
the United States Environmental Protection Agency (EPA)
action level of 148 Bq/m3 (4 pCi/L). This study indicatedthat the distribution of radon concentrations in homes could
be reasonably described by a log-normal distribution. The
overall geometric mean (median) of the radon concentration
data was 25 Bq/m3 with a geometric standard deviation of
about 3.1.
33. Information on radon levels in Latin American countries is also becoming available (e.g. [C41, M41, S65]).
According to reference [C41], most indoor radon levels in
Latin America are below 100 Bq/m3. However, levels canbe quite variable in different regions within a country, as
evidenced by data for 17 states in Mexico, which reportmean radon levels across the 17 states surveyed ranging
from less than 40 Bq/m3 to near 200 Bq/m3. The overallmean value reported across the 17 states surveyed was
111.6 Bq/m3.

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

Figure I. Distribution of indoor radon levels in homes in the
United States (after reference [M1]).
1 000

EPA* (148 Bq/m³, 4 pCi/L)
100
RADON CONCENTRATION (Bq/m³)

Mean (46 Bq/m³, 1.25 pCi/L)Median (25 Bq/m³)

10

LEGEND:All housing units
Single-family housing units
Multifamily housing units

1

NOTE:
* EPA’s recommended remedialaction level for indoor radon
0.1
1 2

5

10

20 30 40 50 60 70 80

90 95

98 99

CUMULATIVE PROBABILITY (%)

34. A study of indoor and outdoor radon levels in Brazil
reported that a variability in radon levels of about 50% in asingle day could be measured on an hourly basis, with thehighest values in the morning and the lowest values in theafternoon. The authors also reported a seasonal variability of
two orders of magnitude in outdoor levels, with the highest
levels in the dry winter season and the lowest levels in thewet summer months [M41]. The mean indoor radon concentration within the urban area of Rio de Janeiro is 40 Bq/m3,
while the indoor radon concentrations within urban and rural
areas in Poços de Caldas are 61 and 204 Bq/m3, respectively[M41].
35. Several Japanese studies, including two nationwide
indoor radon surveys, showed significant differences [F14,F15, S64]. The first survey of more than 7,000 dwellings,
using passive radon detection, reported a mean radon concentration of 20.8 Bq/m3 [F14]. To investigate potential confounding by thoron, a survey of 900 dwellings used a detector
that could discriminate between radon and thoron. The meanradon concentration from the second survey was 15.5 Bq/m3[T12]. The authors noted a gradient in thoron concentrationin the homes, the concentration decreasing by nearly a factorof 2 with increasing distance, up to about 1.5 m, from walls
made of soil. Radon concentrations are reported as generally
homogeneous in rooms, except for places near the inlet andoutlet of indoor airflow [Z8]. This result should be considered when choosing locations for detection.

205

36. In the 1980s and 1990s, two nationwide surveys [R12,
W21] of indoor radon decay products were performed
in China using various grab sampling measurements(e.g. scintillation cell method and two filter methods) or
short-term measurements (2–4 days of activated charcoal
adsorption). The arithmetic mean (AM) of the indoor radon­concentration was 24 Bq/m3 and the geometric mean (GM)was 21 Bq/m3 [U2]. Since 1996, China’s housing situation has greatly changed. A new survey of indoor radonin 26 cities and regions during 2001 and 2005 used alphatrack detection (exposure period of 3–6 months) [S67]. The
AM of indoor radon concentration was 43.8 ± 37.7 Bq/m3
and the GM was 34.4 ± 1.95 Bq/m3, with the median of theconcentration at 32.9 Bq/m3. These results are significantly
higher than those of previous surveys. The radon concentration in some high-rise buildings exceeds the national
action level of 400 Bq/m3.
37. Traditional soil-brick and mud-wall structure houses
are still popular in the countryside of China. At least 100 million people are currently living in such houses. Surveys ofthoron and its decay products in Chinese traditional soil
structure houses have been conducted. Fan et al. observed
a thoron EEC of 4.0 Bq/m3 (n = 56) in 1991 in Shaanxi soilhouses [F20]. More recently a survey using improved alphatrack monitors was completed in Guangdong, Gansu, Yunnan, Xinjiang, and Guizhou provinces: thoron and its decay
product concentrations (AM) were 318 Bq/m3 and 3.8 Bq/m3(n = 148, maximum value 15.8 Bq/m3), respectively [S68].
Applying a dose conversion factor of 40 nSv/(Bq h m–3), theaverage annual effective dose from inhalation of 220Rn was
estimated to be in the range 1.1–4.4 mSv. The dose contribution from thoron is obviously significant in this situation.
38. Surveys of radon and thoron decay products were­carried out in Fujian province in China, where the naturallevels of 238U and 232Th in soil are elevated (about 53.4 Bq kg
and 116.8 Bq/kg, respectively) [Z5]. Radon and thoron
decay products were measured in homes and various otherlocations, including outdoors, in mineral processing plants,
underground and in railway tunnels. The average radon andthoron decay product concentrations (expressed as EEC)
measured in 204 dwellings were 12.9 Bq/m3 and 0.87 Bq/m3,
respectively. The mean outdoor level of radon decay productconcentrations (expressed as EEC) derived from measurements made at 180 locations was 9.69 Bq/m3. The authors
also assessed potential exposure of people living in Fujian
province, and concluded that thoron (and decay products)
contributed about 20% of the estimated effective dose of
1.28 mSv/a from radon and thoron combined.
39. Measurements of outdoor (1,100 samples) and indoor(1,050 samples) levels of radon and thoron decay products
were made at a university campus in northeastern Japan overa period of 4 years [K20]. The authors analysed the datausing a variety of statistical methods. A seasonal variation
was observed, with higher radon concentrations for autumn
and winter, both outdoors and indoors. A seasonal variationof thoron levels was not as clear.

206

UNSCEAR 2006 REPORT: VOLUME II

40. Shang et al. [S67] report the results of a countrywide
survey of 3,098 homes in China carried out with alpha track
detectors between 2001 and 2005 by the Chinese National
Institute for Radiological Protection. Measurement sites were
statistically distributed in the municipalities (Beijing, Shanghai and Tinjing) and 15 provinces, using a sampling rate of
0.09 in 10,000 houses across the country. The authors reportan arithmetic annual average radon concentration of 43.8 +37.7 Bq/m3 and a geometric mean of 34.4 + 2.0 Bq/m3, andnote that about 6.4% of the measured houses had levels ofabove 100 Bq/m3, although only 0.7% were above 200 Bq/m3.
The authors also report that in dwellings made of soil or mud,
exposure from thoron may be significant, and that a study ofthoron has been designed.
41. Additional data on indoor levels of thoron in Europe
and Asia are also reported in reference [S41], which notesthat the doses from thoron and its short-lived decay products can be comparable to, or even larger than, the dosefrom radon and its short-lived decay products. Others also
reported data on thoron and thoron decay products for someareas of East Asia (e.g. [T21, Z5]).
42. It was not practical to calculate lung dose directlyfrom thoron gas measurements because the equilibrium
factor (Feq) between the gas and decay products was notwell established. Past dose estimates for thoron were mademainly from filtered air measurements of the thoron decay
product 212Pb. However, much work on methods of measuring radon and thoron and their decay products was carried out in Japan (e.g. [T12, T13, T14, T15, T16, T17, T18,
T19, T20, T34, T41, Z3]). A Japanese study [I15] observedthat it was important to consider the influence of thoron on
the measure­ment of RDPs. A great deal of work was also
carried out in Japan concerning the measurement of radon
and thoron and their decay products, including passive andcontinuous systems, and the measurement of various factors(e.g. Feq and the fraction fp of decay products attached to particulates) which are important in assessing lung dose [K28,
T42, T43, T44, T45, T46, T47, T48). Shang et al. also investigated the effect of thoron on the measurement of RDPs in
Chinese cave dwellings [S69]. Since thoron will be present
in many homes, if a detector is used that responds to thoron
and radon without distinguishing between them, the contribution from thoron will be attributed to radon. For example,
in testing the sensi­tivity of one thoron monitor, the averagethoron level in 20 homes in southern India was 168 Bq/m3[Z3]. More detailed information on the average thoron levelsin these homes is provided in reference [T33].
43. Evaluation of exposure to radon and thoron and theirdecay products thus must take account of the actual activity
concentrations of the various alpha-emitting radionuclides
from the two series in the air that is inhaled. As noted previously, the total alpha particle energy yet to be released by
the decay of inhaled radon or thoron is reported in terms ofpotential alpha energy concentration (PAEC), with units ofeither J/m3 or WL (working level). This quantity can be calculated once the activities of the individual radionuclides are

determined. In most cases, the individual activities are notdirectly measured. Thus the exposure rate must be indirectly
determined using assumptions on concentration ratios, i.e.
equilibrium factors, which lead to the determination of the
EEC. The environmental factors that influence concentration
ratios in each of the radioactive series are of great ­significancefor assessments of both exposure and dose [M33, U2].
44. Many measurements of RDPs have been reported.
These suggest that a rounded value for the equilibrium factorof 0.6 may be appropriate for the outdoor environment [U2].
Ramachandran and Subba Ramu [R6] reported variationsin indoor equilibrium factors. The UNSCEAR 2000 Report
also noted that there is a wide range of values from individual measurements. This is understandable given the manyenvironmental factors, including exhalation rates and atmospheric stability conditions, that influence the various activity
ratios. The range of the equilibrium factor for outdoor radonis from 0.2 to 1.0, indicating a high degree of uncertainty inthe application of a typical value to derive EECs [U2].
45. Measurements of both thoron and radon gas using
passive alpha track detectors have been reported (e.g. [B16,
D11, G13, I8, I9, L37]). Measurements were obtained over a
2-year period for both thoron gas and its 212Pb decay productin four locations — three indoors and one suburban ­outdoor
location [C22, H20, H21]. Tokonami et al. [T48] provide
information on the contribution from thoron for several
radon detectors. The authors also provide a comparison ofsmall indoor surveys in the Gunna prefecture in Japan andin Kovagoszlos, Hungary. The average radon/thoron ratio
from the Japanese survey was 1.3, compared with a meanof 4.5 in the Hungarian survey. The authors concluded thatmeasurements without discrimination of radon isotopes
have the potential to affect risk estimates. Sugino, Tokonami
and Zhuo [S70] report radon and thoron concentrations in
offices and dwellings of the Gunna prefecture. The averageradon concentrations in offices were about 29 Bq/m3, higher
than the 17 Bq/m3 reported for dwellings. As for reference[T48], the mean ratio of thoron and radon was estimated as
approximately 1.3.
46. The UNSCEAR 2000 Report [U2] assumed average
equilibrium factors obtained from 212Pb/220Rn ratios of 0.003
outdoors and 0.02 indoors to derive estimates of a dose­conversion factor for thoron EEC and recommended a valueof 40 nSv/(Bq h m–3) for dose estimation.
47. Determinations of the equilibrium factor for radonindoors generally confirm the typical value of 0.4 previously
assessed by the Committee [U5, U6]. While indoor measurements show a range from 0.1 to 0.9, most are within 30% of
the typical value of 0.4 [H22, R6]. A study [H22] in seven
North American houses showed that the equilibrium factor
varies significantly with time, typically by a few tens of per
cent. Measurements were carried out over a 4-year period togain an understanding of the characteristics of radon and its
decay products in air-conditioned office buildings in Tokyo,
Japan. The equilibrium factor was evaluated during working

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

hours and over the whole day in this survey; there was ­little
difference between the two conditions. These values were
within 30% of the assumed typical value of 0.4 [T34].
Although the measurement of radon gas concentration may
be a surrogate for direct measurement of the decay ­productconcentration in the determination of exposure, EECs or
PAECs estimated using this assumed typical value may be inerror, frequently by several tens of per cent, though rarely by
as much as a factor of 2.
48. More caution should be exercised in assuming the
average values of the equilibrium factor for dose assessment from inhalation of thoron decay products. An objection
to the use of thoron gas measurements for dosimetric purposes is that thoron may not be well mixed in the indoor air
because of its short half-life. As indicated previously, somedata indicate that indoor thoron concentrations vary with the
distance from walls and floors [Z8]. In many samples, thethoron concentrations in the centre of the room or more than
1 m from the surface of building material containing 224Ra
were as low as in outdoor air, while the thoron concentration near the surface of the building material was more than
10 times that in the centre of the room. Only where a room
fan is used would thoron be well mixed and a large variation of the thoron concentration in the room not be found[M40].
49. Equilibrium factors for estimating 220Rn EEC
are given above as single values. Because of the large­spatial variations in thoron concentrations in a room,these ­single values should be regarded as being subject
to large uncertainties. Thus the use of an equilibrium
factor for thoron should be limited to situations wherelarge spatial ­variation is not found. On the basis of theresults of simultaneous measurements of thoron concentration and thoron EEC by long-term passive methods,
Yonehara et al. were unable to find a good correlation
between thoron concentration and the EEC [Y9]. However, Yamada et al. were able to find good agreement in astudy of 265 cave ­dwellings [Y10].
50. Although only a limited number of measurementsof thoron in indoor air are available, several investigations
reported both radon and thoron EECs. While acknowledging
the uncertainty noted earlier, this allows some generalizations to be made from the derived ratios of the radon andthoron EECs. On the basis of the physical characteristics ofradon and thoron, model entry rates to buildings and a ventilation rate of 0.7 h–1, the International Commission on Radiological Protection (ICRP) estimated expected concentrations
in buildings [I5], which in terms of EEC are 2–50 Bq/m3 forradon and 0.04–2 Bq/m3 (mean = 0.5 Bq/m3) for thoron. This
corresponds to a thoron/radon EEC ratio of 0.03 [U2].

207

51. Table 1 provides a summary of concentrations ofradon in indoor air determined from surveys. Many of thedata in table 1 are from annex B of the UNSCEAR 2000
Report [U2]. Data carried forward from reference [U2] areindicated as such. Original sources of data in [U2] are fromUNSCEAR surveys of Natural Radiation Exposure andliterature as cited in [U2]. Since the publication of [U2],
the Committee has carried out three surveys of the naturalradiation environment, in 2001, 2004 and 2006. Summary
results from these surveys are provided in table 1. Not surprisingly, many of the indoor radon data were collected forareas or regions where indoor radon levels were thought
to be ­elevated. Thus, where the survey results provided to
the Committee make such a judgement possible, a note isgiven in the last column of table 1 indicating whether thedata provided in the table may be considered as some formof national average or whether they are more indicative ofa regional or local area of elevated levels. In some instances
the national authorities provided this judgement, while inothers the judgements were made by the Committee. In all
instances, precedence was given to data provided to theCommittee, as opposed to literature values. The data in
table 1 show considerable variability both within countries
and from country to country, with, for example, reported
nominal geometric mean indoor levels ranging from
<10 Bq/m3 in Egypt and Cuba, to more than 100 Bq/m3 in a
number of European countries, and to above 600 Bq/m3 inparts of the Islamic Republic of Iran.
52. The UNSCEAR 2000 Report [U2] gives worldwide
arithmetic mean values of 46 Bq/m3 (unweighted) and39 Bq/m3 (population-weighted). Worldwide geometric meanvalues of 37 Bq/m3 (unweighted) and 30 Bq/m3 (populationweighted) with corresponding geometric standard durationof 2.2 (unweighted) and 2.3 (population weighted) are alsopresented in [U2]. Given the wide variety and disparity ofdata currently available, no attempt was made to update thenominal values from [U2], and the values provided in [U2]are retained for the purposes of this report.
53. The UNSCEAR 2000 Report [U2] gives an annual
per caput dose estimated at 1.15 mSv from exposure tonatural sources of radon. This value is still appropriate.
The UNSCEAR 2000 Report also gives an annual dose of
0.1 mSv from natural sources of thoron [U2]. While thisvalue is still reasonable, data collected for the present study
indicate that the levels of thoron (and hence doses fromexposure to thoron and its decay products) are highly variable and that thoron may provide a larger contribution tonatural background dose than previously thought. Dosesfrom radon and thoron represent approximately half ofthe estimated dose from exposure to all natural sources of­ionizing radiation.

208

UNSCEAR 2006 REPORT: VOLUME IITable 1

Concentrations of radon in indoor air

Region/country

Population(10 6)

Indoor radon (222Rn) (Bq/m3)
Arithmetic
mean

Geometricmean

Maximum
value

Notes
Geometric
standard deviation

AfricaAlgeria [U2]

28.78

30

140

–

Egypt [U2]

63.27

9

24

–

Ghana [U2]

17.83

340

–

North America
Canada [U2]

29.68

34

14

1 720

3.6

Canada [L47]

32.27

28.35

11.2

1 720

3.9

Mexico107.03

140

90

1 193

United States [U2]

269.4

46

25

[M39]
3.1

National survey [M1, U18]

2Countrywide averagee

South America
Argentina

38.75

35

Brazil186.40

81.95

310.0

[C41]

Chile [U2]

14.42

25

86

–

Cuba

11.20

7.7

Ecuador

25

5.2

211

15.3

3.3

200

Countrywide averagef[Z11]

Paraguay [U2]

4.96

28

51

–

Peru

27.97

32.29

50.20

[C41]

Venezuela

26.75

52.50

346

[C41]

National average based on
sampling 3098 dwellings [S67]East Asia
China1315.84

43.8

34.4

596

China [U2]

1 232

24

20

380

 Hong Kong SAR
[U2]

6.19

41

Taiwan

22.89

10.0

8.5

63.5

0.6

Countrywide averagef

India [U2]

944.6

57

42

210

2.2

–

Indonesia

213.67

35.1

35.1

165

1.2

Countrywide averagee

Japan [U2]

125.4

16

13

310

1.8

–

Kazakhstan

14.83

Republic of Korea

48.85

53.4

Malaysia [U2]

20.58

14

20

–

Pakistan [U2]

140.0

30

83

–

Philippines

75.90

23

22

62

1.13

Countrywide averagee

Philippines

76.57

23

23

62

±6

Countrywide averageg

Russian Federation
Thailand [U2]

2.2

140

–

Countrywide averagee

5 000
43.3

1 350

1.8

50–60
58.7

23

–

Countrywide averagee

[Z11]
16

480

1.2

–

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

Region/country

Population(10 6)

Indoor radon (222Rn) (Bq/m3)
Arithmetic
mean

Geometric
mean

209

Notes

Maximum
value

Geometric
standard deviation

1.3

West Asia
Armenia [U2]

3.64

104

216

Islamic Republicof Iran

63.76

82

3 070

Countrywide average

Islamic Republicof Iran

795

2 745

31 000High-background areas

600

1 000

High-background areas

Islamic Republicof Iran
KuwaitPalestine (Gaza) [Y7]1.69 [U2]

14 [U2]

0.95

34

105

–

16

36

–

44

520

Saudi Arabia [A25]
Syrian Arab Republic
[U2]

14.57

10.6

119.2

0.74

Countrywide average

North Europe
Denmark

5.2

59a

39a

1 200

Estonia [U2]

1.47

120

92

1 390

Finland

5.2

120

84

20 000d

Iceland0.3

10Lithuania

3.73

49

Lithuania

3.49

Norway [U2]
Sweden2.2Countrywide average

2.1

Countrywide average

26

Countrywide averageg

38

1 900

Countrywide average

55

36.5

636

Countrywide average

4.35

73

40

50 000

–

8.88

108

56

84 000Countrywide average

15

190

–

West Europe
Austria [U2]8.11

Belgium

10.22

48

38

12 0002Countrywide average

France [U2]

58.33

62

41

4 690

2.7

–

89.3

53.5

4 964

Population-weighted mean of
63 Bq/m3 based on 12,261
measurements in dwellings

Germany [U2]

81.92

50

40

>10 000

1.9Countrywide average

Ireland

3.84

89

57

7 000

2.4

Countrywide average

Liechtenstein

0.03

80

Luxembourg

0.22

110

70

2 500

2.0

Countrywide average

Netherlands [U2]

15.58

23

18

380

1.6

–

Switzerland

6.71

75

41

10 000

142b
73c

81b
59c

15 000b
15 000c2.6b
1.8cCountrywide average20

14

17 000

3.2[C26]

England [G10]9050Wales [G10]8448

Switzerland
United Kingdom [U2]

58.14

e

1 098

Countrywide average

Average (20 Bq/m3) populationweighted average

210

UNSCEAR 2006 REPORT: VOLUME II

Region/country

Population(10 6)

Indoor radon (222Rn) (Bq/m3)
Arithmetic
mean

Geometricmean

Maximum
value

Notes
Geometric
standard deviation

East Europe
Belarus
Bulgaria

31.8

221

8.10

Czech Republic

118

Countrywide average

22

250

2.1

Countrywide average

94.4

70 000

1.84Countrywide average442

20 000High-background area

214

20 000High-background area

124

70 000High-background area

112

20 000High-background area

136

6 000

High-background area

214

6 500

High-background area

Hungary [U2]

10.05

107

82

1 990

2.7

Poland

38.12

49.1

Poland38.17

49

Romania

22.55

25.0

564

Countrywide averagee

Slovakia [U2]

5.35

87

3 750

–

Countrywide averagee

1 300
31

3 260

–

2.3

Countrywide averageg

South Europe
Albania [U2]

3.40

120

105

270

Croatia [U2]

4.50

35

32

92

Cyprus [U2]

0.76

7

7

78

Greece [U2]

10.49

73

52

490

55

44

1 700

2.4

Countrywide averageg

Greece

2.0

–
–

2.6

–
–

Italy [U2]

57.23

75

57

1 040

2.0

–

Italy

57.3

70

52

1 036

2.1

Countrywide averageg

Montenegro(Yugoslavia)

0.60

184

110

1 128

2.74

Countrywide averagee

Portugal [U2]

9.81

62

45

2 700

2.2

–

Slovenia [U2]

1.92

87

60

1 330

2.2

–

Spain

36.72

90.38

45.69

15 400Countrywide averagee

Spain

0.001

748.5

242.6415 400High-background arease

Spain

40.84

90.4

45.7

15 400

Spain

0.02

610.0

2.9

Countrywide averageg
High-background areasg

1 400.0
Oceania

Australia [U2]

18.06

11

8

420

New Zealand

3.81

21.5

19.5

80

a Population-weighted average.
b Upper value, unweighted data.
c	Lower value, data weighted for floor dependence, population distribution, errorof exposition and mobility.
d Annual mean.

2.1–
Countrywide averageee	Data from UNSCEAR Global Survey on Exposures to Natural Radiation Sources(2001-2006), submitted in 2001.
f Ibid., submitted in 2004.
g Ibid., submitted in 2006.

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

C.

Workplaces

workers involved in the nuclear fuel cycle, radon exposure
from mining and milling is a relatively important contributor
to the per caput dose.

54. The spectrum of workplaces other than mines whereradon can present a hazard is large. While it includes belowground workplaces such as subways, tunnels, undergroundparking, stores, caves, spas and closed-out mines open tovisitors, the majority of such workplaces, such as factories,
shops, schools and offices, will be above ground.

58. The total number of workers exposed to human-madesources and enhanced natural sources was given in the
UNSCEAR 2000 Report [U2] as 11.1 million. Approximately
4.6 million workers were exposed to human-made sources
at an annual average (effective) dose of about 0.6 mSv.
Some 6.5 million workers were exposed to enhanced natural
sources at an annual average dose of about 1.8 mSv, of whichapproximately half was from radon. The estimate for radonin above-ground workplaces (for example in the ­phosphateindustry) is still considered to be crude.

55. Other workplaces where large quantities of materials with elevated radium concentrations are stored or pro­
cessed, for example phosphate fertilizer production [R11] ormonazite sand mining, can similarly exhibit elevated radon
and thoron levels [A30, C42]. One study [S71] reporting onradon and thoron concentrations in the monazite production
area of a rare earth facility in Kerala, India, found radon­levels to be below the detection level of about 1.7 Bq/m3 (seereference [S72] for discussion of monitoring techniques),
but the c­ orresponding thoron levels were 5.9 kBq/m3.

59. In the case of occupational exposure, there are­several situations where workers who have the highest
radiation doses receive a significant contribution fromradon. These include the situations listed in table 2 (fromreference [U2]). For workers involved in nuclear power
production, those involved in the mining of uranium typically receive the highest collective doses; a significant partof that exposure is from radon inhalation. The group of
workers in the category of above-ground workplaces (seetable 2) are the second largest group identified in annex Bof the UNSCEAR 2000 Report [U2]. These workers were
estimated to receive an average annual effective dose of
4.8 mSv. This is the largest average annual dose received byany type of worker and was due entirely to radon. According to the UNSCEAR 2000 Report [U2], radon inhalationis also a significant contributor to the doses to the othercategories of workers in table 2. While information to fully
update table 2 is not available, a paper by Liu et al. [L46]indicates that there are about 6 million miners in Chinese
coal mines alone, which nearly doubles the number of coalminers reported in reference [U2].

56. Underground workplaces, including mines other thanuranium mines, especially coal mines (e.g. [B17, J8, S43,
U16, V5]), can accumulate high radon levels in the same
way as natural caves or abandoned mines. High radon levels
in underground workplaces will not be limited to only thoseareas where elevated levels were found in above-groundworkplaces. The experience of the Newfoundland fluorspar
miners illustrates this. The levels of 226Ra in the host rockwere low, but the miners were exposed to elevated levels ofradon and short-lived decay products arising from radon in
the groundwater, which entered the mine and subsequently
the mine atmosphere (e.g. [A14, M16, M17]).
57. Exposure to environmental sources of radon from­mining and mineral processing is common, and the inhalation of RDPs can be a significant exposure pathway. For
Table 2

211

Situations where doses from radon are significant [U2]
Source/practice

Number of monitored workers

Average annual effective dose (mSv)

Nuclear fuel cycle (including uranium mining)

800 000

1.8

Mining (other than coal and excluding uranium mining)

760 000

2.7

Coal mining a

3 910 000

0.7

Mineral processing

300 000

1.0

Above-ground workplaces (radon)

1 250 000

4.8

a See paragraph 59.

D.

Measurements of radon and radon decay products

60. It is well established that the inhalation of the shortlived decay products of radon (222Rn) and their subsequentdeposition along the walls of the various airways of the bronchial tree are the main pathways of radiation exposure of thelungs [U2]. As discussed elsewhere in this report, the lung

exposure arising from the decay products of thoron (220Rn) is
also of increasing interest. Traditionally, the potential alphadecay energy per litre of air (referred to in units of working
level month (WLM)) was the measure of exposure to RDPs
used in evaluations of exposures in mines. Miner epidemiological studies use data based on measurements of this type.
Later, time-integrated radon measurement techniques were212

UNSCEAR 2006 REPORT: VOLUME II

developed and are the method of choice for modern studies
of residential radon. The discussions of each miner cohort(section IV) or residential case–control (section V) epidemiological study include a brief description of the methods
used for measuring radon and its decay products; in addition,a few general comments are provided below. Further information is available in published reports (e.g. NCRP Report
No. 97 [N15]).
61. Various techniques were used in the past to assess residential radon exposure, including: instantaneous grab samples
and subsequent analysis in a scintillation cell; accumulation
on a charcoal absorber with subsequent gamma spectroscopy;
various solid state detectors; and a variety of other techniques,
including continuous measurements [I12, N15]. Overall, longterm track etch radon measurements are widespread and arealmost universally used for residential epidemiological ­studies.
In recent years, increasing attention was given to measurementsthat assist in the reconstruction of past exposures, including
the measurement of polonium (210Po) activity on glass surfaces
or in volume traps (e.g. [F1, F9, L31, M33, N11, N15, P15]).
Bochicchio et al. [B41] describe a comprehensive quality
assurance programme for radon measurements carried out aspart of a residential radon case–control study.
62. A paper on the use of (outdoor) radon levels in meteor­
ology provides additional information on environmentalradon monitoring and experimentally available lower ­limitsof detection [Z7]. Tokonami et al. [T42] conducted an intercomparison of measurement methods for radon, RDPs andparticle size using a radon/aerosol chamber. The authors
report good agreement (typically within +5%) across institutions and methods for continuous radon monitoring. However, the ratios of the radon concentrations measured by­various institutions to those measured by alpha track detectors ranged from about 0.71 to 1.34. The authors noted thatthese types of detector (alpha track detectors) are often usedin surveys and that quality control is always needed.
63. Much of the miner epidemiology is based on ­exposurescharacterized by relatively short-term (grab) measure­ments

Table 3of WL in one or more areas of a mine (combined with anevaluation of the hours worked in the same location).
According to reference [L13], the first radon gas measurements in United States mines were in 1949, and the first
RDP measurements were in 1951. In 1973, the American
National Standards Institute (ANSI) published a consensus
standard for radiation protection in uranium mines in the
United States [A22]. According to this standard, a monitoring system for RDPs must be capable of measuring theannual accumulated exposure in WLM within an uncertainty
interval of 50% at the 95% confidence level. To satisfy the
ANSI criterion, uranium mining companies in the UnitedStates adopted a procedure based on periodic measurementsof the air concentration of RDPs. All of the measurement
methods in routine use at the time involved drawing a known
volume of air through a filter. After a specified time delay,
the activity on the filter was analysed. The concentration ofshort-lived RDPs, in units of working levels (WL), was calculated from the measured activity on the filter. The mostcommon method was that of Kusnetz [K23], who deviseda procedure for estimating WL from a single alpha count.
(The Kusnetz method generally involves a 5-minute sampleof air drawn through a glass fibre filter paper at a rate of
1–2 L/min; the filter paper is counted for total alpha activity after a delay of 40–90 min and the concentration in WL
is determined from the total alpha count using a correction
factor based on the decay time [K23].) For control purposes,
a single measurement was used to establish whether or nota work area was safe for occupancy by miners. For routine
dosimetry, the concentration estimated from the grab sample
was assigned to the appropriate work area. As discussed laterin section IV, similar practices were followed in many uranium mines outside the United States.
64. Twelve grab sampling methods for measuring airborneradon decay product concentrations were analysed by Schiager
et al. [S56] to determine whether they would satisfy the ANSI
standard [A22]. The evaluations considered six independentsources of uncertainty that together, according to reference[S56], comprised the overall uncertainty of the method. The
results of their evaluation are summarized in table 3.

Uncertainties in grab sampling (adapted from reference [S56])
Nature or source of uncertainty

Total uncertainty(Percentage of measured value)

Variations in airborne RDP concentrations

36

Inherent errors of the method

4–20

Precision based on counting statistics (dependent on concentration)

1.4–9.3 a

Human error

1–4

Estimation of occupancy time

4

Record-keeping and data transcription

1.5

Overall uncertainty

37–41

a Range for 10 out of 12 methods analysed at 0.3 WL.

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

65. For comparing the uncertainties involved in different grab sampling methods, Schiager et al. [S56] assumed
that all of the sources of uncertainty were independent,
multiplicative and normally distributed. The total relative
uncertainty in an individual measurement was estimated,by conventional propagation of error techniques, as thesquare root of the sum of the squares of the individual
relative uncertainties. The result was an approximation,
since the shape of the distribution of errors was not determined. In spite of the limitations of the analysis, Schiager
et al. [S56] determined that the estimated uncertainties are
entirely adequate to indicate the nature and magnitude of
the various sources of potential errors and the adequacyof the sampling methods used for meeting the ANSI
criterion.
66. Makepeace and Stocker [M35] presented a statisticalanalysis of WL measurements in Canadian uranium mines.
No attempt was made to establish the accuracy or precisionof the individual measurements. A total of 2,427 observations were obtained at 33 mine locations. The number of213

observations per location varied from 22 to 188, and the sampling periods ranged from 4 to 12 days. The combined data at
each location had a coefficient of variation ranging from 5 to
95%, with an arithmetic mean value of 30%, consistent with
the ANSI standard. Francis et al. [F18] evaluated the effects
of autocorrelation on workers’ daily exposures and concluded
that sampling programmes that rely on measurement on consecutive days (a possibility in the uranium mine environment)
can result in biased estimates of exposure if autocorrelation
is present. They suggested a random sampling strategy be
employed where day-to-day correlation is high.
67. The possible role of exposure to thoron and its decay
products is of increasing interest, and a number of authors
report that the contribution of thoron and its decay products can be a significant component of the total exposure(radon plus thoron); thoron can thus be a source of error in­residential radon studies that do not distinguish the two contributions to exposure (e.g. [C21, T11, T17]). Future measurement studies should therefore consider the contributionof both radon and thoron.

II.

DOSIMETRY

68. The health risk associated with radon arises from theinhalation of the short-lived decay products and the consequent dose to critical cells of the respiratory tract. Estimatesof the absorbed dose to the critical cells of the respiratory tract
per unit radon exposure can be derived from an analysis of
information on aerosol size distribution, unattached fraction,breathing rate, fractional deposition in the airways, mucous
clearance rate and location of the target cells in the airways.
Such estimates are model-dependent and necessarily subject
to all the uncertainties associated with the input data as well aswith the assumptions built into the particular model. The dose
calculation procedure and assumptions are described in the
UNSCEAR 2000 Report [U2]. The magnitude of the risk fromexposure to RDPs was quantified in epidemiological studies of
the increased rate of lung cancer among uranium mine workers
and more recently in residential case–control studies.

radioactive material. Such models incorporate data on respiratory tract deposition, including data on deposition ofultrafine particles in nasal passages (e.g. [H22, I3, I13, J1,
J2, N8, N10]). These models, widely used for the assessment of dose for most inhaled radionuclides, show that thedose per unit intake of RDPs is dependent upon the site ofdeposition within the respiratory tract. The site of deposition in turn is strongly dependent on the particle size ofthe airborne RDP, particularly for those particles of below
10 nm diameter (typically the “unattached fraction” of
the RDP, i.e. the fraction not attached to ambient particulates). A weighted dose conversion factor for a particular
exposure location can be derived from measurements of
RDP size distribution combined with the particle-size­dependent dose conversion factors, calculated using oneof these dosimetric models.

69. ICRP Publication 65 [I2] recommended the use of
the risk factors determined from epidemiological studies
of uranium miners as the preferred method for converting RDP exposure to effective dose. Since that time, as
discussed in section VI, numerous case–control studies ofresidential exposure to radon suggest risk factors that are
generally consistent with those from miner epidemiology.
These case–­control studies provide direct evidence of risk
from residential radon, removing the dependence on dosi­
metric adjustments from conditions in mines to conditions
in homes. Thus, as discussed in section V, the assessment ofrisks from exposure to radon in the home can now be basedon the evidence from residential case–control studies.

72. Various techniques are used to measure size distributions of RDPs associated with fine (5 nm) and ultrafine(0.5 nm) particles. These are extensively reviewed ­elsewhere(e.g. [N8]). Many of these methods rely on ­fractionationof the sub-micrometre radioactive particles by diffusion
processes, using sets of wire screens, or by inertial andimpaction processes, using cascade impactors. The lattersystems are also capable of resolving particle sizes exceeding 1 µm. These systems have been used to measure RDP
size distributions in homes, workplaces and uranium mines[G1, G2, R7, S44, S57, T23, T25]. A simpler techniqueinvolving the measurement of the unattached fraction is
widely used [H23]; this method relies on the separation of
the ultrafine and accumulation modes (particles in the size
range 50–500 nm, which grow on inhalation, affecting thepattern of deposition in the lungs) using a diffusion sampler, usually a single wire screen. When used with the optimized configuration, a system with a set of wire screenscan measure size distributions in both modes concurrently[F16, F17]. Tokonami et al. [T49] describe two methods ofmeasuring the particle size distribution of RDPs and report
data measured in a mine in Japan, as well as in a wellcontrolled radon chamber.

70. The ICRP has provided guidance on a dose conversion
convention for radiation protection purposes [I2]. For radon,
the ICRP recommends the RDP conversion convention that
all exposures are combined on a dose- and risk-equivalentbasis. The use of a single value for the dose conversion factor
implicitly assumes: (a) that the distribution of RDP particle
size for occupational exposure, in particular the fraction andparticle size of the ultrafine mode, is not too different fromthe particle size distribution in uranium mines; and (b) for
exposures of the general public, that the differences in aerosol
conditions are offset by lower breathing rates for members of
the public, particularly children. The current ICRP conversion
convention is not applicable to thoron decay products.
A.

B.

Dosimetry of radon and thoron

73. A number of researchers have used the ICRP dosi­
metry model to assess the doses from radon and thoron andtheir decay products (e.g. [B35, I7]). The ICRP advises
against using the human respiratory tract model for risk
estimation [I3].

Dosimetric models

71. Dosimetric models can be used to estimate theradiation doses arising from the inhalation of airborne
215

216

UNSCEAR 2006 REPORT: VOLUME II

74. Tokonami et al. [T36] report measurements of physical parameters related to dose assessment in an actual room
where environmental conditions such as ventilation rate,
air-conditioning and the operation of an air cleaner werevaried. Using the data and the ICRP Publication 66 respiratory tract model, dose conversion factors were calculated
under different situations and compared with each other. The
paper investigated the wide variation of dose conversion factors with environmental conditions and concluded that themost sensitive parameter is the unattached fraction of decay­products (fp).
75. Other dosimetric models have also been described [H3,N10], and current models are relatively consistent in theirdose estimates, particularly in showing a strong dependenceof the radiation dose per unit intake on the size of the inhaledradioactive particles (aerosol in the range 0.8 nm to >1 µm).

Table 4Year

The publication of James [J5] contains an illustration of thedose conversion factors according to various models.
76. The UNSCEAR 2000 Report [U2] listed, in table 26 ofannex B, the principal dosimetric assessments for the lungdose from deposited RDPs. For convenience, this table isreproduced here as table 4. Graphs of the dose coefficient
as a function of median inhaled aerosol diameter, breathing
rate and unattached fraction are reproduced from reference[U2] in figure II. The current RDP dosimetric models use thestandard values for tissue and radiation weighting factors toconvert tissue dose to effective dose. The effective dose estimates vary, but are within a factor of 3 higher than the estimates derived from the epidemiological approach used by
the ICRP conversion convention. Considering the uncertainties in both the epidemiological and dosimetric approaches,
this agreement is remarkable.

Principal dosimetric assessments of lung dose from deposited RDPs [U2]
Investigator

Parameter values
Unattachedfraction

Breathing rate
(m3/h)

Target region

Model type

Dose factora[nGy(Bq h m‑3)‑1]

1956

Chamberlain andDyson [C1]

0.09

1.2

Average in 45 µm
epithelium

Cast of trachea and bronchi

11

1959

ICRP [I4]

0.1

1.2

Mean tracheobronchialregionDeposition retention assumptions

6.7

1964

Jacobi [J1]0.25

Basal cells (30 µm)

Findeisen–Landahl 6-region anatomical
model

24

1964

Altshuler et al.
[A3]

0.085

0.9 basal

Cells (22 µm)

Findeisen–Landahl 6-region anatomical
model

32

1967

Haque and
Collinson [H3]

0.35

Basal

Cells (30 µm)

Weibel dichotomous airway model

71

1972

Harley andPasternack [H5]0.04

0.9 basal

Cells (22 µm)

Weibel dichotomous airway model

5.7

1980

Jacobi and Eisfeld
[J2]

0.1

1.2

Mean epithelium

Weibel dichotomous airway model,
correction for upper airway turbulent
diffusion [M2]

8.9

1980

James et al. [J6]

0.1

1.2

Mean epithelium

Yeh–Schum anatomical model [Y2]

14

1982

Harley andPasternack [H6]0.07

1.1 basal

Cells (22 µm)

Same as Jacobi and Eisfeld [J2]

6.4

1982

Hofmann [H10]

0.2

0.9

Mean epithelium

Same as Jacobi and Eisfeld [J2]

11

1991

National ResearchCouncil [N10]0.16

1.2 basal

Cells (35–50 µm)

Yeh–Schum anatomical model [Y2],correction for upper airway turbulent
diffusion

21

1996

Harley et al. [H4]0.1

1.2 basal

Cells (27 µm)

Nikiforov and Schlesinger [N12] anatomical model, airway deposition fromempirical data from human airway casts

9

1998

Marsh and Birchall
[M2]

0.08

0.8

Bronchial cells: basal(35–50 µm), secretory(10–40 µm) Bronchiolar
cells: secretory(4–12 µm)

ICRP lung model [I1]a	Per unit 222Rn concentration (EEC). WLM converted to Bq h m–3 using 0.27 × 10–3 WL (Bq/m3)–1 and 170 h per working month.

8.5
19
14

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

217

Figure II. Absorbed dose in bronchial epithelial cells per unit exposure to radon decay products as a function of aerosol
size [U2].
70
60

40
30
20

20
10

0
200

600

400

800

0.21
0.11
0.05
0.02

30

10
0

For breathing rate 0.6 m3/h
Unattached fraction fp

40

1.2 m3/h (oral)
1.2 m3/h (nasal)
0.6 m3/h (nasal)
0.3 m3/h (nasal)

50

DOSE COEFFICIENT[nGy (Bq h m-3)-1]

DOSE COEFFICIENT[nGy (Bq h m-3)-1]

50

For unattached fractionof 0.05

0

1000

0

MEDIAN AEROSOL DIAMETER (nm)

200

400

600

800

1000

MEDIAN AEROSOL DIAMETER (nm)Note: Extract Figure XIV from UNSCEAR 200, Annex B

77. It is not possible to assess the radiation dose frominhalation of thoron decay products by epidemiologicalmeans, and therefore it must be estimated using dosi­metricmodelling. In annex A of the UNSCEAR 2000 Report[U2], a conversion factor for thoron decay products of
40 nSv (Bq h m–3) was used. According to reference [U2],this value was intended to include the dose to organs otherthan the50lungs due to the transfer of 212Pb from the lungs.
For breathing
rate 0.6 m3/h
Table 5 provides a summaryof the principal
dosimetricassessments of lung dose from deposited thoron decay
products and supports the continued use of a conversion
factor of 40 nSv (Bq h m–3)–1. Marsh and Birchall [M42],
in a comment on the review of thoron dosimetry issues[N16], report thoron dose conversion factors based on thelatest ICRP biokinetic models and provide a range of doseconversion factors from 1.1 mSv/WLM to 3.8 mSv/WLM,
which encompasses the full range of previously estimatedthoron dose conversion factors.

Unattached fraction fp

40

DOSE COEFFICIENT[nGy (Bq h m-3)-1]

0.21
0.11
Table 5 Principal dosimetric assessmentsof lung dose from deposited thoron decay products
0.05
30
0.02
Year

Investigator

20

1956

Parameter values

Target region

Unattachedfraction 212Pb

Breathing rate
(m3/h)

0.02

0.3

10 Chamberlain andModel type

Effective dosea[nSv(Bq h m‑3)‑1]

Average in 45 μm epithelium

Cast of trachea and bronchi

30

Mean tracheobronchial region

Deposition retention assumptions

43

Basal cell generations 2–15

Weibel43

Basal cell generations 2–15

Weibel

35

Bronchial fromwhole lung × 0.06

Weibel64

Dyson [C1]1959 0 ICRP [I4]
1973
1980

0

200

400

0.02

600

1.2

800

Harley and
0.02
0.9
MEDIAN AEROSOL DIAMETER (nm)Pasternack [H31]
1.2

1000

Jacobi and Eisfeld
[J2]

n.a.

1981

ICRP 32 [I10]

n.a.

1.2

Bronchial fromwhole lung × 0.06

Based on Jacobi and Eisfeld [J2]

73

1982
(1993)

James et al. [J16](James [J5])

n.a.

1.2

Bronchial basal cell

Birchall–James model using
either Weibel or Yeh–Schum [Y2]model

34–103

1983

NEA 1983 [N8]

0.02

1.2

Mean bronchial

Jacobi and Eisfeld [J2]

36

2000

UNSCEAR [U2]

n.a.

1.2

Whole body

ICRP 50 [I5]

40

2001

Porstendorfer[P12]

0.005–0.02

0.75

Bronchial

Modified from ICRP 66 [I3]weighting by basal and secretorycell density

53

Note: Extract Figure XIV from UNSCEAR 200, Annex Ba	Per unit 220Rn concentration (EEC). WLM converted to Bq h m–3 using 3.6 × 10–3 WL (Bq/m3)–1 and 170 h per working month.

218

UNSCEAR 2006 REPORT: VOLUME II

C.

Dose conversion factors

78. The health risks from exposure to radon and thoron
are principally due to the inhalation of the short-lived decay
products and alpha particle irradiation of the bronchial airways. Radon and thoron decay product exposure rates are
specified by the measure of potential alpha energy concentration (PAEC), given in units of J/m3 or working levels(WL), and the equilibrium-equivalent concentration (EEC),
given in Bq/m3. The potential alpha energy concentration
is derived from a linear combination of the activities of theshort-lived decay products in each radon decay series (seeparagraph 122, annex B, UNSCEAR 2000 Report [U2]).
The constants in the linear combination are the fractional
contributions of each decay product to the total potentialalpha energy. The EEC (in Bq/m3) can be converted to the
PAEC by the relationships:

included in ICRP’s calculations, though cited in them. Their
inclusion yielded a mean ERR coefficient of 3.08 (J h m–3)–1(1.09% WLM–1), about 80% of the 3.79 (J h m–3)–1(1.34% WLM–1) value estimated by the ICRP.
82. Alternative methods to ICRP’s weighting by personyears can also be used to arrive at a mean risk estimate.
One alternative considered by Lowe and Chambers [L7]was based on the confidence interval (CI) associated witheach estimate: in this approach, the estimates with the leastassigned uncertainty are given the most weight in estimating an overall mean. The method of combining values with­different CIs was to weight each value by the inverse of its
variance. It was assumed that the 95% CI about each estimate approximated to a range of ±2 standard deviations
about the estimate, and that the variance was the square of
the standard deviation. This weighting method gave a mean
ERR of 0.73 (J h m–3)–1, or 0.26% WLM–1.

1 Bq/m3 = 5.56 × 10–9 J m–3 = 2.7 × 10–4 WL (222Rn);and
1 Bq/m3 = 7.6 × 10–8 J m–3 = 3.64 × 10–3 WL (220Rn).
79. For occupational exposure to inhaled 222Rn decay
products, the ICRP 65 [I2] recommended the use of a ­single
factor (conversion convention) to relate the 222Rn decay
product exposure to the effective dose to an individual.
This conversion convention is based on a comparison of therisk to a uranium miner, based on epidemiological studies,with the risk to a radiation worker from an effective dose of
1 Sv, in other words, comparison of the radiation detriment
coefficient (risk per unit dose) with the miner detriment(risk per PAEC exposure). For worker exposure, this ­factor
is 1,430 mSv (J h m–3)–1 (rounded to 1,400), 5.06 mSv/
WLM (rounded to 5 mSv/WLM), or 7.95 nSv (Bq h m–3)–1(rounded to 8 nSv (Bq h m–3)–1) EEC (tables 7 and 8, ICRP
Publication 65 [I2]).
80. In recommending that a similar approach be used formembers of the public, ICRP Publication 65 [I2] assumed
that the lung cancer risk per unit exposure in a home was thesame as that in an underground uranium mine, in order toderive a conversion factor for members of the public. Since
the detriment coefficient for the public is greater than forworkers (7.3% Sv–1 against 5.6% Sv–1), the derived conversion convention for members of the public was calculated
to be 1,100 mSv (J h m–3)–1, or 3.88 mSv/WLM (rounded to
4 mSv/WLM), or 6.1 nSv (Bq h m–3)–1 EEC.
81. In developing ICRP Publication 65 [I2], data from epidemiological studies of seven miner cohorts were used toderive a central estimate of the excess relative risk (ERR)of exposure to RDPs. The ICRP (table A.2 of reference[I2]) estimated a mean ERR coefficient of 3.79 (J h m–3)–1(1.34% WLM–1), by averaging the results from the sevenselected cohorts, weighted according to person-years of riskfor each cohort. Lowe and Chambers [L7] considered three
additional epidemiological studies (the Port Radium uranium miners, the Chinese tin miners and the Newfoundland
fluorspar miners, as discussed in section IV) that were not

83. Stather [S62], in a discussion of the dosimetric andepidemiological approaches, arrives at a similar conclusion and suggests that the difference of about a factor of 3 is“surprisingly good”. The Committee agrees with this viewand simply notes that the calculated doses are in reason­
able agreement with risk factors derived from epidemiology,
uncertainties in both approaches considered.
D. Uncertainties in dose conversion factors
84. Uncertainties are present in both epidemiological(see sections IV and V) and dosimetric approaches, and thedosimetry of inhaled RDPs is quite complex and depends on
many factors (biological, physical and behavioural). Many
authors have reported estimates of doses, including references [A3, C1, H4, H5, H6, H10, J1, J2, J5, J6, N1, Y2, Y3,
Y4, Y5], and both sensitivity (e.g. [M2]) and uncertainty
analyses (e.g. [M32]) have been conducted. It may be thata single dose conversion factor cannot adequately cover thevariety of natural and occupational exposure situations.
85. Marsh et al. [M32] described a detailed parameter
uncertainty analysis for the weighted equivalent lung dose(absorbed dose averaged over the lung that is weighted forthe relative biological effectiveness of alpha radiation) per
unit exposure to RDPs in the home. The authors commented
that the ICRP (para. 356 of reference [I3]) recommended
that the risks from residential radon be based on epidemiological studies of miners, from which a conversion factor(effective dose per unit exposure to radon) of 5 mSv/WLM
for workers was estimated. In contrast, Birchall and James[B18] calculated a conversion factor of 13.4 mSv/WLM forminers based on the ICRP’s human respiratory tract model[I3], a factor of 2–3 times larger than that estimated fromminer epidemiological studies. In carrying out the parameter
uncertainty analysis, Marsh et al. [M32] assumed a dosi­
metric tissue weighting factor of 0.12 for the lung and analpha radiation weighting factor of 20. Their analysis considered various aerosol parameters, target cell parameters,

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

and parameters such as breathing rate and fraction of breathing through the nose, related to the characteristics of individuals at home. Using the ICRP’s weighting factors for
exposure at home, the mean ratio of the distribution of milli­sieverts per working level month was found to be about 15.
It was ­further concluded that a conversion factor of as low
as 4 mSv/WLM was extremely unlikely from a dosimetricperspective [M32].
86. In a separate analysis of the physical parameters and(dose) conversion factors for RDPs, Porstendorfer [P12]­calculated a range of (dose) conversion factors of 4.2–
11.5 mSv/WLM, depending on aerosol concentration andother factors. Homes with higher aerosol concentrations had
lower dose conversion factors. Porstendorfer [P12] noted
that, while dose conversion factors calculated for homes with“high” aerosol levels were comparable to those derived fromminer epidemiological studies, the discrepancy between thedose conversion factors derived from miner ­epidemiology anddosimetry remained for homes with ­“normal” aerosol concentrations. Dose conversion factors for exposure to thoron
decay products of 2–3 mSv/WLM were also calculated.
87. Nikezic et al. [N13] discussed the importance of theabsorbed fraction (of alpha particles in sensitive cells) whenestimating dose conversion factors for RDPs. The authors
noted that the most important parameter in estimating theabsorbed fraction was the depth of the sensitive cell layers.
In another paper [N14], the ICRP’s human respiratory tract
model [I3] and microdosimetric considerations were used toinvestigate the dose conversion factors for RDPs. The authors
concluded that having the alpha particles deposit their energy
only in the nuclei of sensitive cells reduced the dose conversion
factor from 15 mSv/WLM to about 11 mSv/WLM.
88. All of the above calculations assumed the standard
alpha radiation weighting factor of 20. Brenner et al. [B19]suggested that a quality factor of 20 may be too large, on
the basis that RDP alpha particles deposit most of theirenergy in a region of relatively low biological effectiveness,and recommended a value of about 10 for residential radon
exposure. The suggestion of Brenner et al. [B19] seems quite
reason­able on the basis of their review of the extensive data
available at present on the in vitro transformation of cultured
mouse cells. However, Brenner’s suggestion would also be
likely to raise questions at the tissue level, where the cell loss
by killing may stimulate cell renewal that eventually promotes the development of cancer (i.e. increasing the weighting factor). In ICRP Publication 92 [I11], the ICRP notedthat the current radiation weighting factor (wR) of 20 for
internally deposited alpha emitters can serve as a guideline,
and suggested that for specific situations such as exposure toradon and its progeny “more meaningful weighting factors
can be derived”, whether based on specific assumptions about
target cells and dosimetric models or on epidemiology.
89. Little [L21] compared lung cancer risk in the survivorsof the atomic bombings in Japan (using data from reference[P13]) and the Colorado Plateau uranium miners (using data219

from references [R8] and [H9]). Models of ERR were used,
and time since exposure, smoking and sex were considered.
Little found that, although there are statistically significant
differences between the two data sets in how ERR variedwith time since exposure, these differences were no longer
statistically significant when only male atomic bombing
survivors were used as the basis for comparison with the
Colorado Plateau miners. Little concluded that the conversion factor based on the atomic bombing survivors was 18(95% CI: 6.1, 110) mSv/WLM, using a model with exponential adjustments for the effects of radiation with time since
exposure and age at exposure, and 19 (95% CI: 6.2, 160)
mSv/WLM, using a model with adjustments for the effectsof radiation proportional to a power function of time since
exposure and attained age. The absence of smoking data for
the Japanese atomic bombing survivors was acknowledged
to be a potentially important confounding factor. Little’s
estimates compare with the range developed by Birchall andJames [B18] of 17.2–22.5 mSv/WLM (with 95% CI extending from at least two times smaller to at least two times
larger) using the ICRP lung model [I3] and with the ICRP’s
estimate of 5 mSv/WLM [I5] based on the Japanese atomic
bombing survivor data and the uranium miner data. Overall,
Little concluded that the various estimates of the dose conversion factor are “very close” to that predicted by dosimetry and “statistically compatible” with the epidemiological­derivation of the conversion factors [L21].
E. Exposures in homes and in workplaces
other than uranium mines
90. Both the BEIR IV and the BEIR VI Committees [C19,
C20] used dosimetric models to examine the comparative
doses for exposures of miners and people exposed at home.
As a follow-up to the BEIR IV report, the United States
National Academy of Science compared the dosimetry ofradon in mines with that in the home [N10]. This study
expressed differences in exposure–dose relationships interms of a “K-factor”, defined in reference [C19] as theratio of dose per unit exposure at home to the dose per unit
exposure to a male miner. Thus, for K < 1, the dose per unit
exposure at home is less than in a mine, and for K > 1, thedose per unit exposure at home is larger than for exposurein a mine. As reported in reference [N10], the K-factors forchildren and infants, calculated for a wide range of exposure scenarios, were somewhat larger than those calculated
for adults, but nevertheless did not exceed 1. BEIR VI [C20]
used the concept of the K-factor and included various environmental and physiological factors. BEIR VI reviewed
measurements of RDP activity size distributions for datacollected in Germany, the United Kingdom and the UnitedStates (annex B of reference [C20]). These size distributions were used with the RDP dosimetric model to estimate
the K-factor for exposure in homes. The values obtained
were close to 1 for male and female adults and for children. James et al. [J15] described the technical basis for thecomparative radon dosimetry applied to BEIR VI [C20],
including a thorough review of the basis for the BEIR VI

220

UNSCEAR 2006 REPORT: VOLUME II[C20] Committee’s choice of unity for the K-factor. Jameset al. concluded that a K-factor of unity is appropriate, but
noted that it is not yet established that the conditions in
the houses considered in the BEIR VI analysis are representative of conditions in houses in other regions acrossthe United States (including the effects of seasonal andclimatic conditions) [C20].
91. No similar review of RDP activity size distributions
has been undertaken for occupational exposure. While it
might be expected that RDP exposure in uranium mines
would be consistent with the ICRP conversion convention,
namely 8 nSv (Bq h m‑3)–1, there are a significant numberof workers exposed to elevated levels of environmentalradon in workplaces other than uranium mines (buildings,
other kinds of mine, caves, etc.). A series of RDP aerosol
measurements carried out in Australia in a range of workplaces [S45] showed that some workplaces have aerosol
conditions that differ markedly from those found in operating uranium mines, particularly those operating during the1950s and 1960s. Adjustment factors for these workplaces,
also referred to as K-factors, were derived from the Australian measurements for occupational exposures and are in therange 1–2.6. Use of the ICRP conversion convention in these

exposure situations would lead to a significant underestimation of radiation doses. Thus it is important to understand thefactors that affect the estimation of dose in workplaces otherthan uranium mines.
92. As discussed in the UNSCEAR 2000 Report [U2], theCommittee adopted the dose conversion factors set out in
ICRP Publication 65 [I2], which had been based on epidemiological evidence. Dosimetric evaluation of the absorbed
dose to the basal cells of the bronchial epithelium per unit
exposure gave values in the range 5–25 nGy (Bq h m–3)–1,
and a value of 9 nGy (Bq h m–3)–1 was estimated for averageindoor conditions. Using a tissue weighting factor of 0.08
for the bronchial and bronchiolar regions and a radiation
weighting factor of 20 for alpha particles, the effective dose
per unit equilibrium-equivalent concentration (EEC) became
15 nSv (Bq h m–3)–1. The epidemiological approach provides
a value of 6 nSv (Bq h m–3)–1, a factor of 2½ times lower,
and the dosimetric evaluations provide dose coefficientsin the range 6–15 nSv (Bq h m–3)–1. The UNSCEAR 2000
Report concluded that the value used by the Committee inearlier evaluations [U5, U6], 9 nSv (Bq h m–3)–1, is within
this range, and recommended that this value continue to be
used in dose evaluations [U2].

III.

EXPERIMENTAL STUDIES

93. Information from animal experiments and from cellularand molecular biology is helpful in understanding the underlying mechanisms of cancer induction, potential interactions
among agents and the uncertainties associated with extrapolating risks from exposure in mines to residential exposure.
The UNSCEAR 2000 Report [U2] discussed the biologicaleffects of low doses of ionizing radiation (annex G) and the
combined effects of exposure to radiation and other agents(annex H), and provided much information from animal
experiments, studies of DNA, and cellular and molecularresponses to various forms of radiation. This section provides a concise summary of experimental results, including
more recent data than reported in reference [U2].
A.

and major bronchi, and in other portions of rat lungs after
exposure to radon and/or RDPs. The respiratory tracts ofanimals that inhaled radon plus its decay products contained
125 times more radioactive material than those of animals
that inhaled radon alone [C44].
96. Beginning in the mid-1950s, at the University ofRochester in the United States, Morken and Scott initiated apioneering series of experiments (e.g. [M13]) to evaluate thebiological effects of inhaled radon and radon decay ­products
in mice; later experiments used rats and beagle dogs. The
essentially negative biological results of these studies suggested that alpha irradiation alone was relatively inefficientin producing tumours in the respiratory system. NCRP
Report No. 78 [N7] provides a comprehensive summary ofmany of the early animal data.

Animal experiments

94. Animal studies have been conducted for several­decades to identify the nature and levels of the uranium mine
air contaminants responsible for producing the lung cancers
observed among uranium miners (e.g. [C40, M29]). Many of
the initial studies were concerned with early effects or shortterm pathological changes (e.g. [R4]). Exposures were based
primarily on radon gas concentrations, and provided little
or no information on the radon decay product concentrations that contribute the greatest radiation dose to the lung.
The early studies (e.g. [K10]), in which lung tumours were­produced, were methodologically or statistically inadequate
to show an unequivocal association of lung tumours with
exposure to radon and/or RDPs.

97. In the late 1960s and early 1970s, other studies in France(Compagnie Générale des Matières Nucléaires (COGEMA))
and the United States (Pacific Northwest Laboratory (PNL,
later PNNL)) were initiated and later proved successful inproducing lung tumours from RDPs. The French investigators exposed rats to RDPs, either alone or in combination
with stable cerium, uranium ore dust or cigarette smoke, toproduce tumours in the lung (e.g. [C4, P4]). Later, the potential co-carcinogenic effects of various environmental andindustrial airborne pollutants, e.g. minerals from metallic
ore mines and diesel exhausts, combined with radon and/orradon progeny exposure were also investigated [M31]. The
later United States studies were designed to systematicallydetermine the pathogenic role of RDPs, either alone or invarious combinations with uranium ore dust, diesel engine
exhaust and cigarette smoke. These studies involved lifespanexposures of beagle dogs and Syrian golden hamsters, andchronic exposures of rats (e.g. [C13, C14]). A joint review
of PNNL (United States) and the Commissariat à l’énergie
atomique (CEA)–COGEMA (France) animal experimental
data was published in 1999 [C40]. Bronchial dose models
were published for the Syrian golden hamster, rats (Long–
Evans, Wistar, Sprague–Dawley, Fischer) and beagle dogs
[D23, H39, H40].

95. In the 1950s, there was growing concern that theincreased incidence of respiratory cancer observed in European uranium miners would also be found in United Statesminers. This led to the initiation of systematic studies
in the United States to identify the agents responsible forincreased incidence of lung cancer in miners and to develop
exposure–response relationships in animals. Investigators
at the University of Rochester began to focus attention on
the biological and physical behaviour of RDPs as well ason their contribution to the radiation dose to the respiratory
tract [B1, M12, M13]. Shapiro [S33] exposed rats and dogs
to several levels of radon alone and in the presence of RDPs
attached to room dust aerosols. He also showed that thedegree of attachment of RDPs to carrier dust particles wasa primary factor influencing the alpha radiation dose to the
airway epithelium. This dose was further demonstrated to be
due primarily (>95%) to the short-lived RDPs rather than tothe parent radon. Cohn et al. [C8] reported the relative levels of radioactive material found in nasal passages, trachea

98. A review of the animal studies through 1970 appearedin the final report of subgroup IB, Interagency Uranium
Mining Radiation Review Group [R4]. That report, which
addressed the early acute radon toxicity studies, concluded(as had an earlier Federal Radiation Council report [F2]) thatexperimental work prior to the 1970s had not demonstrated
that pulmonary carcinomas could be produced in animals ina systematic way from controlled exposures to radon and its
221

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UNSCEAR 2006 REPORT: VOLUME II

decay products. Since that review, discussions of the biological effects of inhaled radon and RDPs in animals haveappeared in ICRP Publication 31 [I1] and NCRP Report
No. 78 [N7]. A more detailed review of animal studies wasprovided in the most recent NCRP SC65 Report [N11].
99. Gilbert et al. [G3] analysed data on the risk of lungtumours in rats exposed to radon. Male, specific-pathogenfree rats were exposed starting at about 90 days of age to
RDPs (and uranium ore dust) at levels of from 20 WLM
to more than 105 WLM, and at exposure rates of 10 WL to103 WL. (The experiments are reported by the authors tohave resulted in an estimated dose of about 5 mGy/WLM atthe cellular level.) The authors used a time-dependent hazard
model to accumulate risk and examined several exposure–
response functions, the simplest being a simple linear model.
They concluded that the rat data were in “reasonable agreement with a linear exposure–response model” and estimatedan overall linear (risk) coefficient of 237 per 106 WLM. In
addition, their analysis showed “no observed evidence of
exposure-rate effect below 1,000 WLM”.
100. Bijwaard et al. [B20] used equations based on a twomutation biological model for two large data sets of radonexposed rats (10,000 in total). The improvement in fittingthese data sets separately compared with fitting them jointly
was statistically insignificant, indicating that a joint fit­represented the data well. The joint solution exhibited a firstmutation rate two orders of magnitude larger than the second
mutation rate, which was strongly suppressed by a mutation
killing term acting at high exposure rates. Maximum cancer
incidence occurred for exposure rates of 1–10 WLM/d, whichis in good agreement with reference [M10]. An inverse exposure-rate effect was evident at higher exposure rates, withincidence orders of magnitude larger than for lower exposure
rates. A very pronounced effect of age at exposure was also
observed. In all cases, ERR ranged between 0.007WLM–1
and 0.025 WLM–1. The proposed model compared reasonably well with a similar model derived for Colorado uranium
miners [L23], when the data for rats were scaled by the ratioof human to rat lifetimes.
101. Heidenreich et al. [H8, H26] used a two-step clonalmodel, based on Luebeck et al. [L9] and Moolgavkar et al.
[M11], to investigate the risk of lung cancer induction in rats
exposed to radon. The authors concluded that only fatal lungtumours among the rats could be used for generalizations
to models for lung cancer induction in humans. This model
for fatal tumours showed an inverse dose-rate effect at average exposure rates of above 20 WL, but below 10 WL thelung cancer risk per unit exposure decreased with increasing
duration of exposure. Finally, on the basis of their analysis,
the estimated ERR for rats at low exposure rates was in therange from 0.003 WLM–1 to 0.012 WLM–1, depending on the
exposure periods; this is of the same order of magnitude as
the ERR seen in humans. Because a statistical test had given
a strong indication that the results from different rat strains
should not be pooled together, Kaiser et al. [K25] carried out
separate risk analyses for two rat cohorts: the PNNL cohortof Wistar rats, and the combined cohort of Sprague–Dawley
rats at the CEA and at AEA Technology Plc (AEAT), United
Kingdom. The study was restricted to fatal tumours. Usinga refined technique of age adjustment [H34], the lifetime
absolute risk was standardized with the survival function forcompeting risks in the control population. The age-adjusted
excess risks for both strains of rats were of similar magnitude, despite the higher lifetime excess absolute risks perunit exposure (LEAR at 1 WLM) in the European cohort
because of the very low mortality in the control group.
102. Rats were exposed to tobacco smoke and radon(1,000 WLM) at the CEA. When animals were exposed tocigarette smoke prior to radon exposure, a slight decrease inlung carcinoma incidence was observed compared with rats
exposed to radon only, but when the cigarette smoke exposure occurred after the radon exposure, there was a highly
significant factor of 4 increase in lung carcinoma incidence.
This resulted mainly from an increase in squamous-typetumours. These data were used to estimate smoke-dependent
parameters in the biologically based two-stage clonal expansion model [H36]. Although smoke had no effect on tumour
initiation, an effect was seen on tumour promotion. Promotion by cigarette smoke was also seen in levels of adenomatosis in the PNNL studies. It appears that preneoplastic
lesions induced by radon are promoted by cigarette smoke.
103. Monchaux and Morlier [M28] reported the resultsfrom a study of the effect of exposure rate on lung cancer
induction in radon-exposed Sprague–Dawley male rats. The
study was conducted at relatively low cumulative exposuresof about 100 WLM, which is comparable to lifetime exposures in high-radon houses or to current underground mining
exposures. The risk of lung cancer in rats decreased withpotential alpha energy concentration (PAEC), i.e. exposure
rate, confirming the results obtained at lower exposures[M30]. These results and those from former experiments[M31] indicated that the risk of lung tumour induction inrats was maximal for cumulative exposures ranging from 25
to 200 WLM and PAEC ranging from 50 to 150 WL, i.e.
exposure rates ranging from 5 to 25 WLM per week. These
data suggest that there is a “watershed” at cumulative exposures of about 50 WLM. Below this exposure, decreasing theexposure concentration (WL) or protracting the time overwhich the dose is delivered results in a reduction in the lungtumour risk. Above this level, the reverse is true: decreasing exposure concentrations or protracting the exposure time
results in an elevated lung cancer risk [M29]. These results
were confirmed by Collier et al. [C38] in a series of lifespanexperiments in which the effects of radon and its decay products were investigated at different total doses, dose rates andunattached fraction. Collier et al. [C16, C34] also reported
on studies of the factors that affect the risk of inducing lungtumours in rats exposed to radon and RDPs.
104. The results of rat experiments conducted in parallelat the CEA and AEAT (as described in Work Package 4 ofreference [T30]) have been reviewed by various individuals.
The studies were carried out specifically to investigate theANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

effect of exposure rates on induction of lung cancer at cumulative exposures of about (100 WLM) (0.36 J h m–3). Except
for the fact that rat exposures in the CEA experiments werecarried out for a working day without a carrier aerosol andthe AEA Technology experiments were for continuous exposure with carnauba wax as a carrier aerosol, the experimental designs were similar. The joint analysis comprised morethan 4,000 exposed rats and 1,500 non-exposed control rats.
The authors calculated both relative and absolute risks andfound, in general terms, increased risks of lung cancer bothwith increasing cumulative exposure and with increasing
exposure rate. These results indicated that, at low exposurescomparable to those in modern mines or high-radon homes
and for cumulative exposures of up to about 100 WLM, therisk of lung cancer in rats decreased with decreasing exposure
rates, confirming earlier results [L43, M30] at lower cumulative exposures. A parallel analysis of these experimental datawith data from European uranium miner cohort studies was
performed [L44, T30]. It confirmed that epidemiologicaland animal data are consistent in showing an increase of risk
with cumulative exposure protracted at low exposure rates.
105. Mitchel et al. [M9] reported an experimental study
using a nose-only breathing system with male Sprague–
Dawley rats exposed to one of two concentrations of naturaluranium ore dust (44% U, at 50 mg/m3 or 19 mg/m3) without significant radon exposure. They concluded that chronic
inhalation of ore dust alone posed a risk of lung cancer inrats that was directly proportional to dose rate.
106. Although it was established that the rat constitutes a­valuable model to study radon-induced cancers, Petitot et al.
proposed a system which is designed to allow direct exposure ofisolated cell populations cultured in vitro to radon and its decay
products [P18]. One advantage of this is that cells cultured invitro are irradiated directly by a natural radon emanation and a
mixture of RDPs in exposure conditions similar to those used
for inhalation exposures of rats (in vitro studies). This new
method could help to identify biological markers of irradiatedcells in radon-induced cancers by an in vitro approach.
107. Overall, animal data support the conclusion fromepidemiology that exposure to radon and its decay products is carcinogenic. Moreover, animal data also confirm theobservations from epidemiology that the risk from exposure
to radon and its decay products increases with increasing
cumulative exposure, even for protracted exposures at low
exposure rates.
B.

Biomarkers

108. Traditionally, biomarkers have been used to determine
exposures or doses. Jostes [J14] provides an overview of
the use of biomarkers as measures of effects of exposure toradon. It is becoming increasingly apparent that biomarkers
of effect are potentially of great value in evaluating potential
health impacts. The following discussion provides an overview of selected studies of biomarkers involving ­exposure223

to RDPs. Readers interested in a more extensive discussionof this subject are referred to annex C, “Non-targeted anddelayed effects of exposure to ionizing radiation”.
109. In biodosimetry, the emphasis has traditionally beenon assessing doses per se; however, an important perspective
nowadays is to think of these biomarkers as possible surrogates for measuring the body’s integrated response to radiation damage, rather than as simply indicators of absorbed
dose. The focus of this section is on the potential use of
biomarkers of effects of radiation exposure rather than on
biological dosimetry per se.
110. In addition to interactions with DNA, ionizing radiation also damages other cellular components. Cellularresponses to various forms of radiation include structuraland functional changes to cells and cell organelles. In addition to the morphological signs related to cell death, several
reversible alterations are seen in the structure of different cellorganelles. Radiation-induced changes in the supra­molecularorganization of the membranes, including the plasma membrane as well as different cell organelle membranes, can playa significant role in the development of radiation effects.
Various morphological alterations of nuclear chromatin (e.g.
changes of fine structure, development of chromosome aberrations, etc.) are thought to originate from radiation-induced
damage to the supramolecular organization of DNA and/or
nuclear proteins [N3, S35]. Changes in chromosomes are
considered to be useful as biological indicators or even biological dosimeters of radiation injury [A26, A27, A28, A29,
B2, C3], and can be evaluated qualitatively and/or quantitatively by various techniques, such as morphological analysisof metaphase chromosomes, fluorescence in situ hybridization (FISH) and the scoring of micronuclei. ­Brenner et al.
[B32] argued that there was a need for a biomarker to distinguish between the effects of exposure to high-linearenergy-transfer (LET) radiation and other ­carcinogens. Theysuggested that exposure to high-LET radiation produced
a distinctly low ratio of stable inter­chromosome to intra­
chromosome aberrations, and recommended this ratio as a
candidate for such a biomarker.
111. Miller et al. [M7] exposed cultures of C3H 10T½ cells
either to microbeam irradiation using the Columbia University microbeam or to broadbeam irradiation in order toinvestigate oncogenic transformation rates, and demonstrated that cells exposed to exactly one alpha particle each
had a significantly lower response than cells exposed to a
Poisson mean of one alpha particle each. Miller’s group concluded that cells intersected by multiple alpha particles (asmay occur in high-exposure miner studies) contributed mostof the response. This suggested that extrapolating from thehigh-exposure conditions of miners could result in an overestimation of the risk of cancer induction at domestic levelsof radon exposure, in which situation essentially no target
cell is ever traversed by more than a single alpha particle.
112. Brenner and Hall [B10] commented on a pheno­menon
known as the inverse dose-rate effect. They noted that it is224

UNSCEAR 2006 REPORT: VOLUME II

widely observed for radiation with medium to high LET,
such as neutrons and alpha particles, and that a given dosedelivered over a longer time (i.e. protracted) had a greater
effect than an acute exposure at the same total dose. Theysuggested that the dose-rate effect was significant only whena certain combination of dose, dose protraction and radiation quality was present. When uranium miners are exposed
to low WLs for a long period of time, the possibility of the(inverse) dose-rate effect occurring exists, but such dose-rate
effects in cases of typical domestic exposures to radon areunlikely, because the average cell is traversed by one or zeroalpha particles in a lifetime. Hence when risk estimates fordomestic exposure are extrapolated using data from miners,
the radon risk may be overestimated.
113. Similar to the conclusion reached by Brenner and Hall
[B10], Jostes [J14] noted that epidemiological studies haveshown that radon is a risk factor for both smoking and nonsmoking miners, and that it is reasonable to suppose, on thebasis of molecular and cellular considerations, that exposure
to RDPs in the home also poses a risk of cancer induction.
He also suggested that, while an inverse dose-rate effect has
been seen in miners, such an effect may not apply to residential exposures, since the majority of lung cell nuclei would
experience no alpha hits and only a few cells would receive
even a single alpha hit.
114. Mutations induced in mammalian cells following
irradiation with alpha particles have been studied using
microbeam methods to irradiate the cytoplasm of individual
human–hamster hybrid cells (AL cells) [W11]. The ­aiming
point of the microbeam was 8 µm from the ends of the major
axis of each cell nucleus. The probability of a ­scattered alphaparticle hitting the nucleus was 0.4%. Irradiation of the cytoplasm produced gene mutations in the nucleus in a process
mediated by free radicals. While the microbeam irradiation induced minimal toxicity to the irradiated cells, it was
effective in inducing mutations in the nucleus. An analysisof the mutational spectra induced by nuclear versus cytoplasmic irradiation suggested that different mechanisms ofcancer induction were operative. Two approaches were used
to investigate whether reactive oxygen species (ROS) mediated the process of mutagenesis through cytoplasmic radiation. The first involved the use of an antioxidant (di­methyl
sulphoxide) to scavenge ROS, and the second involved
the addition of a drug (buthionine-S-R-sulphoximine) todeplete endogenous scavenger sulphydryl groups. In the first
approach, the induction of mutations was suppressed; in thesecond, the induction of mutations was enhanced. Indicators of cellular ability to scavenge ROS may have potential
as a biomarker of effect. Observations from in vitro studies­further demonstrate that cytoplasmic irradiation is clearly arisk factor.
115. An earlier study had also demonstrated that radiation
damage from alpha irradiation reflected in the patterns of sister chromatid exchange (SCE) in human diploid lung fibro­
blast cells was independent of the number of alpha particles
traversing a nucleus [D3]. The much larger cross-sectional

area of the cell relative to that of the nuclear target was suggested to be an important consideration. It was hypothesizedthat the effect was mediated by the free radicals formed in
the cytoplasm as the result of traversal by alpha particles. In
support of this hypothesis, a paper subsequently described
the production of superoxide anions and hydrogen peroxideby alpha particles traversing the cytoplasm [N1].
116. In recent years, evidence has accumulated that­suggests that both directly hit and bystander cells may show
effects of exposure to alpha radiation (e.g. [M36]). ­Goodwin
and Lehnert [G8] found from in vitro studies that a low doseof alpha radiation increased levels of SCE, an indicator ofgenetic damage in the lungs of mice and humans. In addition,
the amount of ROS increased. SCE and ROS levels were also
increased in non-irradiated cells. The authors concluded that
harmful effects of radiation are induced in bystander cells tothe same extent as in irradiated cells. There was evidence that
the ROS response triggered the up-regulation of cytokinesthat mediate the bystander effect [G8]. The authors anticipated that an important next step would be to show whether
this work applied in vivo and whether carcinogenesis andother disease processes were a result of the bystander effectfrom ionizing radiation.
117. Azzam et al. [A15] postulated that the reaction of
non-irradiated cells is due to gap junction-mediated intercellular communication (GJIC) of a damage signal between
irradiated cells and their neighbours. The levels of a protein(p21Waf1) induced by stress were compared in cells competent in GJIC and not competent in GJIC. In the competentcells, more cells showed elevated levels of the protein than
could have been directly intersected by an alpha particle. The
expression of p21Waf1 correlated with micronucleus formation(indicating DNA damage) and with increased phosphorylation of the p53 gene. While GJIC is certainly one mechanism
by which a signal is communicated from an irradiated cell toa non-irradiated cell, there is also evidence for soluble factors being secreted into the culture medium to communicate
the bystander signal [M43, S73].
118. A bystander effect and gene induction in non-irradiated cells occurred for situations where only a small fractionof the cell nuclei were traversed by an alpha particle [L5].
Oncogenic transformation can arise from the passage of single alpha particles through cell nuclei [M7]. The role of ROS
in mediating DNA damage and cell-related effects is of greatinterest in understanding bystander effects. The bystander
effect and other non-targeted effects are discussed at greater
length in annex C, “Non-targeted and delayed effects of
exposure to ionizing radiation”.
119. Little and Wakeford [L35] discussed radon-induced
lung cancer and the bystander effect in C3H 10T½ cells
exposed to alpha radiation. These authors fitted a model of the
bystander effect [B21] to experimental data [M7, S47] andto epidemiological data for various residential radon studies, for Colorado Plateau miners [L10] and for the combined
analysis of BEIR VI [C20]. The best estimate of the number

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

of neighbouring cells that contributed to the bystander effectwas between 0 and 1.0 (with an upper 95% confidence limit
of between 1 and 6.5), and therefore the bystander effectseen in the experimental C3H 10T½ cell system probably
did not play a large role in the radon-induced lung cancer
seen in the epidemiological studies of humans. The authors
also found that the ERRs in the Colorado Plateau miner data
were statistically indistinguishable from those derived fromresidential radon studies.
120. One potential biomarker of radon exposure is interleukin-8 [N2]. When alpha particles hit a cell, the production of ROS increases; this causes the production and releaseof interleukin-8 to increase. However, there are triggers for
ROS production other than alpha radiation; asbestos, ozone
and cigarette smoke can all cause an increase in ROS levels
and therefore in interleukin-8 levels.
121. In an early study of chromosome aberrations in­uranium miners, Brandom et al. [B30] studied cultures of
peripheral blood leucocytes in 15 uranium miners and 15 normal age-matched non-miner male controls. The ­miners’ work
experience ranged from 1 to 20 years, and cumulative exposures ranged from 10 to 5,400 WLM. Chromosome abnormalities were observed in 3 (0.28%) of the control cells and37 (2.3%) of the miner cells. The differences were statistically significant and were considered by the authors to bebiologically important.
122. Smerhovsky et al. [S58] described a study of chromosome aberrations in radon-exposed miners in the Czech
Republic. The study included 1,323 cytogenetic assaysof peripheral blood lymphocytes from 225 mine workers exposed to RDPs at levels ranging from about 1.7 to662 WLM. Seventy-five of the workers were reported as
non-smokers. In total, some 36 lung cancers were observedin this group. Kaplan–Meier survival analysis was used toinvestigate cancer incidences and Cox regression was used
to model associations between chromosome aberration frequency and cancer incidence. The Kaplan–Meier survival
analysis indicated a significant decrease in survival times
dependent on frequency of aberrant cells and on frequencyof chromatid breaks. The Cox regression showed that thefrequency of aberrant cells was significantly related to therisk of cancer.
123. A pilot study on German miners determined the­presence of biological markers related to radon exposure inuranium mines [P3]. The researchers looked at two categories
of markers: markers in the blood and markers in broncheo­
alveolar lavage fluid. They found that former miners’ leucocytes had a decreased ability to repair DNA. In addition,
there were chromosome aberrations in their lymphocytes andan increased frequency of micronuclei in lung macrophages.
There was an increase in the levels of tumour necrosis factor
alpha in the miners compared with the control group, andthis was weakly correlated to their radon exposure. Further
study is needed to establish whether these measurements
are reliable biomarkers and whether individual cancer risk

225

can be predicted on the basis of the markers. A study of therelationship between residential radon and the occurrence ofchromosome aberrations in peripheral blood lymphocytes
was reported in reference [O3]. This study demonstrated
an excess in the number of cells containing dicentric and/or centric rings for 61 people living in dwellings with radon
concentrations of above 200 Bq/m3 compared with a control
group from the researchers’ laboratory (53 people). However, no statistically significant difference was seen between
the control group and people exposed to radon concentrations of between 230 and 13,000 Bq/m3. Uncertainty in
exposure was suggested as a possible confounding factor forthis result [O3].
124. Another study done on German uranium miners tried
to identify a specific genetic defect caused by alpha radiation that led to lung cancer [W4]. The authors were unable tofind a mutation of the tumour-suppressor gene p53 that wascharacteristic of a radon-induced cancer.
125. Hussain et al. [H28] described studies of the mutability of codons 249 and 250 of the p53 gene patterns in normal
human bronchial epithelial cells from a 15-year-old male
who had never smoked. The cells were either unexposed
or irradiated to a total dose of 4 Gy (equivalent, according to the authors, to 1,460 WLM of exposure to RDPs). In
this study, exposure was from alpha particles from 238Pu in
six equal fractionated doses. The authors found that alpharadiation selectively increased mutation frequency in bothcodons 249 and 250 but noted that interindividual variability
argues against extrapolation of their results based on a single
donor.
126. Yngveson et al. investigated the association between
residential radon exposure and p53 mutation in lung tumours[Y6]. Their study included 83 lung cancer cases in nonsmokers and 250 lung cancer cases in smokers. Lung cancer cases were selected with time-weighted average radon
concentrations of below 50 Bq/m3 or exceeding 140 Bq/m3.
Molecular analysis was carried out on samples obtained fromthe pathology departments where the cancer cases had been
diagnosed. Statistical analysis was carried out to investigate
associations between exposure to radon, tobacco consumption and the presence of p53 mutations. A non-statisticallysignificant odds rates (OR) for increased mutation prevalence
was indicated for those exposed to high levels of residential
radon (OR = 1.4; 95% CI: 0.7, 2.6), especially among nonsmokers (OR = 3.2; 95% CI: 0.5, 15.5). Mutations of p53
were also found to be associated with smoking status and,in the case of non-smokers, with exposure to environmental
tobacco smoke.
127. Although biomarkers such as mutations in the TP53
gene (which encodes the p53 protein) are recognized asimportant biomarkers of radiation damage, not all studies
arrive at this conclusion. Besides the study in [W4], a more
recent study of p53 in serum samples from former uranium
miners found no correlation between p53 protein concentrations in serum and exposure to ionizing radiation (measured

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UNSCEAR 2006 REPORT: VOLUME II

in WLM), and the authors concluded that there was no benefit in screening for p53 or p53 antibodies at the present time[S7]. A study by Vahakangas et al. [V3] of lung tumours inminers found that their TP53 gene mutations were different
from those typically seen in lung cancers caused by tobaccosmoke: there were no G:C to A:T transitions in the coding
strand, and the mutations were mostly transversions andsome small deletions, which are very uncommon in humanlung cancers.
128. Albertini et al. [A2] studied the viability of using
HPRT mutations in human T-cells as markers of the qualityof radiation to which a person was exposed. They lookedat the different mutations produced by high- and low-LET
radiation. With high-LET radiation such as that from radon,
the mutations were mostly small partial deletions, withfewer than 2% being total gene deletions. There were relatively more breaks in the DNA strand because of the energy
transferred from the alpha particle to the strand. A high proportion of the breaks were lethal, so alpha radiation is efficient at killing cells. In the case of low-LET radiation, 10%of the mutations were total gene deletions and most of thedamage to the DNA molecule occurred because of secondary ionization, not from the initial collision. These or relateddifferences may eventually allow researchers to differentiate
between high- and low-LET radiation.
129. Alpha particles, like other high-LET particles, induce
the tumorigenic phenotype into BEP2D cells in the lungs[Z1]. This is achieved by deleting suppressor genes. Thus,by looking at the levels of the products of these genes, it
may be possible to detect the early development of cancerous cells.
130. Jostes [J14] suggested that the cellular response
to alpha radiation may depend on the repair status of theaffected cell. This was based on studies by Schwartz et al.
[S8] and Shadley et al. [S32], which compared the inductionof chromosome aberrations in repair-proficient versus repairdeficient cells. In reference [S8], it was found that a cell linethat was repair-deficient for single-strand DNA breaks wasmore prone to aberration induction than the parental repairproficient cell line. In reference [S32], the cell line used was
repair-deficient with respect to double-strand DNA breaks,
and it was less prone to aberration production when alphaparticles were used than when X-rays were the inducers. It
also appears that there is an adaptive response to radiation.
Jostes [J14] reported that when human peripheral blood lymphocytes were exposed to a small priming dose of low-LET
radiation and later to a challenge dose from radon, the numberof chromosome aberrations was smaller than expected [W8,W9]. However, when alpha particles and X-rays were usedat the same time, the number of micronuclei was higher than
the anticipated additive effect [B4]. Taylor et al. [T1] suggested a p53 mutation hot spot in radon-associated lung cancer, noting that Jostes [J14] had also compiled results fromvarious studies on mutations at the HPRT locus in Chinese
hamster ovary cells exposed to alpha radiation. In general,
half of the mutations were complete gene deletions, with theother mutations composed roughly equally of partial deletions and rearrangements (~25%) and undetectable changes(~25%). The proportions changed depending on cell type[J14].
131. In a study using comparative genomic hybridization,
Dano et al. investigated gains and losses of genetic material in a series of radon-induced rat lung tumours [D13,D19]. Frequent losses occurred at various locations homologous to human chromosome bands. These regions are frequently (30–80%) deleted in human lung cancer and containtumour suppressor genes or proto-oncogenes such as MET,CDKN2A, CDKN2B, FHIT and RB1, and genes yet to be
identified. Frequently observed gains involved chromosomes
homologous to those in human encoding MYCN and MYC
oncogenes. The genetic similarities between rat and humanlung cancer suggest common underlying mechanisms fortumour evolution in both species and provide an opportunity
to study early events in carcinogenesis. Moreover, cytogenetic and molecular genetic analyses of radon-induced rat
lung tumours could help to better understand the development and progression of radon lung cancer in humans.
132. Smoking remains the predominant cause of lung cancer. The effects of smoking per se and of exposure to environmental tobacco smoke (ETS) [A24] are of great interestfor both miner and residential epidemiological studies ofradon [A23]. On the basis of in vitro studies, Piao and Hei[P7] suggested that the combined effects of radiation andsmoking are additive for both low- and high-LET radiation.
133. Bennett et al. [B31] reported a molecular epidemiological study of gene–environment interactions in promoting
lung cancer in women who never smoked. The study was
designed to assess the risk of lung cancer from exposure to
ETS, radon and dust, as well as family history and occupational exposure in 106 white women with lung cancer who
never smoked. The authors also carried out genetic analysis for cancer susceptibility genes. Odds ratios and 95% CIs
were calculated by multiple logistic regression. A dose–
response relationship was found between ETS exposure andincreased lung cancer risk in women with a common genetic
deficiency, loss of GSTMI enzymatic activity, with the trend
test significant at the 2% level. This suggested that ETS exposure could possibly double the risk of lung cancer in nearly
half of the white women in Western nations. According to­Bennett, loss of GSTMI enzymatic activity occurs in about
50% of the white population in Europe and North America.
134. Alavanja [A23] reviewed biological damage fromexposure to tobacco smoke and from radon, and concluded
that strong similarities existed between the biological­damage caused by the two agents. He also noted a protective effect arising from cruciferous vegetable consumption,
at least in part attributable to their antioxidant properties, inboth smokers and non-smokers [A23].
135. Traditionally, biological effects from irradiation of
a population of cells are considered to arise as a result of

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

unrepaired or misrepaired damage to DNA in irradiatedcells. It has been noted previously in this section that radiation effects can also occur in non-irradiated cells (e.g. [L5,
L35, M7]), a phenomenon commonly referred to as the
bystander effect (e.g. [M36, M37, M38]). As the bystander
effect is thoroughly discussed in annex C, “Non-targeted anddelayed effects of exposure to ionizing radiation”, a detailed
discussion here is unnecessary, other than to note that while
the existence of the bystander effect raises many questions
about traditional dosimetric modelling, risk factors derived
from epidemiology will already implicitly take into accountany contribution from the bystander effect, thereby supporting the use of epidemiological evidence for risk estimation.

227

136. At the present time, considerable uncertainty
remains about the effect on risk estimation of the “new”
biology outlined above and discussed in greater detailin annex C. Future dosimetry will need to consider theeffect of mechanisms of carcinogenesis such as those discussed above. In the meantime, it should be noted thatthe effects of such mechanisms are implicitly includedin the results of epidemiological studies. Although several potential biomarkers of radon exposure have beenstudied, chromosome aberrations appear to be the mostpromising at this time, owing particularly to the possible“signature” of high-LET exposures and the correlation
with cancer risk.

IV.

EPIDEMIOLOGICAL STUDIES OF MINERS

137. Studies of underground miners exposed to radon
form the current basis for estimating risks from radon and its
decay products. The UNSCEAR 2000 Report [U2], BEIR VI[C20] and others have reviewed the epidemiological studiesof miners, but scant attention was given to the basis for the
exposure estimates. Miner studies are reviewed in this section. Particular attention is given to the sources and effects ofuncertainty in estimates of risk to miners.
A.

Service (USPHS), in cooperation with the United States
Atomic Energy Commission and the state health departments of Arizona, Colorado, New Mexico and Utah began
studies of exposures in uranium mines. By 1950, medical
studies had been initiated [A12], and uranium ­miners were
subject to routine medical examinations. Beginning in 1954,
medical examinations were performed every three years on
all uranium miners who could be reached and who agreed toundergo examination [C11].

United States: Colorado Plateau miners

142. The initial study population, some 90% of these socalled “Colorado Plateau” miners, consisted of all miners
examined in 1954. The USPHS began to collect data onradiation exposure, smoking history and mortality for theseminers. Several analyses of these data, with different periods of follow-up, were published [A4, A10, A12, A16, A17,
H17, L13, L14, L15, W1, W2, W3, W12].

1. Introduction
138. The discovery of radioactive ores in the Colorado
Plateau dates to between 1881 and 1887. The ores contained vanadium, uranium and a small quantity of 226Ra[H13]. Before there was a demand for uranium and radium,vanadium was mined on a small scale. Radium production became more important during the period 1916–1923.
However, United States radium production eventually lostits competitiveness owing to the availability of radium fromhigh-grade ores from the Belgian Congo, and during theperiod 1930–1945, the vanadium content of the Colorado
Plateau ores was the principal objective of their mining. Subsequently, through the defence initiatives associated with theSecond World War, the emphasis turned more towards themining of uranium [L13].

143. Holaday [H11] recommended that all states adopt a
tentative “working level” of 10–10 Ci/L of radon in equilibrium with its decay products [S31]. For practical purposes,
the recommended standard was equivalent to 1 WL, or
12 WLM per year of exposure. Holaday indicated that thisrecommendation was adopted as a guide by official agencies in most of the uranium mining states. Holaday [H11]documented decreasing exposure to RDPs in the Colorado
Plateau mines: in 1961, only 21% of the mines studied had
WL measurements of below 1 WL, but by 1967, 70% of themines fell in the <1 WL range.

139. The uranium mining industry had expanded somewhat by 1949, and by 1950, some 500 miners worked in the
Colorado Plateau area mining uranium ores, mostly in small
underground workings with an average production of about
1 ton (907 kg) per person-day [H12]. According to ­Holaday[H12], employment peaked in 1960, when about 5,800 underground miners were employed. By 1967, employment had
declined to approximately 2,800 miners, who were then producing approximately 3 tons per person-day. ­Holaday notedthat by 1967, the occupational health field station had some
information on exposures in over 1,200 different ­mining
operations; however, an unknown additional number ofoperations were never surveyed.

144. Before 1964, the road systems serving the mines in
the Colorado Plateau area were inadequate, and since mostmines were located in remote areas, supplies of fresh waterand electricity were often insufficient [H13]. Holaday andDoyle hypothesized that this insufficiency of electrical
power contributed to high RDP levels during the early daysof mining [H13]. In those days (early 1950s), the mines weretypically shallow [H11], were usually entered by a horizontal
adit or incline, and were almost always ventilated by natural
draught only. In most mining operations, the miners would
work throughout the mine in a variety of activities — drilling, mucking, handling the ore or setting timbers, etc. Thispattern of working was established when the mining was forvanadium.

140. Cooper [C11], writing in a special supplement of the
Journal of Occupational Medicine, noted that in 1953 or 1954,
large deposits of primary uranium ores were discovered in theMoab, Utah, area and in the Grants area of New Mexico.

145. Uranium was identified in ores in the Shiprock areaof New Mexico around 1918. At that time, it received little
attention. By 1950, however, there was considerable interest in uranium, and various uranium outcroppings werediscovered in limestone and sandstone areas in the Grants

141. In 1949, as a result of concern about the possible healthhazards of uranium mining, the United States Public Health
229

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UNSCEAR 2006 REPORT: VOLUME II

mineral belt [S5]. At various times during the period 1950–
1960, some 60 mines were in operation. Uranium market
fluctuations starting in the early 1960s resulted in various
expansions and shutdowns. In the period 1966–1978, theannual output of uranium in the Grants area of New Mexico
represented approximately 45% of United States uranium
production.

1955–1968, the total number of mine-years for which there
were more than five measurements was 1,313 [L13]. An
earlier study by the Advisory Council of the United States
National Academy of Sciences [N4] observed, “Exposure
values assigned to the period before 1956 are highly unreliable, being based almost entirely on estimates rather than
measurements of concentrations” ([N4], p.7]).

146. The New Mexico Health Department began monitoring RDP levels in New Mexico mines in the late 1950s,
with the aid of the state mine inspector’s office. According to Samet et al. [B47, S1, S3, S4, S5, S74], the maximum permissible concentration of RDPs decreased from 25
to 10 WL in 1960, to 5 WL in 1963, to 3 WL in 1967, to
1.75–2 WL in 1969, to 1.4 WL in 1973, and finally to 1 WL
in 1976. The Grants clinic, which opened in 1957 to serve
the miners, handled 80–90% of the pre-employment andfollow-up medical examinations of miners, and kept recordsof the movements of most miners in this area. The majorityof the investigations of New Mexico miners were examinations of morbidity and mortality among uranium miners andparallel investigations of the miners’ exposures, with information on the exposures being published in the third annual
report of the study by the University of New Mexico [S4]. A
1989 report provided the results of a case–control study toinvestigate lung cancer risk in a cohort of the New Mexico
underground uranium miners [S1].

149. The exposure estimates used by Lundin et al. [L13]were based on RDP concentrations derived in one of threeways:

2. Radon and radon decay products
147. According to Lundin et al. [L13], radon gas measurements in United States mines were first made in 1949, and
RDP measurements were first performed in 1951. In 1952,
an effort was made to survey all operating uranium mines,
and during that year, RDP measurements were made in 157
mines and radon gas measurements in 79 mines. In 1953–
1954, the number of surveys was limited, but by 1955, manyof the larger uranium mining companies initiated their own
air sampling programmes. By 1956, the mining companies
were performing most of the mine survey work, while agencies continued their own control programmes [L13]. These
measurements were carried out across the four states of Arizona, Colorado, New Mexico and Utah.
148. Lundin et al. [L13] emphasized the large number ofmeasurements made in the mines: “From the entire 1951–
1968 period nearly 43,000 measurements of RDP concentrations were available to characterize the approximately 2,500
uranium mines from which ore was shipped” (p. 31). However, this represented typically less than one measurement permine-year averaged over 18 years of study. The distributionof these WL measurements, moreover, was far from uniform(from no measurements to a fairly large number of measurements in a single mine). Lundin et al. tabulated the numberof mines in which five or more RDP measurements weremade in any one year. This amounted to 116 mine-years forthe period 1951–1954, during which more than half of the
cumulative exposure of miners was received. For the period− “Measured”. Values were derived directly fromone or more measurements of RDP concentrations in any given mine in a given calendar year.
Before 1955, such measured data included only a
few of the work areas in the mine.
− “Extrapolated”. Extrapolated concentrations wereobtained by extrapolating between measurements
made in the same mine in other years, if not morethan two years before or after the year to which
the extrapolation applied. In some cases, regional
extrapolations were also made for mines having
no direct measurements; this was done by assigning a concentration value equal to the average forthe other mines in the vicinity.
− “Guesstimates”. These were provided by Holaday
for mines in which RDP concentrations could not
be obtained by any of the extrapolation techniques
discussed in Lundin et al. [L13]. The “guesstimates” were based on general knowledge of theairborne concentrations that occurred in similar
mines during the same time period and on expert
knowledge of long-term trends.
150. The miner database of the United States Public HealthService/National Institute of Occupational Safety and Health(USPHS/NIOSH) contains a great deal of information onminers’ exposures [S13]. There are significant temporal andspatial trends in RDP concentrations. Figure III illustrates
the decrease in RDP concentrations over time. The figure
is a plot of annual RDP concentration (WL) by years, estimated using the different methods described by Lundin et al.
[L13]. The WL values estimated by the different methods
are approximately comparable and show a similar temporaldecrease. The high measurement average in 1951 was due toa high WL (895 WL) averaged over 8 samples at the Freedom Mine in Salt Lake, Utah. Figure IV illustrates the extent
of variability in measured WL values for Colorado Plateau
miners.
151. Given the uneven distribution of measurements on a
per-mine basis and the large variability between mines, it
may not be possible to estimate exposures for a specific year
in those mines where no measurements were taken. SENES[S13] estimated variabilities at the four levels of “area estimate” by the pooled standard deviation of measurements,
under the assumption that there was no spatial or temporal
trend in the variation of exposure rate. At each level (i.e.

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

231

Figure III. Average radon decay product concentration for underground uranium mines in the Colorado Plateau area from1950 to 1968, inferred using different methods [L13].

AVERAGE RADON DECAY PRODUCT CONCENTRATION (WL)

1 000

100

10

1

0.1
1950

1955

Legend1960
YEAR
Measurement
Guesstimate

1965

1970

Extrapolation

Figure IV. Range of average radon decay product concentrations measured in underground uranium mines in the Colorado
Plateau area [S13].

AVERAGE RADON DECAY PRODUCT CONCENTRATION (WL)

1 000

100

10

1

0.1
1950

1955

1960
YEAR

1965

1970

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UNSCEAR 2006 REPORT: VOLUME II

local, district, etc.), only areas that met the criteria of having
more than three mines and 10 samples were included in thecalculation; the natural logarithm of measured concentration
Table 6was used with sample size as a weighting factor. The results
are listed in table 6 together with the equations used in thecalculations to rate the large variability in WL values.

Estimated variabilities at different levels of area estimate [S13]Level of area estimate

SymbolStandard deviation of natural logarithm (WL)

Localitya

SL

3.16

Districtb

SD

3.23

StatecSS

3.39

Coloradod

SC

3.66

(E − E )

2

S =
2
X

m

α
α

mα

( Ny − 1)

α

,

where
E is the natural logarithm of measured concentration (WL);
is an average over E weighted by sample size;
N is the number of measurements;
X = L, D, S and C, denoting locality, district, state and Colorado levels, respectively;α = (ly), (dy), (sy); m = mine; y = year.
Summation over α denotes summation over both indices indicated in parentheses. In the case of the Colorado level, summation over α reduces to a single summation
over variable y only.
a Locality: reflects variability among mines within the local mining area.
b District: reflects variability between mines within an entire mining district.

c State: reflects variability between states (Arizona, Colorado, New Mexico, Utah).
d Colorado: reflects variability among Colorado mines.

152. Schiager and Hersloff [S6], reporting on interviews
with Holaday, noted that, while large uncertainties existed in
the estimates of the RDP concentrations, these were the bestthat could be made and there was no conscious bias injected
into the estimates of the early exposure conditions. This is
in contradiction to other statements, which indicated that,
where uncertainties existed, WL were deliberately overestimated [L13]. If, on average, exposures were overestimated,then on average the risk per unit exposure would be under­
estimated. As noted previously, mining companies began
conducting the majority of exposure measurements after
1956. Lundin et al. [L13] felt that during this period, i.e. the1950s and 1960s, there was a possibility of a bias towards
underestimation of exposure, and therefore these authors
tried to avoid this potential bias: “...efforts were made to
exclude company measurements from data after 1960 fromuse in the epidemiological study of uranium miners” (p. 31).
However, since the bulk of the exposures upon which risk
estimates are based occurred prior to 1960, this selection hadvery little impact on the results overall.

more reliable. However, there is still likely to be considerable uncertainty in exposure assessment up to what one mightrefer to as the “modern period”, i.e. 1967 to the present.

153. In summary, most WL exposure estimates for Colorado
Plateau miners for the period before 1950 seem unreliable, as
essentially no data exist and all the estimates were based on
extrapolation or “guesstimates”. Exposures estimated for theperiod 1950–1956 were considered more reliable, althoughstill highly uncertain. After 1956, the mining companies
themselves started to make measurements on a systematic
basis and, at least in the large mines, measurements became

3.

Exposure estimation

154. Wagoner et al. [W2] concluded that the excess respiratory cancer rates among uranium miners were not attributable to age, smoking activity, heredity, urbanization, selfselection, diagnostic accuracy, prior hard rock mining orore constituents. They attributed the excess risk to airborne
radiation.
155. Miners who were examined in the Colorado Plateau
area during the period 1951–1960 and for whom sufficient
records existed made up the cohort studied by Wagoner et al.
[W2]. The exposure calculation was performed individuallyfor each miner; the miners were then categorized into groups
according to exposure levels. The average RDP concentration (WL) and duration of work underground in working
months (one working month is taken as 170 hours) were multiplied to arrive at average exposures (WLM) for each groupof miners; additional details were provided in a subsequent
paper by Wagoner et al. [W1]. At that time, approximately
12,000 RDP measurements were available for the approximately 1,200 mines under study. The RDP concentrations
used in the calculations were derived as follows:ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

“When there were multiple measurements in the mineduring a calendar year, an average annual exposure wascalculated. Measurements from non-work areas were
excluded from all calculations. When no measurements
were available for a mine during a calendar year, estimates were made from the average of measurements
for the same mine during the preceding and following
calendar year or other mines on the basis of geographic
proximity, similar ore bodies, physical layout, ventilation and control efforts of regulatory agencies” ([W1],p. 184).
156. More detailed measurements began in 1967 with theintention of presenting these data for review by the Federal
Radiation Council (FRC) [F2] so that the FRC guidelines forthe control of radiation hazards in uranium mining could be
updated. The results of this update were presented in a paperby Lundin et al. [L14], who examined a cohort of 3,414 whiteand 761 non-white underground uranium miners who had
undergone medical examinations in 1950–1961. The results
indicated that uranium miners who smoked had an excess
lung cancer risk ten times greater than non-smoking miners.
Prior hard rock mining experience had little effect on lungcancer mortality overall but was suspected of contributing
more significantly to risks in the lower exposure categories.
157. In 1971, Lundin et al. [L13] reported on the cohort
with exposures updated from the start of mining up to theend of September 1969. The cohort contained 3,366 whiteand 780 non-white uranium miners with at least one monthof underground mining prior to 1 January 1964. Procedures
were also developed by Lundin et al. [L14] for estimating
RDP exposure from hard rock mining other than uranium
mining.
158. As previously noted, radiation levels in uranium
mines dropped sharply after 1967. This fact, combined witha drop-out rate of 10%–50% per year for the original cohort
after 1960, justified the assumption that exposures received
after September 1969 contributed only a relatively small
additional exposure to this cohort.
159. According to Lundin et al. [L13], more than 50% of
the collective exposure of some 2.8 × 106 person-WLM wasreceived prior to 1955. For the calculated cumulative exposures (expressed in person-WLM) received up to 30 September 1960 by 3,325 men, more than 25% was received by
1,325 men and was based on actual measurements [L13].
Although the converse was not explicitly stated, it may beinferred from reference [L13] that, for the other 2,000 men
in this cohort, less than 25% of the cumulative exposure wasbased fully on measured data. This emphasizes strikingly theuncertainties in the early exposure data; however, it does notnecessarily imply any bias in exposure estimates and hencein risk estimates. The duration and periods of exposure ofindividual miners were determined in some cases from
employment records, but in most instances they were determined from interviews with the miners themselves. Annual
exposures for full-time miners were assigned on the basis of233

RDP concentrations averaged throughout the mine in which
each miner worked. For calculating annual exposures, “itwas assumed, unless we had information to the contrary,that a man worked in the mine at which he was found forsix months before and six months after the census or questionnaire date (6 month rule). When a man was known to
work in two different mines at less than a one year interval,
the period of employment during the interval was equallydivided between the two mines” [L13].
160. Full-time mining was assumed to consist of 12 fullmonths underground with no adjustment for vacation or illness. This assumption was an important reason for believingthat exposures were overestimated; however, no adjustments
were made for overtime work or ‘moonlighting’ by any of theminers [S6]. People familiar with the uranium mining industry in the 1940s and 1950s recalled that many miners worked
exceptionally long hours. A standard working week until
1960 was at least 48 hours. Many miners working entirely
under production contracts spent 50–60 hours undergroundeach week. It was not unusual for a miner to work a regularfull-time shift for an established company and then spend hisdays off developing a mining claim of his own. Since estimates of cumulative exposures (i.e. WLM) were based onmonths rather than actual hours worked, the method tended
to underestimate exposures. It is likely that any potential bias
in exposure estimates resulting from not making allowancesfor vacation and sickness was more than compensated by
not making allowances for underground time exceeding thenormal (48 hour) working week.
161. The 1968 report of the National Academy of Sciences[N4] suggested that the problem of determining radiation
exposure for individual uranium miners was also complicated because official mine records did not necessarily show
the actual job assignment. Only the miner himself, and toa lesser extent his immediate supervisor, knew the areas inwhich he worked. The report also raised the possible problem of exposure from previous mining, and further noted,“there is some uncertainty in the average working values
even in mines in which numbers of measurements havebeen made.” Measurements consisted of spot sampling at
a particular time and location, and therefore reflected only
the conditions that existed at that time and location. (This isimportant since exposure in WLM is the product of time inworking months, a working month nominally being taken
as 170 hours, and RDP concentration in WL.) The report
recognized that workplace conditions (dust levels, WL, etc.)
varied with the nature of the operations being carried out,
e.g. blasting, the ventilation provided and the amount of ore
uncovered.
162. According to a review by the USPHS epidemiologicalstudy ([J10], p. 1,266), the basic information examined forindividual uranium miners was the following:− Mines worked;− Dates of employment in mines;− Fractional time in mines;

234

UNSCEAR 2006 REPORT: VOLUME II− Radon decay product concentration levels measured or estimated for each mine as a function of
calendar year.
From these data, cumulative WLM values were calculated
for each individual miner. Cumulative WLM values were
calculated for both underground uranium mining alone andfor uranium mining plus other hard rock mining. The average exposure of the group of miners in the USPHS study wasof the order of 800 WLM [C18].
163. Some effort was made to estimate the exposure ofminers to RDPs accumulated since 1967. As reported inreference [S13], exposure data were collected and examined at 23 separately managed large mining operations andalso for more than 180 small mines representing in totalsome 25,000 miners. The major conclusions of this study(for 1967–1985) were: the average working time for theunderground uranium miners studied was 3.3 years, andwas 5.9 years for those who worked more than 2 years; theaverage lifetime exposure (i.e. 1967–1985) of undergrounduranium mine employees was 3.6 WLM; and the average
yearly exposure per employee who worked in undergrounduranium mines in this period was 1.2 WLM.
164. Many New Mexico miners were included in the cohort
study above. An independent epidemiological study of New
Mexico uranium miners was initiated in 1977. This study
was performed by the University of New Mexico under the
direction of J.M. Samet. The focus of Samet’s study group
was retrospective analyses of a cohort of 3,055 underground uranium miners whose first underground experienceoccurred prior to 1971. An outline of the approach used toestimate exposure in the New Mexico study is provided inreferences [M34, S4, S75]. According to Samet et al. [S75],
two large databases were developed to profile exposures to
RDPs of Grants area miners:“The first comprises WLM measurements from 1957
to 1967. A total of 20,086 individual readings areavailable from the 186 visits made during the 11 years;
after 1960, mine index values, which weight individual
measurements by number of persons exposed, weregenerally reported. The second includes all individual
WLM reports by companies for 1967 through 1982”[S75].
165. The mean WL values reported by Butler et al. [B47]
were 42 in 1954, decreasing greatly to about 4.3 in 1955, andsubsequently generally declining to about 1.7 in 1967. Theseauthors also reported that the mean annual WLM exposure
was below 10 from 1969 onward [B47]. It is perhaps worth
noting that the (weighted) mean of the WL measurements
was found to provide the best indicator of the total mineindex values, which weight measurements by numbers of
personnel exposed [B47, L35, S1, S4].
166. The employment history of United States miners was
documented by various means. Miners were interviewed

about mining experiences during medical examinations.
In addition, supplemental information from subsequentannual uranium miner censuses, records of official agencies and mail questionnaires was available [L13]. Efforts
were also made to account for other hard rock (OHR)­mining experience, as many of the underground uranium
miners in the Colorado Plateau had previous mining experience. For example, some had hard rock mining experience from other western mines, while others were coalminers from the eastern United States. However, the lackof radon or RDP measurements from these other mining
activities may result in an underestimate of the miners’
actual exposure.
4.

Other hard rock mining exposures167. Radon is a normal constituent of mine atmospheres,
including those of non-uranium mines. For example,­Holaday [H14] reported the results of several early measurements of radon in non-uranium mines in Colorado and
New York. Radon levels surveyed in 35 metal and clay
mines in Colorado ranged from 10 to 2,100 pCi/L. In the
New York mines, radon concentrations ranged from 2 to
110 pCi/L. According to SENES [S9], about 40% (1,433
out of 3,359) of white male Colorado Plateau miners accumulated radiation exposures from OHR mining. To investigate this, exposures from OHR mining were calculated
for miners whose exposures in uranium mines were lessthan 120 WLM. Following the procedure employed in
the USPHS investigations as described by Lundin et al.
[L14, L15], a miner was assumed to have completed his
OHR mining activity by the year preceding the start of
his underground uranium mining. Exposures were basedon assumed RDP concentrations in OHR mines of 1.0 WL
prior to 1935, 0.5 WL from 1935 to 1939 and 0.3 WL from1940 onward.
168. When these OHR mining exposures were added
to the exposures from uranium mining, the dates usedin the previous analysis (i.e. the dates when each miner
reached the limit of each exposure category) had to berecalculated. The assumed starting date of OHR mining
was calculated on the basis of the total exposure and theappropriate rates. Next, from the OHR mining exposure
rate and the known dates at which different uranium mining exposure levels were reached, the miner’s exposure
history was classified according to exposure rates andthe number of months at each rate. The exposures wereassumed to be accumulated at a constant rate during eachperiod at a particular exposure level. The dates on which
each miner reached a certain level of accumulated exposure were seen to be earlier when the OHR data weretaken into account. The addition of OHR mining exposures caused a shift of the miners, and hence the lungcancers (and person-years), into the higher exposure
categories. The average uranium and OHR mining exposures from the NIOSH cohort (with follow-up through1985) are shown in table 7 [S13].

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

Table 7235

Average uranium and OHR mining exposures by exposure category — white miners [S13]

Exposure category for
uranium mining (WLM)

Average time spent
underground(years)Average exposure for
uranium mining (WLM)

Average exposure for
uranium and
OHR mining combined(WLM)

Average exposure rate while underground(WLM/year)

<120

1.11

51.7

68

46.58

120–359

3.33

233.4

249

70.09

360–839

5.24

572.6

594

109.27

840–1 799

7.47

1 229.6

1 252

164.61

1 800–3 719

10.66

2 515.5

2 534

235.98

3720+

14.22

5 787.3

5 808

406.98169. Although the impact of OHR mining on the overall cohort average WLM is small, the effect for individual­miners can be appreciable. Table 8 summarizes the estimated exposures of miners from uranium mining only andfrom OHR mining, for those whose uranium-only exposure was estimated at less than 120 WLM [S9]. When OHR­mining exposure was added, 5 of the 10 miners shiftedfrom the 0–120 WLM category into the next higher exposure ­category. The average exposure of this group increased

Table 8from 51 to 101 WLM. These data suggest that, particularlyin the low exposure categories that are of greatest interest
to present-day miners, a potential bias towards underestimation of exposure is introduced if OHR mining exposure isneglected. In the absence of an agreed procedure for estimating OHR mining exposures, studies could be done either
on miners with no pre-1950 exposure (which minimizes thiseffect) or with the use of categorical variables denoting thepresence or absence of pre-1950 exposure.

Effect of OHR mining for lung cancer cases with less than 120 WLM uranium mining exposure [S9]SENES ID number

Uranium mining exposure (WLM)

OHR mining exposure (WLM)

Total exposure (WLM)

1911

8

0

8

1011

13

136

149

1092

28

0

28

4017

30

0

30

3321

33

0

33

3633

42

0

42

3297

42

187

229

3588

44

62

106

2023

68

29

97

1563

73

68

141

50

82

0

82

534

83

122

205

560

119

44

163

Average

51

50

101

170. Limited data are available on the characteristics ofmine atmospheres in the past. For example, the attachedfraction of RDPs will depend on a number of characteristics of the mine atmosphere, among them WL, dust levels,
dust particle size and humidity. Such data are important for
both dosimetry and epidemiology. Useful data are summarized in various publications, including references [C10,G1, N7, N10, S23]. A 1957 report on the control of radonand its decay products [H27, U16] compared the operating
conditions in uranium mines with those in conventional hardrock mines. While few data were presented, the report statesthat “dust counts in uranium mines indicate concentrations
from 5 to 20 million particles per cubic foot of air” and “the­silica content of uranium ore (carnotite) ranges from 50 to236

UNSCEAR 2006 REPORT: VOLUME II

75 percent” [H27]. It is important to remember that the
majority of measurements in mines are from the mid-1960s
onward, and that the major portion of exposure of miners
dates from earlier times when few actual measurements weremade.
5. Epidemiological analyses
171. In 1950, the United States Public Health Service(USPHS) began to collect various data on uranium miners in
Arizona, Colorado, New Mexico and Utah, including radiation exposure, smoking history and mortality. Numerous
analyses of the data have been published [G4, H9, H17, L13,
L16, W12]. The miner data were incorporated into analyses
carried out by BEIR IV [C19], Lubin et al. [L10] and BEIR
VI [C20], among others.
172. Hornung and Meinhardt [H17] reported a proportional hazards analysis of the Colorado Plateau cohort, originally described by Lundin et al. [L13], with follow-up to
31 December 1982. The cohort consisted of 3,366 (white)
miners with 256 lung cancer deaths, a median (cumulative)
exposure of 10.3 WLM and a median duration of employmentof 48 months underground. This study evaluated several risk
models and chose a power function model since it provided
the best fit to the data, and permitted analysis of the effects ofseveral temporal factors and smoking. The study estimatedexcess relative risk (ERR) to be 0.9–1.4% WLM–1, comparedwith a previous smoking-related estimate of 0.31% WLM–1
reported by Whittemore and McMillan [W12]. The relative risk increased with age at exposure and decreased withincreasing time since exposure (a reduction of about 55%
10 years after cessation of mining compared with miners
with the same exposure, smoking history and age) [H17].
Hornung et al. [H9] conducted further studies of modifiersof lung cancer risk in the Colorado Plateau miner cohort;follow-up to 31 December 1990 added an additional 121
lung cancer deaths, bringing the total number of lung cancer
deaths to 377. This analysis confirmed the earlier finding ofa strong interdependence of relative risk and age. It also suggested an exposure-rate effect resulting in a concave downward dose response; i.e. the relative risk was reduced at low
exposure rates and low cumulative doses. For risk estimation, the authors recommended the use of ERR after stratification to lower exposure rates (<10 WL) and cumulative

Table 9exposures (<800 WLM). The authors also briefly discussed
the potential for errors in the exposure estimates to bias epidemiological analysis, and suggested that such errors might
result in the underestimation of the ERR.
173. A 1992 report [S9] described exploratory analyses
performed on white male underground uranium miners withfollow-up to 31 December 1985. The purpose was to investigate the effect of other hard rock (OHR) mining experience and smoking on the risk of lung cancer. The report
also included an analysis of a subgroup of miners withcumulative exposures of below 2,000 WLM, using Poisson
regression to estimate model parameters using iteratively
reweighted non-linear regression. Simple linear ERR andabsolute risk models were among the dose responses evaluated. The results are shown in table 9. While the regressions in table 9 are all significant (p < 0.01), on the basisof the F-ratio (a statistical test to see whether the amountof variation explained by the regression is significant), theabsolute risk model accounts for more of the variability inthe data for all miners and smoking groups than the relativerisk model. The effect of OHR mining and the joint effect of
OHR mining and smoking were also investigated (table 10).
The miner groups included all of the miners with no exposure from OHR mining, those with OHR mining experience but no uranium mining exposure prior to 1950, thosewith neither OHR mining experience nor pre-1950 exposure in uranium mines, and those with only OHR mining
experience. All of the regressions in table 10 are significant(p < 0.01). On the basis of the F-ratio, the results in table 10show that, in this case, the relative risk model accounted
for more of the variability in the data than the absoluterisk model. The absolute risk to miners who never smoked(2.8 × 10–6 a–1 WLM–1) was about half that calculated for“all” smokers (6.6 × 10–6 a–1 WLM–1). In this analysis, therisk was calculated relative to the baseline risk in a nonsmoking reference population. Finally, the authors fitted
the parameters of the BEIR IV model with the exposure–
response model using the data for white Colorado Plateau
uranium miners. The parameter estimates for time since
exposure and attained age were comparable to factors presented by BEIR IV; however, the basic risk factors (b2in BEIR IV terminology) estimated in the regressions ofColorado ­Plateau data alone (not correcting for exposure in
OHR mining) were an order of magnitude lower than thoseestimated for the combined cohorts by BEIR IV [C19].

Summary of regression analyses by smoking category [S9]Category

Number ofminersNumber of lungcancers

All miners

3 359

Non-smokers
Ex-smokers

Excess absolute risk

Excess relative risk

Slope
(PY/WLM)aIntercept(PY–1)a

F-ratiob

Slope(WLM–1)

Intercept

F-ratiob305

5.0 × 10–6

–0.18 × 10–3

3 960

2.3 × 10–3

0.30

619

493

14

3.1 × 10–6

–0.61 × 10–3

139

21.1 × 10–3

–3.8

203

318

38

7.5 × 10–6

1.1 × 10–3

45

9.2 × 10–3

1.2

55

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

Category

Number ofminersNumber of lungcancers

All smokers

2 754Light smokers
Heavy smokers

Excess absolute risk

237

Excess relative risk

Slope
(PY/WLM)aIntercept(PY–1)a

F-ratiob

Slope(WLM–1)

Intercept

F-ratiob282

5.7 × 10–6

–0.30 × 10–3

1 954

2.2 × 10–3

0.30

607

501

29

5.2 × 10–0.35 × 10

63

3.5 × 10

–3

1.3

11

1 910

215

5.5 × 10–6

–0.0035 × 10–3

7 882

1.8 × 10–3

0.29

264

–6

–3

a PY = person-year.
b	A statistical test to see whether the amount of variation explained by the regression is significant.

Table 10 Summary of regression analyses by smoking category for miners with no OHR mining exposures [S9]Category

Number ofminersNumber of lungcancers

All miners with
no OHR mining

1 926

Never smoked

Excess absolute risk

Excess relative risk

Slope
(PY/WLM)aIntercept(PY–1)a

F-ratiob

Slope(WLM–1)

Intercept

F-ratiob

152

5.5 × 10–6

–0.92 × 10–3

599

2.6 × 10–3

–0.031

1 266

338

9

2.8 × 10–6

–0.64 × 10–3

344

20.7 × 10–3

–4.1

438

Ex-smokers

177

22

8.6 × 10–6

–1.7 × 10–3

56

9.7 × 10–3

3.4

27

Light smokers

318

15

5.1 × 10–3–0.80 × 10

108

3.3 × 10

1.1

11

Heavy smokers

1 015

99

5.9 × 10–6

–0.84 × 10–3

977

2.4 × 10–3

–0.15

498

All smokers

1 528

138

6.6 × 10–6

–1.3 × 10–3

310

2.6 × 10–3

–0.033

1 034

–6

–3

a PY = person-year.
b	A statistical test to see whether the amount of variation explained by the regression is significant.

174. Stram et al. [S60] reported an analysis which used ameasurement error correction of lung cancer risk based on
fitting a multilevel statistical model to the Colorado Plateau
uranium miner cohort data within the same mine, locality
and mining district. The authors used two subcohorts fromthe cohort of 3,347 white miners employed for at least one
year in the period 1950–1960 as defined by Roscoe [R8].
The first cohort (referred to as the 1950 cohort) included
2,074 miners with 263 lung cancer deaths who had theirinitial mining experience commencing in 1950 or later. The
second cohort (referred to as the 1952 cohort) included
2,388 miners with 209 lung cancer deaths. The authors
noted that the reason for the selection of the 1952 cohort
was that systematic measurement of radon in mines did notstart until 1952. The approach to (exposure) error correction was based on a computation of the exposures (WLM)
for each year and mine of interest. The authors investigateda number of models for lung cancer mortality, including asimple linear ERR model of the form 1 + βχ(t), where β
is the ERR per 100 WLM and χ(t) is the miner’s cumulative workplace radon exposure up to 2 years prior to theattained age. In addition, the authors investigated several
models, including a simplified BEIR VI model, of the effectof smoking, attained age and time since exposure. For

the simple linear risk model, the authors report β = 0.28
(SE = 0.075), and for the exposure error adjusted model,
β = 0.44 (SE = 0.14), both for the 1950 cohort. Similarly, forthe 1952 cohort, the authors report β = 0.33 (SE = 0.1) using
the uncorrected exposure and β = 0.54 (SE = 0.2) for the
exposure error adjusted model. For both cohorts, the error
correction increased the risk estimates by about 60%. The
authors found a submultiplicative relationship between radon
exposure and smoking, both with and without error correction. The authors also observed an exposure-rate effect; this,however, diminished after correction for exposure measurement error. With their simplified BEIR VI model, the effectof low exposure rate (0–15 WL) was essentially the samewith and without measurement error correction.
175. Gilliland et al. [G4] reported a study of the exposure to
RDPs and lung cancer risk in non-smoking uranium ­miners.
The authors used case–control methodology and conditional
logistic regression analysis to investigate the relative risk ofdeath as a function of cumulative exposure to RDPs. Theirfindings are in close agreement with a parallel analysis ofminers reported by Lubin et al. [L10]. The authors concluded that non-smoking miners were indeed at increased
risk of developing lung cancer.

238

UNSCEAR 2006 REPORT: VOLUME II

176. Luebeck et al. [L16] noted that with their biologically
based model and parameters applied to the Colorado Plateau
miner cohort, an inverse dose-rate effect was not seen with levels of exposure typical in residences. The ERR was estimatedto be about 0.0078 (95% CI: 0.0036, 0.0165) per WLM for
25 years of residential radon exposure at a level of 150 Bq/m3.
This is consistent with the value reported from a ratio analysis
of eight epidemiological analyses of residential radon exposure[L4]. Finally, Luebeck et al. [L16] noted that the comparable
risks in the BEIR VI report, expressed as lifetime risk, werehigher than their own estimates by a factor of 2–4.
177. The Navajo of the south-west United States wereinvolved in the mining and milling of uranium ores in the
Colorado Plateau area from the 1940s to the 1970s. The 1995
study of Roscoe et al. [R9] updated an earlier study [A4] ofmortality among Navajo uranium miners. The 1995 study
reported on a cohort of 757 Navajo miners with vital status
followed from 1960 to 1990 and a mean cumulative exposure of 755 WLM accumulated over an average of 8.3 yearsof underground work. The exposures were based on the workof Lundin et al. [L13]. A life table approach based on mortality data for non-white men in New Mexico and Arizona,direct standardization of rates and internal comparisons
between the exposure categories were used to analyse thecohort. All exposures were lagged five years, to represent a
reasonable minimum period for the induction of lung cancer.
A Cox regression analysis was used to account for simultaneous risk factors and the use of external mortality rates. The
time-dependent regressors considered in the model included
cumulative exposure, log cumulative exposure, duration of
exposure, log exposure rate, log cumulative pack-months ofsmoking, time since first exposure and others.
178. Standardized mortality ratios for a number of causesof death (heart, circulatory and digestive diseases) were lowered. Of the diseases examined, only the values for lung cancer, pneumoconiosis and “other respiratory diseases” were
elevated. The mean exposure to RDPs among the 34 deaths
observed from lung cancer (versus 10.9 expected) was
1,517 WLM. Smoking status for the miners with lung cancer
was similar to that for the entire cohort of Navajo miners. A
log-linear model and a linear model in cumulative exposure
fitted the data equally well, and yielded ERRs of 13.8 and 9,
respectively, for a cumulative exposure of 400 WLM relative
to no exposure. Unlike the case of the white miners, smoking
was not strongly associated with lung cancer risk in the Navajo cohort. The authors attributed the excess non-malignant
respiratory disease (standard mortality ratio of 1.4) to be due
mainly to pneumoconiosis and exposure to silica and otherworkplace contaminants rather than to radon.
179. Latency is an important consideration in ­evaluating
potential lung cancer risk following exposure to RDPs (e.g.
[C20, L10, N7, N11]). Langholz et al. [L39] reported on
an investigation of methods to assess latency effects andan analysis of latency in Colorado Plateau uranium miners
using a nested case–control methodology with 263 lung cancer deaths. Of these, 239 cases were matched to 40 controls

each and the remaining 24 cases with fewer than 40 controls (all controls were used). The relative risk of lung cancer
increased for about 8½ years and then decreased, reachingbackground levels after about 34 years. The decline in risk
with increasing time since exposure was much more pronounced in persons over 60 years of age. Hauptmann et al.
[H30] reported on the use of splines (piecewise polynomial functions) to analyse latency in the Colorado uranium
miner cohort and reported similar results, with ERR > 0 for
the period from 9 to 32 years prior to the identification oflung cancer. The ERR reached a maximum of about 0.6 for
100 WLM about 14 years after exposure and decreased toabout 0.02 thereafter [H30].
6.

Evaluation

180. Estimated exposure rates for individual miners in the
Colorado Plateau area have large uncertainties due to: variations between workplaces within mines (even for those mines
where RDP concentrations were measured); the necessity touse “guesstimates”, extrapolations or estimates from othermines in the area; uncertainties in OHR mining exposures
before (and possibly after) employment at uranium mining
facilities; and discrepancies in the work histories of those in
the sample group examined. There are significant discrepancies in the work histories of workers who worked underground part-time. Furthermore, the average WL values may
not apply to this type of worker, as they may have worked
in areas subject to lower ventilation rates than those for theother workers. Reconstruction of exposure histories fromcompany records is probably not feasible, because workers tended to work in numerous mines, for which few or norecords are available. However, it should be possible to study
a statistically valid sample of miners and from these investigations to draw conclusions concerning the uncertainty associated with the estimates of exposure for the cohort per se.
181. Overall, notwithstanding limitations in the exposure
data, the Colorado Plateau cohort of uranium miners is anextremely valuable resource for risk estimation. It provides
one of the most substantial bases for risk estimation for groups
exposed to RDPs, and the best information on smoking histories. The prominent strengths of this group include: the size ofthe cohort; the extent of follow-up; the considerable amountof exposure information for the periods of interest (prior tothe mid-1960s); information on smoking; and the possibility
to assess the effect of OHR mining. Although no systematic
bias was identified in the estimates of the exposures for thiscohort, the uncertainties in the exposures of individual miners
are very large, particularly for the early years of mining.
B.

Canada: Ontario uranium miners
1.

Introduction

182. Uranium mining in Ontario started in the early to mid1950s in the Elliot Lake area. Uranium production developed

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

rapidly to reach a peak during 1957–1960, and declined justas rapidly after 1960. To illustrate this, there were 2 operating uranium mines in 1955, 15 in 1958 and 5 in 1964.
The period during which the largest exposures occurred for
Ontario uranium miners was relatively short (10 years orso). In total, more than 16,000 men were employed in the
Ontario uranium mines at various periods between 1955 and1977 [M8]. By 1988, only two uranium companies, both in
Elliot Lake, remained in operation, and by 1993, only one ofthese, Stanleigh, remained in operation. It ceased operationsin 1996.

239

methods of mining and ventilation were incorporated. Considerable information about exposure conditions in Ontario
mines is given in Ham [H1] and in Muller et al. [M19].
185. Ventilation flows changed in Ontario mines over thesame time period. McCrodan [M4] reported ventilation rates
for two of the Elliot Lake mines (table 11).
186. The Ontario investigators recognized that estimatesof RDP concentrations involved some uncertainty, especially for the early years of mining for which relatively few
reliable measurements were available. The epidemiologicalinvestigators worked together with mine ventilation engineers who were familiar with the Ontario uranium minesover the early years of operation to develop two estimatesof RDP concentrations, “standard WL” and “special WL”,
in an attempt to bracket the uncertainty [M19]. Muller andhis co-workers suggested that the standard WL estimate wasmore representative of a miner’s exposure, while the special
WL estimate was probably on the high side.

183. Muller and co-workers have carried out a number ofstudies (e.g. [K13, M3, M8, M15, M19, M21, M22, M23,
M24, M25, M26]) on some 50,000 men who worked in one
or more mines in Ontario. The Ontario miner population was
subdivided into gold miners, nickel/copper miners, iron ore
miners, uranium miners, other ore miners and a mixture ofminers. Men who worked at the Eldorado Mining Company’s
Port Hope plants or in Eldorado mines in Saskatchewan or
the Northwest Territories, as well as those with documented
exposure to asbestos, were treated separately. This left acohort of about 15,000 miners with exposure in Ontario
uranium mines. Men were considered to have entered thestudy at the time of their first medical examination (between
1 January 1955 and 1 January 1978) if they reported having
worked for a half-month or longer in dust exposure in an
Ontario uranium mine. Miners left the study on 31 December 1981, or at death if this occurred earlier. Miners with uranium mining exposures outside Ontario and asbestos miners
were excluded.

187. In 1985, a reconstruction of early underground uranium mining environments was undertaken [D7]. Mining
practices during the late 1950s and early 1960s in the Elliot
Lake area were reproduced in reconstructed undergroundmine areas. Extensive measurements were made during thesetests. RDP levels were recorded under a variety of operating
conditions. Under continuous ventilation with compressed air(the most favourable condition), exposures during the 1950s
were likely to have been not more than double the exposures
received by miners today. However, in unventilated areas, theexposure levels could be as much as 10 times higher than inmines today [D7]. While this work reduced the uncertainties
associated with the concentrations of RDPs, the uncertainties in ventilation practice and in the times spent in different
locations in the mine remained. The 1985 study concluded
that a typical raise miner in the 1950s, relative to miners ofaround 1985, would have been exposed to RDP levels threetimes higher, about equal levels of thoron decay products
and gamma radiation, and much higher levels of uranium
and quartz dust. Further information concerning the atmospheres of the Elliot Lake mines and the levels of radon andthoron decay products is given in reference [D8]. The measurements made during the reconstruction also confirmed thatthe exposure values used for the early years (specifically forraise miners) did not overestimate the working exposures atthat time.

2. Radon and radon decay products
184. A certain number of workplace exposure data for operating uranium mines in Ontario became available starting in
1955, although systematic measurements were not started
until 1958. According to Ham [H1], the Ontario Departmentof Mines issued codes in 1957 requiring that various measurements, including measurements of RDPs, be taken in themines. Thus, however infrequent, measurements were madein each operating mine from that time on. Records of RDP
levels for 1954–1955 show that average RDP concentrations
ranged from 3 to 7 WL and the average exposures of full-time
underground miners from 36 to 84 WLM/a. The levels varied(up and down) over time as new mines developed and new

Table 11 Ventilation rates at Elliot Lake mines (adapted from reference [M4])Year

DenisonQuirke I

cfmc

cfm/tond

cfmc

cfm/tond

1957–1958

200 000

33

300 000

54

1960

–

–

300 000

70a

1961

350 000

59

–

–

240

UNSCEAR 2006 REPORT: VOLUME II

YearDenisonQuirke I

cfmc

cfm/tond

cfmc

cfm/tond1965

450 000

75

–

85

1965

530 000

88

1965

530 000

150a

1968

–

250 000

125a

1969

–

1971

650 000

108

400 000

200b

1975

650 000

108

a Reduced tonnage.
b Mine reopened.
c Cubic feet (of ventilation air) per minute (1 cfm = 4.72 × 10–4 m3 s–1).

3.

Exposure estimation

188. According to Muller et al. [M19], some 131,000
measurements of RDP concentrations were made over theperiod 1955–1981, using the Kusnetz method. These data
were obtained over a period representing some 141 mineyears of operation, which corresponds to 929 measurements per mine-year of operation. Up to 1977, there wereapproximately 55,000 measurements representing some
126 mine-years of operation; this corresponds to approximately 430 measurements per mine-year of operation
over this period. Expressed another way, in the period
1955–1977, approximately 1.7 measurements were madeper average man-year worked, or, on average, one WL
measurement for every 9 WLM accumulated by the study
cohort.
189. About 23% of the assigned WLM exposure of thecohort was based on extrapolated values of RDP concentrations, particularly during the early years of mining. The
mean period of extrapolation was approximately 1.1 years.

Table 12d	Cubic feet (of ventilation air) per minute per short ton (907.19 kg) hoisted tosurface.

The key observations concerning exposure estimations for
the Ontario study taken from references [M23, M24] aresummarized below:
1955:

2 uranium mines; sporadic measurements;

1958:	15 uranium mines; systematic measurements,
assuming 80% of time spent in working
areas (headings, stopes, raises) and 20% in
travelways;
1955–1981:	131,000 RDP concentration measurements
over 141 mine-years of operation (an average of 929 per mine-year, though fewer in theearly years); mean extrapolation period (pre1958) of 1.1 years (representing 23% of thetotal collective exposure expressed in personWLM); “standard WL” and “special WL” bymine and calendar year, with “standard WL”
considered more representative.
Table 12 shows the distribution of collective exposure basedon extrapolated and measured WL values.

Exposure (WLM) of Ontario miners based on extrapolated WL values [M23, M24]
Calendar year

Extrapolated standard WLM

Total standard WLM

Per cent WLM based on extrapolation

1954

843

843

100

1955

4 276

4 276

100

1956

20 80620 806100

1957

52 26352 263

100

1958

9 736

87 04811

1959

10 61781 733131960

12 90255 27523

1961–19770

185 079

0

TOTAL

111 443487 323

23

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

190. It is not clear how the upper bounds for the exposures of uranium miners from 1955 to 1977 in reference[M19] were generated. Unfortunately, no justifications or
explanations were given; however, as noted earlier, theexposures used for the early years were not likely to havebeen overestimated [D7].
191. The Ontario mining industry experienced frequentchanges in market conditions. Mining companies and miners
moved from mining one type of ore to mining another, andfrom one location to another [M19]. According to Muller
et al. [M15], the miners were all requested to fill in detailed
employment information concerning their first 60 monthsof mining experience so that those involved in mining otherores could be identified and excluded from the study.
192. For the period 1955–1977, full-time miners wereassumed to spend 80% of their working time undergroundand 20% on travelways. Part-time miners were assumed tospend 50% of their time underground in areas with higher
than average concentrations of RDPs. Up to the end of 1967,
and in one mine up to 1 April 1968, WLM exposure values
for each miner were estimated as WLM = WL × WHF ×
Months, where WL is the weighted average of WL measurements in stopes, raises and travelways; WHF is a work
history factor introduced to account for overtime or work
stoppages; and Months are the total number of actual months
worked underground. The Ontario investigators set WHF
equal to unity if normal working hours were maintained.
The WHF was increased or decreased as appropriate if overtime hours were worked or if work stoppages occurred ina particular calendar year. After 1967, the individual exposure estimates were based on time cards, filled out daily byminers, and on WL measurements made in the particular
work locations reported by the miner on his time card.
193. Smoking information on the Ontario uranium miners
is very incomplete. The most recent studies of Muller et al.
[M23, M24] reported the results of a 3% sample of men born
prior to 1939. Smoking information was sought on 226 uranium miners who had no former gold mining experience.
Smoking information was also sought on 80 lung cancer
deaths that occurred between 1955 and 1981 in the cohortof uranium miners with no former gold mining experience.
Of the 80 lung cancer deaths, smoking information was
obtained on 73 men; of these, 72 were smokers or former
smokers [M22].
194. Studies of lung cancer incidence and mortality in
Ontario gold miners reported by Kusiak et al. [K11, K12]
suggested that both radon and arsenic might be causative
factors in lung cancer. Kusiak et al. [K12] noted that available data indicated that dust concentrations in some gold
mining occupations in the 1930s and 1940s were often
above 1,000 particles/mL (p/mL), decreasing over time to an
average of 400 p/mL by 1959 and 200 p/mL by 1967. Geological data confirmed the “anomalously high arsenic levels
where gold is found...” and “that arsenic concentrations...areregionally enriched”. The authors noted, however, that no241

excess of lung cancer could be identified in gold miners whobegan mining gold after 1945 [K12].
195. According to reference [K12], no RDP concentration measurements were made in Ontario gold mines prior
to 1961. RDP levels in gold mines were variable. In somemines, average WL values in inactive areas were 0.3 WL,
while in other mines, levels were below 0.02 WL.
196. Increased ventilation rates and related practices introduced since the 1950s may have had less effect on ore dust
concentrations than on the concentration of RDPs in themine air. According to Ham, dust levels in Ontario uranium
mines decreased by only about a factor of 2 between 1960
and 1975 [H1], while the estimated concentrations of RDPs
decreased by about a factor of 5 over the same period [M21].
On the other hand, data reported by DSMA Atcon Ltd. [D7]and by Duport and Edwardson [D10] suggested that thelevels of ore dust and/or RDPs in the mine air diminished bycomparable amounts over the years in the Ontario mines.
197. In the Elliot Lake mines, thoron decay products also
contributed an appreciable radiological exposure; their­concentrations, however, were less affected by increased­ventilation than the concentrations of RDPs [J9]. Data for two
Elliot Lake mines reported by Chambers et al. [C2] showedthoron decay product levels ranging from 0.1 to 0.3 WL,
with parallel RDP levels ranging from 0.2 to 0.5 WL. Thissuggests a ratio of thoron decay product exposure to RDP
exposure of about 0.5.
198. A great deal of information is available concerning
the working environment in the Elliot Lake mines. Some
data were developed by the mining companies for engineering or regulatory purposes, while other data were developed
through the various research activities of the Elliot Lake
Mining Research Laboratory operated by Energy Mines
and Resources Canada, the Atomic Energy Control Board(AECB) of Canada and the mining companies themselves.
Data are available on various subjects, including: the effects
of using diesel equipment on the characteristics of mine air[B8, K2, K3]; the composition of the mineral dust (70%quantity) [B7, K4]; particle and activity size distributions[B6, B9, D10]; gamma radiation levels [C5, C6]; and arseniclevels in Ontario gold mines [O1]. A 1986 study [S11] examined the potential to use electrostatic precipitation to reduceradioactive aerosols in underground uranium mine atmospheres. This study further summarized available data on dust
loadings, particle size distribution and the attached fractionof the RDPs. Overall, detailed information is available tocharacterize the mine environment for the 1970s onward, butlittle information is available for earlier times.
4.

Epidemiological analyses

199. In the Ontario uranium miner cohort, the averageminer had, in 1984, 1½ years of mining experience and amedian age of 39 years. Overall, there was an average of242

UNSCEAR 2006 REPORT: VOLUME II

15.1 person-years at risk per man. Muller et al. [M26] reported
mean exposures for this group in the range 40–90 WLM.
200. The present cohort consists of men who worked ahalf-month or longer in an Ontario uranium mine between
1 January 1955 and 31 December 1977 [M19]. The exclusion of those with known asbestos exposure or with exposure in uranium mines other than in Ontario reduced thecohort size to 15,984 men. It was discovered that 66% of
these miners had OHR mining experience and that OHR
mining, particularly gold mining, increased the risk oflung cancer significantly [M23, M24, M26]. Exclusion ofminers with OHR mining experience reduced the cohort
size to 5,443 [M19]. The Ontario studies demonstrate theimportance of identifying OHR mining experience, sinceuranium miners with previous gold mining experience
exhibited excess risk of lung cancer even at zero exposure from uranium mining. In men who worked as gold
miners before becoming underground uranium miners,
92 deaths from lung cancer occurred, compared with the55.7 expected [M22].
201. In a follow-up analysis of mortality from lung cancer
in Ontario uranium miners, Kusiak et al. [K13] re-examined
the Ontario uranium miner cohort; mortality follow-up was
extended from 1981 to 1986. An association between excess
lung cancer and RDP exposure was found in the minercohort; this was similar to that found in the same cohort withfollow-up to 1981. The study found that lung cancer mortality in Ontario uranium miners who also mined gold was
related to exposure from both RDPs and arsenic.
202. In a cohort and case–control analysis of Ontario
miners, Finkelstein [F13] investigated the presence of silicosis as a risk factor for lung cancer. A cohort of 382 miners with silicosis and 970 controls were developed fromthe 68,000 workers in the Ontario database. Data wereavailable for 94% of the silicosis cases and for 99% of the­controls. In discussing his cohort analysis, Finkelstein notedthat there was a significant excess of cancer (mainly lungcancer) in miners with silicosis and that men with normal
radiographs had a lower cancer incidence than the Ontario
average. Finkelstein used a case–control methodology andlogistic regression to assess the risk of cofactors. He foundthat silicosis is a highly significant risk factor for lung cancer. In an analysis of cumulative risk, Finkelstein calculated
a weak association between silicosis and lung cancer with
OR = 1.004 (95% CI: 0.9967, 1.011), while in his model ofthe joint effect of cumulative radon exposure and silicosis, hefound no association of lung cancer with radon (OR = 0.995;
95% CI: 0.986, 1.004) and a strong association with silicosis (OR = 6.99; 95% CI: 1.91, 25). BEIR VI [C20] and the
UNSCEAR 2000 Report [U2] reviewed the available data on
the effects of exposure to silica in underground miners andcommented on various studies, including that of Finkelstein[F13]. Overall, there is a range of opinions on the effects
of exposure to silica in the mining environment, and someuncertainty remains concerning the influence of silicosis on
risk of lung cancer.

203. One of the most important observations from theOntario miner study was the use of a “time since exposure”
effect, where the risk of lung cancer decreased with increasingtime since exposure [M22]. NCRP Report No. 78 [N7] previously reported a dose–response model where the excess absolute risk of lung cancer was assumed to be zero before 40 years
and then, following an initial latent period of five years, todecline exponentially with time since exposure. The concept
of declining risk with increasing time since exposure was subsequently incorporated in the analyses of others, including
Lubin et al. [L10], BEIR IV [C19] and BEIR VI [C20].
5.

Evaluation

204. The Ontario miner study is a large well-defined study
cohort with considerable information available on which
to base exposure estimates. Researchers have attempted toassess the effect of uncertainty in exposure through the useof standard and special working levels, which are mainly
estimates. While more could be done in quantitatively evaluating the effect on dose response of uncertainty in exposure
estimates, the cohort provides one of the highest-quality
studies available for radon risk estimation. Future updates
will further increase the value of this cohort.
C.

Czechoslovak miners
1.

Introduction

205. Mining in Jachymov (Joachimsthal), Bohemia (nowin the Czech Republic), started at the beginning of the 16th
century. Thousands of silver miners were involved. Mining
for cobalt, bismuth and arsenic started later in the 16th century. In the middle of the 19th century, uranium was mined
for use in the glass and porcelain industry. Between 1909
and 1925, Jachymov was devoted to pitchblende mining inpursuit of 226Ra, with an average annual production of about
26 g of radium [L6].
206. Mining conditions in the early years were poor. The
mines were usually damp (especially in the spring when thesnow was melting) and cold. Miners had to descend ladders
hundreds of metres to their working areas. Natural ventilation was provided by a gallery between all the mines, and,according to Lorenz [L6], it was generally “sufficient”,
except in dead-end shafts.
207. Although high death rates among miners in theirprime years caused by lung-related diseases were recognized
and recorded as early as the 1550s, no detailed studies of thiswere performed. In 1879, Harting and Hesse (as reported
in reference [L6]) became the first to conduct organizedinvestigations on workers at the Schneeberg mines, which
are across the border from Bohemia, in Saxony (Germany).
They found that 75% of the deaths were caused by malignant growths in the lung, and that the incidence was greater
among miners than among masons or carpenters working

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

in the mines [L6]. From 1869 to 1877, 150 deaths due to­“miner’s disease” were identified in a workforce of 650 men.
The onset of the disease occurred after about 20 years ofwork in the mines. According to Lorenz [L6], Harting andHesse were the first to diagnose miner’s disease as lung cancer but, owing to the high levels of dust in the mines, they
assumed that the inhaled arsenic and poor nutrition were thepredisposing factors for the disease.
208. A more recent paper by Greenberg and Selikoff [G7]reappraises the data reported by Harting and Hesse in 1879,
and of the early efforts to identify the cause of lung disease
among the miners and to measure how much dust the miners
inhaled. One experiment (as reported in reference [G7]) concluded, “a miner inhaled 0.231 g [sic] of dust in a 7 h shift”.
Unfortunately, it is not clear whether the studies of Hartingand Hesse led to any improvement in working conditions.
The identification of RDPs as the primary causal agent wasleft to later investigators.
209. More investigations followed. Rostoski and Saupe, ina 1921 study of selected miners and non-miners (507 people
in all), identified lung cancer as a common cause of deathamong miners [L6].
210. Pirchan and Sikl [P5] observed the first case of lungcancer in Czechoslovak radium workers in 1926. In 1929,
Lowy [L48] reported two deaths of Jachymov miners fromlung cancer. Further study of the issue was commissioned by
the Czechoslovak Ministry of Public Works. In 1929–1930,­Pirchan and Sikl [P5] examined necropsies in 13 of the 19 miners who died during that period. They stated that lung cancer
incidence was highly prevalent in Jachymov ­miners and suggested that “radium emanation” (radon) was the most probablecause. Lorenz [L6], however, later suggested that genetic susceptibility to lung cancer in miners must be unusually high.
211. Once radioactivity was recognized as a major causeof lung cancer deaths among miners, measures were taken
to reduce exposures. Drilling with water wash-out was­introduced around 1930 to reduce the amount of airborne
radioactive dust [S24].
212. Prior to 1932, the mines were ventilated using naturalventilation alone. Beginning in 1932, this was in some cases
reinforced by mechanical ventilation. After the ­Second World
War, mining started again. As early as 1946, the Czechoslovak Ministry of Health started collecting miner mortality
data to better understand the hazards of exposure to RDPs.
Up until 1952, ventilation continued to be mainly by natural
means; however, in the 1950s, artificial ventilation, both general and local, was systematically introduced into all mines.
From 1955 onward, all mines were mechanically ventilated.
From 1956 onward, auxiliary ventilation was also provided
in selected areas where appropriate. In 1954, the Czechoslovak Hygienic Service started undertaking inspections ofworking conditions in uranium mines. Increased efforts to
reduce radioactive contaminants by increased ventilationand tighter safety regulations ­commenced in 1966 [V1].

243213. Working conditions were improved as the knowledgeof the risks of lung cancer in miners from exposures to shortlived RDPs increased. According to Vesely and Sada [V1],a dramatic reduction in radon concentrations took place inCzechoslovak mines during the 1950s. Mean cumulative
exposures of about 310 WLM were reported for a cohort ofminers who started work in 1948–1952. The mean cumulative exposure of miners starting work in the years 1968–1972
was 40 times lower [S24].
2.

Radon and radon decay products

214. Pirchan and Sikl [P5] reported three active pits in
Jachymov prior to 1930. The quantity of radium emanation (radon) found in the air discharged from the Svornost­(Harmony) pit (depth 500 m) was 4 Mache units, from theWerner pit (depth 476 m) 15 Mache units and from the Saxon
Nobility pit (depth 120 m) 10 Mache units. (One Mache unit
is approximately 10 Bq/L, see Behounek [B3]). Pirchan andSikl realized that radioactivity levels varied throughout themine and suggested taking measurements at various locations to assess the exposures to miners [P5]. Lorenz [L6]
made a similar observation.
215. Water is a significant source of radon in underground mines. The Czechoslovak mines were no exception.
Behounek [B3] reported radon in groundwater at levels of up
to 426,000 pCi/L in the Jachymov mining district.
216. Systematic radioactivity measurements were not performed in Jachymov until the late 1940s. From 1949 to 1960,radon concentration measurements were recorded. Hundredsof readings for each uranium mine for each year during thisperiod are available [S18]. The “classical” method was used
to measure radon, i.e. measuring the current with an electrometer in an ionization chamber [S20]. The measurementof RDPs did not begin in Jachymov until 1960 [S20]. In
1968, personal exposure records, which took working place
and work time into consideration, were established. The
records were reported quarterly for each individual [S29].
217. Data on attached/unattached fractions are sparse. Some
data were collected from 1963 to 1965 by the Czechoslovak
Academy of Sciences. These measurements indicated that,typically at that time, the fraction of free unattached radium
A (historical name for 218Po) atoms was approximately 10%
in the Czechoslovak mines. The AMAD (activity medianaerodynamic diameter) was estimated to be between 0.05
and 0.2 µm, with the mean at about 0.1 µm [H2].
218. In 1968 and again in 1973, there were “ventilation
incidents” in underground uranium mines in Czechoslovakia [H2]. In both incidents, mechanical ventilation
stopped for a period of time. During these periods, radon/
radon decay product levels were measured to assess theequilibrium conditions that might have existed in the earlydays of mining, prior to the introduction of mechanicalventilation. On the basis of data collected during these244

UNSCEAR 2006 REPORT: VOLUME II

two incidents, an overall equilibrium factor of approximately 86% for the period 1948–1952 was estimated. It
was noted, however, that in some of the newer mines, thesituation was better. From 1953 to 1959, a nominal equilibrium factor of approximately 55% was assumed. For
the period 1960–1966, a radon equilibrium factor of 36%
was assumed; some of the older mines had higher equilibrium factors and some of the newer mines had lower
equilibrium factors.

219. During the period 1948–1952, when the uranium
mines in western Bohemia were naturally ventilated, 40%of the measurements indicated radon levels in excess of
1,000 pCi/L. The early high radon levels and the improvements that took place are illustrated in ­figure V. The breakin the curve shown in figure V indicates the change fromroutine radon measurements to routine RDP measurements
as the primary basis for the estimation of exposure.

Figure V. Evolution of the average radon concentrations, 1949–1968 (right ordinate axis), and the average potential (alpha)energy concentration, 1969–1981 (left ordinate axis) [H2, K7].

35
50

30

40

25
20

30

15

Radon
20

10
10

Potential α energy

5
0
1945

1950

1955

1960

1965

1970

1975

1980

ACTIVITY OF RADON PER UNIT VOLUME (kBq/m�)

CONCENTRATION OF POTENTIAL ENERGY (10� MeV/L)

60

0
1990

1985

YEARS

220. The free unattached fractions of RDPs shown in table 13
were obtained from a series of measurements in uranium mines
in Pribram, Czechoslovakia, in 1988 and 1989 [H2].
221. Measurements made more recently in an eastern Slovak iron ore mine with high dust levels (poorly

Table 13

ventilated) showed low unattached fractions. In one
series of measurements, 1.6% ± 0.8%, and in a second,larger series, 2.3% ± 1.3%, of (equivalent) RDPs wereun­attached [H2]. According to investigators, Jachymov
mines were thought to have unattached fractions of
6–10% [H2].

Unattached fraction of RDPs in Czechoslovak mines (adapted from reference [H2])Range (%)

Mean (%)

Po

12.0–43.0

27.5

Pb

4.0–15.7

9.4

214

Bi

0.6–10.0

5.3

Overall

5.9–23.0

14.4

218

214

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

3. Exposure estimation
222. Sevc et al. carried out the main Czechoslovak epidemiological studies on uranium miners. The first study wasin 1971 [S17] and there have been several updates [K8, K9,
S18, S19, S20, S25], the most recent being in 2004 [T40].
The study involved miners who started uranium ore ­mining
between 1948 and 1957. Sevc et al. [S20] reported that­“estimates of working levels of RDPs (WL) or radon gas
concentrations were made on the basis of records of the ventilation conditions and practices, emanation rates from different types of ores, and RDPs measurements made in 1960
and later” and also “the values for working level months(WLM) were estimated on the basis of radon gas measurements and from data on the number of months of employment with each mine and within each calendar year of thewhole employment period for each miner.” No further details
were given of how each factor was accounted for. However,
the authors estimated the coefficients of variation for WL
estimates to be <27%, and <30% for WLM estimates.
223. The average annual RDP concentration (WL) in each
shaft was used to calculate each man’s exposure on the basisof his working time. Payroll cards were available for all men
in the study groups. Beginning in 1968, individual personal
dosimetry cards recorded each miner’s exposure. Discussions with the Czechoslovak investigators indicate that the
exposures estimated using time-weighted area measurements are unlikely to differ by more than a factor of 2 fromthose obtained from personal dosimetry [H2].
224. Radon measurements were obtained for four types
of working area: mine workplaces in close proximity to theore (stopes), where the levels were the second highest in themine; hallways and corridors where there was no ore; chimneys (raises), where the levels were the highest; and transport
ways. Where no data existed on a working area, the averagevalue for the entire mine in a given year was used. Typically, 20% of the total number of measurements was madein transport ways and 80% of the measurements were madein the other working areas.
225. Individual miner exposures were calculated on thebasis of job descriptions recorded in personnel cards for
all miners (the use of which was begun in 1948 for payroll
purposes). The estimates of time spent in the workplace arethought to be reasonable, since the miners were under surveillance by a controller and a mine technician, the latterbeing responsible for rating the workers’ pay.
226. Very few measurements of the concentrations ofradon in the workplace are available for the years before
1948. Estimation of the exposures of miners before thisdate would require consideration of many factors, including the radon levels recorded in later years and knowledgeof early mining practices and ventilation systems. However, according to Hamilton et al. [H2], Group S consists ofCzech miners who began mining in the period 1948–1959
and worked for four years or more; it does not include245

miners with pre-1948 exposures, and therefore this aspect
is not important.
227. The employment history of the miners is another very
important factor in estimating exposures. Jachymov is situated in the Erzgebirge (Ore Mountains), where abundant
minerals are found. As discussed earlier, there was a long
history of mining prior to 1909, when pitchblende was firstmined. The early miners probably had previous mining experience (e.g. [L6, P2]). On the other hand, tin mines in the
Erzgebirge were not operated between 1931 and 1939, andduring the Second World War, the mines employed mainly
prisoners [T40]. These facts limit the possibility of previous mining experience among cohort members, and this isconsistent with Sevc’s claim that less than 2% of the epidemiological study group mined non-uranium ores before they
mined uranium [S24].
228. By definition, a “working month” conventionally now
means 170 hours of work. If a miner held more than one job
with a mine in the 1920s–1950s, the working hours spent ineach would need to be taken into account in order to determine his exposure. Sevc et al. [S20] stated, however, that
only the number of months of employment during a year
was taken into consideration. There was no evidence thatthe miners’ actual working patterns were incorporated in the
WLM estimations. In discussions held in 1988, the Czechoslovak investigators commented that the workers spent about
80% of their time in the workplace and 20% in the transport ways. In the early days, the men worked 8 hour shifts(3 shifts per day), 6 days per week. After 1968, the normal
working week was reduced to 5 days. Uranium miners had atotal of 5 weeks of leave per year. By 1990, retirement fromuranium mining was mandatory at age 50, but some retired
uranium miners often continued to work in other (especiallycoal) mines. After 1966, people over 40 years of age were notaccepted into uranium mining as new miners. Later reports[T37, T38] on the Group S cohort gave more details on how
information on jobs was used.
229. On the basis of more recent (preliminary) surveys[T29], Czech investigators estimate the exposure of the general population to RDPs in typical areas of the Czech Republic to be about 0.34 WLM/a, whereas some residents of the
Jachymov area could be exposed to 3–4 WLM/a. However,preliminary evaluations suggest that correction for at-home
exposure, which would have shifted miners to higher exposure categories, did not affect the risk estimates significantly.
The effect of exposure away from work is subject to ongoinginvestigation.
4. Epidemiological analyses
230. Czech investigators have studied several groups ofunderground uranium miners. The most studied group is
the Group S cohort. These miners represent approximately
11% of the underground miners employed in the Jachymov
and Horni Slavkov mines. The cohort originally included

246

UNSCEAR 2006 REPORT: VOLUME II

4,364 men [S17]. Reported differences in the size of thiscohort are due to whether or not emigrated miners were
included. The most recent results for the Group S cohort
included 4,320 miners [C20, T37].
231. In discussing the accumulation of WLM exposure,
Sevc et al. [S20] indicate that person-years at risk wereassigned totally to the final WLM category reached by eachindividual miner, rather than being distributed across each
WLM interval as they accumulated. This affected the estimation of the expected number of lung cancer cases in eachexposure category. This difference was discussed by Kunz
et al. [K8], who re-evaluated the epidemiological data ofreference [S20] and concluded that the distortion caused by­earlier methods was not large. Since 1978, all analyses [K8,K9, S25] have used the conventional approach.
232. In earlier papers (1971–1988), the observed numbersof deaths from lung cancer were compared with the numbers
expected from general mortality data. This was justified by aninvestigation of a random group of 700 miners that showedthat 70% were smokers, similar to the general male populationof Czechoslovakia. However, in publications since 1991, theexpected numbers were modified by incorporating a multi­
plicative parameter that allowed the background mortality toTable 14differ from that of the general population. This approach is
close to the “internal approach” if additional stratification forage and calendar year is used [T2, T26].
233. In the past, papers on epidemiological studies ofCzech miners did not usually indicate the numbers of miners involved, and therefore it has sometimes been necessary
to back-calculate from reported data to obtain the number ofminers in the various groups. However, in the first paper [S17]and since 1988 [S25], the numbers of miners were reported.
234. In 1988 [K7, S25], investigations of other study
groups were reported (table 14). These include Group S,
with subgroups A and B covering the underground uranium
miners whose exposure began during 1948–1952 and 1953–
1957, respectively. These two subgroups represent the maincohort for epidemiological investigations. In addition, investigations of a number of other study groups are also reported:
study Group N, which comprised uranium miners whostarted exposure at levels lower than those in study Group S;
a small study group, Group K, which comprised miners iniron mines in eastern Slovakia; and study Group L, which
comprised miners from the Czech shale clay mines. Studies
of non-uranium miners were completed with a Czech studyof tin miners [T40].

Czech and Slovak studies of miners exposed to radon [S25, T40]

Study group

Type of mine

Location

Cohort size

Exposure (WLM)

Latest reference

S* (= A + B)

Uranium

Western Bohemia

4 320

152

[T38]

N* (= C + D)

Uranium

Central Bohemia

5 622

7

[T38]

K

Iron

Eastern Slovakia

1 056

40

[S25]

L

Shale clay

Central Bohemia

916

25

[S25]

C

Tin

Northern Bohemia

2 466

54

[T40]

235. A more recent study by Tomasek and Placek [T2]
investigated risks to a subgroup of miners whose exposures
were restricted to lower exposure rates. This subcohort had
a total of 419 lung cancers to the end of 1995. A decrease inrelative risk with time since exposure and age at exposure
was observed. Differences in the risk estimates for epidermoid and small cell cancers were also identified, althougheach had a pattern of risk similar to that of lung cancer
overall. The authors found no evidence for non-linearity
or dependence on exposure rate (at RDP concentrations of
below 8 WL), although the average dose-rate effect was seenin the Group S cohort as a whole.
236. A subsequent study by Tomasek [T26] investigatedlung cancer risk in a cohort of 5,002 miners exposed in
two different periods. Exposures of the 2,552 miners in
the older cohort (S) were derived from workplace radonmeasurements commencing in 1949. For the 2,450 miners
in the newer group of miners (N), exposures were basedon individual dosimetric records. For the newer subcohort(N), smoking data were available for most (about 85%) of
the miners. For the older subcohort (S), smoking data wereavailable retrospectively for 279 cases and 410 (nested-in)
controls. Follow-up was to the end of 1999. The analysis
was based on a relative risk model that allowed consideration of time since exposure and attained age or age at exposure. Excess relative risk (ERR) was linearly dependent
on cumulative exposures received more than 5 years previously. The ERR was 0.045 (90% CI: 0.017, 0.140) per
WLM among non-smokers (42 cases) and 0.02 (90% CI:
0.011, 0.035) per WLM among smokers (309 cases); thedifferences between the two estimates were not statisticallysignificant. The lung cancer risk in miners who smoked wasabout 10.8 times that in non-smoking miners (this included
those who had not smoked for the previous 20 years). The
ERR was found to decrease by more than 60% per decade oftime since exposure and simultaneously by more than 40%
per decade of age at exposure.

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

5.

Evaluation

237. In comparison with studies reported in BEIR VI, theS-cohort had (by 1990) the second largest number of lungcancer deaths. Exposure information in the S-cohort is among
the most extensive ([C20], p. 322). Extensive measurementsof radon in all shafts are available almost from the first yearsof exposure. In results published in 2003 [T30], only 4% of
all exposure years are not based on radon measurements. The
most recent results on the Czech uranium miners [T39, T40]
are based on a total of 929 lung cancer cases. This combined
cohort includes a large proportion of miners with exposures
based on detailed personal dosimetric data and provides anopportunity to investigate the role of potential modifiers ofeffect.

D.

247

the survey in a relatively short time. I agree that theerror may be great but it should not be more than ±50%,
which is acceptable compared to other potential sourcesof error, even with a very sophisticated method, as local
variations, seasonal variations etc. But we have tried tomake all necessary corrections for these errors too as
far as we know them. The result of that survey is seenin the table.
Radon decay
products

Number ofmines

Number ofworkers<0.1 WL25

1 121

0.1–0.3 WL

13

1 740

0.3–1 WL

17

1 739

1–3 WL

5

133

Swedish iron ore miners
1.

Introduction

238. Iron ore has been mined in Sweden since medieval times. Originally, there was only open-pit mining, butaround 1910 underground mining started [E1, E2]. Before
1945, ventilation was entirely natural. According to Axelson [A11], some mechanical ventilation was developed during the 1940s and 1950s to prevent water freezing underground. Air was warmed by taking it through old shafts
before it reached the workplaces. Snihs [S36] commented
that in some mines, ventilation air was brought into the minethrough crushed rock; this method, although reducing airborne dust and raising the inlet air temperatures, also pickedup any radon emitted from the old shafts or crushed rock. In
addition, travel time resulted in the ingrowth of RDPs.
239. According to Snihs and Ehdwall [S36, S39], theprimary sources of radon in the Swedish iron mines wereincoming radon-rich water and, to a lesser extent, radonfrom radioactive minerals. The uranium content in the waste
rocks in the iron mines was of the order of 15–20 ppm [S39];
however, relatively high emanation coefficients (30–40%,
measured in accordance with the procedures set out in reference [A20]) for some of the rock were a contributing factor
to elevated radon levels in these mines. The radon problem
having been identified in about 1968, ventilation in the Swedish mines was gradually improved. The ventilation path waschanged to bypass crushed rock and incoming groundwater,
thereby leading to reductions in exposure [S37].

This was the situation in 1969 and 1970. My part of
the work is to make a “qualified estimate” of the radon
decay product exposure during the last 20 years, which
will be rather problematic I suppose.”
241. The Malmberget mine, as it now exists, is actually a
combination of several mines that initially were separate. The
iron ore deposit at Malmberget consists of about 20 distinct,
large ore bodies that outcropped to the surface; in these orebodies several open pits were started. Most of these mines
were in line with each other, separated by low-grade iron ore
formations.
242. Open-pit mining in the Malmberget mine area firstbegan in about 1890. Even later, when the depth required
for mining forced a change from open-pit mining to underground mining (by about 1930 all mines were underground),the bottoms of the mines were still above the general level ofthe surrounding terrain. The adits could therefore be driven
from the side of the mountain to the bottoms of the underground mines, which in turn were connected to the bottomsof the open pits, thereby permitting natural ventilation of
the mines. By 1955, the bottom levels of the mines reached
below the level of the country surrounding the mountain, andthe efficiency of natural ventilation declined as progressively
lower levels were developed.

2.

Radon and radon decay products

240. Snihs [S16] summarized the status of knowledge
about RDP levels in Swedish mines around 1972:

243. Early Swedish mine and miner data are summarizedin annex G of the UNSCEAR 1977 Report [U9].

“We measured only the radon concentration in many
mines to get a rough idea of the radiation problem inthe mines. To get the corresponding radon decay product concentration we then applied the factor 0.5. The
reason why we measured radon only, is that the mine
companies were asked to send samples (in pre-­evacuatedbottles) by post. By that method we were able to make244. According to Snihs [S36], the first radon measurements were made in the early 1950s in the Boliden mine.
However, limited knowledge about radon problems in
non-uranium mines and a lack of experience in taking­measurements delayed the institution of routine radon or
RDP measurements. A general awareness that many nonuranium mines had significant radon levels arose in about

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UNSCEAR 2006 REPORT: VOLUME II

1968. This led to radon surveys in which the general procedure was to take 3–12 radon gas samples per mine during
both winter and summer periods. Samples were collected in
evacuated bottles and sent to the National Institute for Radiation ­Protection (NIRP) in Stockholm for analysis.

the mining areas by these two methods; while the principal source of radon in the Malmberget mine was likely tohave been radon-rich mine water, it is likely that additionalradon entered the mine air because of this and the method ofventilation.

245. Each mine with RDP levels of 0.3 WL or greater
was investigated further by NIRP staff. They visited themine and took RDP samples using the Kusnetz method,
as well as many additional radon gas samples. Typically,
more measurements of radon gas concentrations than of
RDP concentrations were taken. According to reference[S28], the equilibrium between radon and RDPs was foundto vary greatly but was typically 50%.

250. Extensive recirculation of air was widely practised in
the Malmberget mine in the 1950s and 1960s. This recirculation could have permitted the buildup of both radon andthe equilibrium factor during that period. Changes in mining methods in the mid-1960s made it necessary to introduce diesel equipment, which in turn led to the requirement
to improve mechanical ventilation. Overall, Swent and­Chambers [S28] concluded that the pre-1969 WL values
given by Radford and Renard [R2] were likely to have been
underestimated by a factor of above 2, as they did not take
into account: the earlier practice of recirculating air in themine; the lower volumes of air circulated through the minein earlier decades; periods of stagnant ventilation airflow,which occurred during the years when natural ventilation
was the only ventilation method; and the pattern of decline
in the incidence of silicosis in later years, confirming theimprovement in ventilation.

246. To provide a basis for estimating the equilibrium­factor, simultaneous measurements of radon and RDP concentrations were taken. Typically, an average equilibrium
factor was applied to all working areas in a mine or partsof the mine [S37]. According to Snihs and Ehdwall [S39],
the equilibrium factor varied widely, ranging from 0.15 at
the air inlet of the mine to nearly 1 at the air outlet. Average
equilibrium factors in workplaces were typically between
0.4 and 0.6 [S39]. Axelson [A11], however, felt that a more
appropriate typical equilibrium factor was 0.7.
247. Swedish measurements of radon were typically madeby the NIRP using 4.8 L conventional propane containers. The containers were evacuated by the NIRP and subsequently opened at places of interest in the mines. After
sampling, they were sealed and mailed back to the NIRP for
analysis in ionization chambers [S39].
248. Following the first measurements of radon, the mines
were divided into zones by radon or RDP level and subsequently checked according to the following frequency:
	Zone 1:	<10 pCi/L (<0.1 WL), once every
two years;
Zone 2:

10–30 pCi/L (0.1–0.3 WL), every year;

Zone 3:	30–100 pCi/L (0.3–1 WL), once every
six months.
For areas with levels of greater than 1 WL, measurements
were to be taken every three months, according to Snihs andEhdwall [S37].
249. In 1986, Radford and St. Clair Renard reviewed thehistory of mining methods and of general ventilation in themines (reported in reference [S28]). When the mines werefirst converted from open pits to underground mines, theunderground method of extraction was by shrinkage stoping.
Small pillars were left between the large shrinkage stopes.
After the shrinkage stopes were drawn empty of broken ore,the hanging wall was allowed to cave in and fill the opening.
Mining methods evolved from shrinkage stoping, followedby sublevel stoping, and in 1965 by sublevel caving, whichallowed even larger quantities of wall rock to cave in. Until
1973, fresh air was drawn through the broken rock left above

3.

Exposure estimation

251. Snihs [S36] developed his estimate of risk of lungcancer by assuming that RDP levels in Swedish mines measured since 1969 were representative of earlier years. Snihs[S36] made “qualified guesses” for exposures that may haveoccurred in relation to the observed mortality from lung cancer. Snihs and Ehdwall [S39] provided further discussion of
the measurements of radon and RDPs that were started only inthe late 1960s in Swedish non-uranium mines. These authors
noted that the earliest measurements of radon in mine air at
Malmberget were in 1968 but that subsequently extensive
measurements of radon and radon decay product concentrations in air were made by the NIRP and the Swedish mining company LKAB. Radford and Renard [R2], noting thatthe new ventilation system for mines became operational in1972, stated that the reconstruction of past concentrations
depended on the measurements made during the period
1968–1972 and on knowledge of the natural and mechanicalventilation used previously. However, the authors noted thatthe reconstruction of Malmberget exposure data depended
on the “assumption” that ventilation conditions in the minesin 1968–1972 were not greatly different from those in thepast [R2].
252. Key features of the exposure estimation for Swedishiron miners in the study of Radford and Renard [R2] were:− 1930: nominal start of the study;− 1969: first (NIRP) radon measurements;− Until 1973: fresh air drawn through broken rock(preheated);− 1955: mechanical ventilation introduced to replacenatural ventilation;

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES− Until aboutoccurred;

1965:

some

air

recirculation− 1965: diesel engines introduced into the mines;− All exposures “guesstimated”.
Radford and Renard [R2] assumed constant exposures prior
to 1968. This is unlikely because of the changing ventilation
practices (e.g. increased volume and elimination of recirculation) and the introduction of mechanical ventilation prior
to the first radon measurements.
253. As discussed above, the data available from measurements of radon and radon decay product concentrations
make it possible to estimate relatively well the exposure inmines after 1970; however, few if any measurement data
are available from the period before 1969, and early data
must be estimated from a combination of later exposure
data, reconstruction of measurements and considerationof ventilation practices, radon sources and other factors. A
paper by Bergdahl et al. [B45] provides a comprehensive
re-evaluation of radon exposures in the Kirunavaara andMalmberget iron ore mines. The re-evaluation suggests thatradon levels were higher than in the estimate by Radford
and Renard [R2], but not as high as suggested by Swentand Chambers [S28].
254. In the epidemiological study of miners in Malmberget by Radford and Renard, undertaken in 1984, the historical exposure was estimated from data from 1968–1972[R2]. Also in this study the equilibrium factor Feq wasassumed to be 0.7. The same exposure estimate was used
for all underground workers except those that worked in
Koskullskulle (these workers seldom changed their working place; Koskullskulle was a separate mine that was not apart of LKAB until 1953). The radon exposure (expressed as
WLM/a) was estimated in reference [R2] for Malmberget as
follows: 1970–1972: 3.2; 1960s: 4.9; 1950s: 6.2; 1940s: 6.9;
1930s: 4.6; 1920s: 4.8; and 1910 and earlier: 4. For Koskullskulle, the authors [R2] estimated the exposure for the period
1920–1969 as approximately 2 WLM/a and for 1970–1972
as approximately 1 WLM/a.
255. Bergdahl et al. [B45] estimated the exposure during
the period 1925–1972 to have been constant at between 0.8
and 17 WLM/a, depending on the ore body in which thework took place. In cases where the location of the work was
unknown, the exposure was estimated to have been 6 WLM/a.
In Koskullskulle, the exposure in the period 1925–1972 wasestimated to have been 5 WLM/a, with values before 1925
gradually increasing from the 1910 value of 0.8 WLM/a.
256. The principal differences between the exposure matrix
proposed by Bergdahl et al. [B45] and the earlier exposure
estimates of Radford and Renard [R2] are that Bergdahl et al.
used a lower exposure value for 1970–1972, a higher value
primarily during the 1940s and 1950s, and a linear extrapolation back to 1890. Furthermore, Bergdahl et al. [B45] used
substantially higher estimates of exposure for Koskullskulle
than did Radford and Renard [R2].

249

4. Epidemiological studies
257. A number of authors reported epidemiological studies
of Swedish miners, including references [A6, A7, A8, A9,
A11, A13, C7, D1, D2, E1, E2, J11, J12, J13, L2, R1, R2,
S21, S22, S36, S38, S39]. However, the basic reference forthe study of Malmberget iron miners is the paper by Radford
and Renard [R2], which included a description of the basis forthe estimates of exposure to RDPs. This was a ­retrospective
study of lung cancer mortality in a group of 1,415 Swedish iron miners. The total cohort represented 24,083 personyears at risk, with an average exposure of approximately
81.4 WLM. The study cohort included men born between
1880 and 1919, who were alive on 1 January 1970, and who
had worked for more than a calendar year between 1897 and1978. Follow-up of these miners was reported for the period
1 January 1951 to 31 December 1976. The authors estimatedan ERR of 3.6% WLM–1, and an excess absolute risk of lungcancer of 19 per 106 person-years per working level month.
Miners were identified from company and union records ofactive and pensioned miners, which were available for theyears since 1900, as well as from medical records and, in a
few cases, from parish records. The Swedish Government
gives every person a code at birth; this code is included in all
work and hospital records. Every citizen is also required toregister in a local parish of the state church. These requirements helped in locating the miners and their records. Also
available from company records dating back to 1900 was
the total number of man-hours worked underground eachyear in each section of the mine. However, the exact location of the work within the mine was not available. There issome uncertainty as to work status for miners who started
or stopped work or who changed work function midway
through the year. Miners were assumed to work 173 hours
per month on average in the period 1890–1930; 162 hours
per month during the period 1930–1950; and 144 hours permonth from 1950 onward.
5.

Evaluation

258. There were few radon measurements in Swedish
mines and none in Malmberget mines prior to 1969. It was
only after this date that investigations were initiated by
the National Institute of Radiation Protection, Stockholm.
Therefore epidemiological studies of Swedish miners werebased on reconstruction of the radon and radon decay product concentrations to determine the exposure of miners.
Since underground work began in 1932, investigators undertook this reconstruction for a period of 36 years. During
this period, the conditions in the mines would have changed
owing to changes in the mining methods and ventilation
procedures. Two attempts were made to estimate the exposures [B45, R2]. The estimates obtained in reference [B45]for the Malmberget iron miners were higher than those usedin the epidemiological study of Radford and Renard [R2].
However, to date, no epidemiological re-evaluation of thisgroup of miners using the updated exposure data has been
published. These updated exposure data represent a great

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UNSCEAR 2006 REPORT: VOLUME II

improvement in the quality of the data available on the exposure of Swedish miners and will provide an opportunity to
update epidemiological analyses of Swedish iron miners.
E.

Canada: Beaverlodge, Saskatchewan, miners
1.

Introduction

259. The uranium mineralization at Beaverlodge in northern Saskatchewan, Canada, was discovered in 1946, and aprospecting mine shaft (the Ace shaft) was started in 1949.
By 1951, the company concerned, Eldorado, had identified
ore reserves sufficient to proceed with uranium production.
By that time, ore bodies to the west of the Ace shaft had also
been identified. In 1951, the Fay shaft was started as a production shaft serving the western ore bodies. An underground
haulage way also provided access to the Ace ore body. By the1970s, the mine was over 1.6 km deep and extended morethan 5.6 km horizontally. In addition, a number of satellite
mines, including two underground mines, were developed.
Mining production in the area increased rapidly during the1950s, but fell off in the 1960s as the demand for uranium
declined. Eldorado recruited experienced miners from thearea to work at Beaverlodge. The Beaverlodge operations
were closed in June 1982. As the Beaverlodge mine developed, all three shafts were connected underground, and twowinzes (shafts that do not come to the surface) were also
constructed.
260. Mill construction started in 1952, and the first ­uranium
concentrate was produced in early 1953. By this time the
Verna ore body to the east of the Ace shaft was identified;work started on the Verna shaft in 1953, with productioncommencing in 1956.
261. Twelve small satellite mines, most being small open
pits, were also developed during the 1960s and 1970s [G9].
Two underground mines, Hab and Dubyna, were also developed, both located several kilometres north and east of the
Verna shaft.
262. Bloy [B11] noted that the first mining method used
was shrinkage stoping. This was changed to cut and fill, using
waste rock and surface sand as the fill material. In turn, thismethod was replaced by hydraulic tailings fill. The initial
ventilation in working headings was mainly by compressed
air from the drills, with some surface air being supplied
by a 16 inch (~40 cm) metal vent pipe located in the shaft.
Because of the lack of knowledge of the hazards of radonand radon decay products, the standards used to control themine atmosphere conditions were the same as those used ingold mines at the time. This involved the measurement of airborne dust by the konimeter dust sampling method as well asradon measurements. An annual objective of the ventilation
programme was to maintain a ventilation rate of 10–25 cubicfeet per minute (cfm) of air per square foot2 of working face.
2

1 cfm per square foot = 5.24 × 10–3 m3 s–1 m–2.

263. In the cut and fill operation, ore was removed fromthe bottom, and mining advanced upward from one level tothe next. The sand fraction of the mill tailings was used as
backfill in the mine, with miners working off the backfill toremove the next lift. Simpson et al. [S30] investigated thisactivity as a possible source of radon and concluded that “nopositive evidence was found that backfill was a major sourceof radon in the mine”.
264. In addition to the Beaverlodge mine, Eldorado developed and operated a number of satellite mines. In addition,according to Garbutt [G9], there were a number of otheruranium mines operating in the Beaverlodge Lake area ofnorthern Saskatchewan. Eldorado’s policy was to recruit
experienced miners wherever possible. Consequently, as
other local uranium mines closed and Eldorado remained in
production, there were opportunities for miners with nonEldorado working experience to migrate to the Beaverlodge
operation.
265. The other (non-Eldorado) uranium mines in the
Beaver­lodge area operated for only a few years, and all had
been shut down by the late 1960s. Most of these mines were
small operations without mechanical ventilation (the exception being the Gunnar mine) and hence there was a potential
in these operations for high exposures even in a short time
period.
266. Estimation of individual employee exposures did
not begin at the Eldorado Beaverlodge uranium mine until
1967. While some measurements of the concentrations ofradon and radon decay products were made between 1954
and 1967, these were intended for monitoring ventilation
rather than for personal dosimetry. Eldorado began routinely
recording individual exposures in 1968, but considered thatsufficient workplace data were available to assess individual
exposures back to November 1966. The radon and radon
decay product measurements from the earlier periods weresummarized and utilized to provide exposure rate estimatesby occupational grouping and year [F7]. These were point
estimates, primarily based on the median concentrations
recorded during the year, and were subsequently used in theoriginal epidemiological analysis of Beaverlodge miners[H19].
267. The employee exposures were estimated by merging
the annual exposure rates with information from the nominal
roll. The nominal roll contained information on age, duration of employment and type of employment for Eldorado
employees. The resultant exposure estimates were the basis
for many epidemiological studies, including the original
cohort analysis [H19], a case–control study investigating theeffects of smoking and previous work experience [L17], theanalysis of four underground mining cohorts performed by
BEIR IV [C19], and the joint analysis of 11 undergroundminer cohorts in reference [L10] and BEIR VI [C20].
268. On the basis of these exposure estimates, the apparent lung cancer risk observed in the Eldorado Beaverlodge

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

cohort was substantially higher than the lung cancer riskobserved in the Eldorado Port Radium cohort [H18], wherethe ­estimated exposure rates were typically much higher
than those ostensibly observed for employees who worked
at the Beaverlodge mine. Given the substantially different
lung cancer risks observed between the two cohorts, andthe importance of epidemiology for estimation of risk from
RDP exposure, the Atomic Energy Control Board (AECB)
of Canada commissioned re-evaluations of the exposure
rates at both the Beaverlodge [S12, S14] and the Port
Radium uranium mine [S15].
269. Exposure conditions were re-evaluated for underground work areas at Beaverlodge between 1949 and 1968,
when individual employee exposure estimates began. The
re-evaluation of Eldorado Beaverlodge exposure rates suggested that previous estimates were underestimated byabout 50% [S12]. The re-evaluation study noted that exposure rates varied significantly (i.e. by more than a factor of10) between different areas of the mine. As a result of thisworkplace variability, substantial uncertainty in individual
employee exposures resulted when an average mine-wide
estimate was assigned.

251

because this method is complicated, most samples taken
were analysed for radon only.
273. According to Bloy [B11], a few measurements of
RDP concentrations were made in the 1954 survey using
the Kusnetz method. Over time, an increasing proportion of
the measurements were done in terms of RDPs. By the mid1960s, the Beaverlodge ventilation department was relying primarily on RDP measurements to assess ventilation­conditions in the mine.
274. During the period 1954 to mid-1962, mining engineers also began measuring radon gas concentrations in
the Beaverlodge mine as a means to determine the adequacy of mine ventilation, and not necessarily to determine miner exposures to RDPs. Radon concentrations
were ­converted to RDP concentrations by use of equilibrium factors determined in the years 1954, 1956, 1959 and1961 from the simultaneous measurement of radon andradon decay product concentrations. A large amount ofdata from the mine operating statements and radiation logbooks was captured in an electronic database by the 1991
SENES analysis [S12].

270. A detailed investigation of Eldorado Beaverlodge
records was conducted to further improve exposure estimates
for a case–control group [S12]. Previous mining experience(including experience in gold and other uranium mines) was
noted for several of the underground employees; however,these records were largely incomplete, and exposure estimates for this experience were not calculated. Improvements
to the Eldorado Beaverlodge exposure estimates involved
reviewing stope production records for the presence of­individuals from the case–control group. Exposure rates specific to the area and the time where the individuals worked,
rather than the mine-wide estimates, were then assigned tothe identified individuals. The individual exposures based on
specific mine areas were higher by a factor of 2–3 than those
based on mine-wide conditions.

275. Prior to mid-1962, most of the radiation measurements were for radon. In the years 1954, 1956, 1959 and
1961, as mentioned above, paired measurements were taken
where radon decay product and radon concentrations were
measured at the same time in the same workplace. The average radon/radon decay product equilibrium factors were
calculated from these data. Equilibrium factors for the earlyyears (1954–1956) appear to be generally lower than for later
years (1966–1968), and the equilibrium factors for the later
years at high RDP concentrations (>2 WL) approached andsometimes exceeded the theoretical maximum value of 1.0.
It was expected that the equilibrium factors for later years
would be lower because of generally improved ventilation
conditions. Radon concentrations without RDP measurements were multiplied by the equilibrium factor to estimate
the corresponding RDP concentrations [S12].

2. Radon and radon decay products

276. The total number of radon and radon decay productmeasurements taken per workplace per year was generally less than 12 and frequently as low as one. The average­during the period 1954–1968 was about four measurements
per workplace per year.

271. The first measurements of RDP concentrations were
performed in 1954 at Beaverlodge; further measurements
were made in 1956 as part of surveys of radiation levels,
dust levels and general ventilation conditions [F7]. These
initial surveys eventually led to a programme of radon andradon decay product measurements. Data from the 1954
survey indicated that simply turning the compressed air
on or off gave rise to substantial changes in workplaceconcentrations. Whether a miner worked within this envelope of fresh air or in the “unventilated” region outside
it was thus a very important factor in estimating his true
exposure.
272. Early measurements were for the purpose of providing
data for ventilation control. Originally, only the Tsivoglou
method was available for RDP measurements. However,277. The mine-wide underground estimates of exposurein units of WL produced in the 1991 SENES study [S12]are compared with the previous estimates in figure VI on
the basis of various interpretations of exposure data, either
WL measurements or WL inferred from radon measurements. The estimates are somewhat higher than those calculated using the medians of available exposure (WL) data,
which included the estimates used by Howe et al. [H19],
which in turn were based on those determined by Frost [F7].
­Figure VII shows that there is considerable variation in WL
values throughout the year — in winter, levels are higher
than in summer — and in different years.

252

UNSCEAR 2006 REPORT: VOLUME II

Figure VI. Comparison of radon decay product concentration estimates with the previous estimates for the Eldorado
Beaverlodge mine [S12].
9LEGEND:
Unpublished Eldorado estimates (based on means of WL data)

AVERAGE ANNUAL RDP CONCENTRATION (WL)8

Unpublished Eldorado estimates(based on WL means weighted by tonnage production)

7

1977 Eldorado back-calculation of WL(based on medians of WL data)

6

Frost 1983 (based on medians of WL and radon-derived data)
Frost 1989 (based on means of WL and radon-derived data)

5

Frost 1991 (mine-wide average WL)

4
3
2
1

1972

1971

1970

1969

1968

1967

1966

1965

1964

1963

1962

1961

1960

1959

1958

1957

1956

1955

1954

1953

1952

1951

1950

1949

0

YEAR

Figure VII. Seasonal variation of radon decay product concentration in Beaverlodge area mines operated by Eldorado [S12].
The high concentration in 1967 coincided with a fan failure.
LEGEND:

2.5

1963

AVERAGE RDP CONCENTRATION IN MINE EXHAUST (WL)

1964
1965
1966
1967

2.0

1.5

1.0

0.5
J

F

M

A

M

J

J
MONTHS

A

S

O

N

D

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

253

278. Several factors changed over time at the Eldorado Beaverlodge mine. The calendar years from the start of operation(1949) onward can be subdivided into five periods based on a
combination of the production and ventilation characteristics.
The period 1949–1953 covered the development of the minefrom early shaft sinking to the startup of the mill. The period
1954–1957 was a period of rapid expansion of the Eldorado
Beaverlodge mine as well as of many other uranium minesin the Beaverlodge area; mine ventilation increased over theperiod concurrent with substantial development work. The
period 1958–1962 was one of relatively little development
at Eldorado, with most of the other area mines closing. The
period 1963–1967 saw the development of two areas in the
Beaverlodge mine that were not ventilated by flow-throughmethods; during this time, all other mines in the area wereclosed. The period from 1968 onward covers the remainderof mine operation, during which the ventilation was increased
and individual personal exposure estimates were maintained.

280. An attempt was made [S12] to correct for these effects
by excluding data for stopes and development headings
where the mine operating statements did not identify any
work activity. This did not mean that work was not going onin those areas. Therefore this procedure underestimated theexposures of radiation technicians and other workers who
might, for various reasons, have been in those areas.

279. The 1991 SENES report [S12] discussed several
sources of uncertainty in the WL estimates, including thelack of measurements in some of the early years (pre-1954).
For these years, the workplace WL values were assumed tobe the same as those estimated for 1954 [S12]. For the years
in which no data were recorded, a linear interpolation of thedata from the nearest years for which data were available
was assumed. Even when measurement data are available,
the measurements were focused on detecting deficiencies inventilation as opposed to evaluating worker exposure.

282. Figure VIII illustrates, using data for the Fay area of themine before 1963, the variability in the workplace data and theeffects of agglomeration of the data. Moving from a mine-wide
stope average (left side of the figure) to individual work areas,a dramatic increase in the difference between the minimum
and maximum values (a factor of greater than 10) is evident.
Agglomeration of data reconstruction relied on work history
files from Eldorado. It was possible that the work histories ofthe study cohort were deficient with regard to non-Eldorado
employment. This observation is important, because miners

281. To correct for this, SENES [S12] weighted the reported
WL values by the recorded level of activity as measured in
man-shifts. However, there was oversampling in stopes relative to other areas with respect to the number of man-shifts
worked. Other problems in assessing exposure were: thereporting, in early measurements, of radon concentrations
rather than WL values; the uncertainty about the radon/radon
decay product equilibrium factor, which led to uncertainty inthe actual WL value; and the considerable variability in data
within and among workplaces and with time.

Figure VIII. Estimates of mean radon decay product concentration (WL) by level of aggregation of stoping workplaces, for
1963, Fay area of Beaverlodge mine in northern Saskatchewan [S12].

MEAN RADON DECAY PRODUCT CONCENTRATION (WL)

11
10
9
0455
STOPE

8
7
6
FAY WESTUPPER STOPES

5
4
3
2

FAY EAST
LOWER STOPES

1
0
ALL STOPING
WORKPLACESMINIMUM
MINE AREA

MAXIMUM
MINE AREA

1009
STOPE
MINIMUM MAXIMUMWORKPLACE WORKPLACE254

UNSCEAR 2006 REPORT: VOLUME II

could have worked in other mines for which no radiation exposure data were available. According to Howe et al. [H19], all
exposures prior to 1 November 1966 were recorded as single
lifetime totals rather than as separate annual exposures; thismakes them less useful for epidemiological purposes. Howeet al. [H19] further indicated that the (annual) median value
was used to describe (annual) average WLs.
3. Exposure estimation
283. In 1967, Eldorado began maintaining personal recordsof RDP exposure for full-shift underground workers. Workers’ time cards indicated the hours spent in each workplace.
These cards were consolidated into monthly printouts ofmanpower in each working place. These printouts, together
with measurements made in the workplaces and travelways,
were used to estimate monthly RDP exposures.
284. In 1970, the computer record system was expanded toinclude cumulative (i.e. lifetime total) exposure in the exposure summary reports. At this time, records of the previous
measurements made for ventilation purposes (as opposedto the measurements made for the purpose of estimatingminers’ exposures) were used to back-calculate exposures to
1 November 1966.
285. In September 1971, all maintenance, technical, supervisory and other personnel who had received some RDP
exposure were added to the exposure roll. Their exposures
were calculated back to 1 November 1966 using mine average working levels, hours worked per year, and a factor toaccount for the portion of time spent underground.
286. In the mid-1970s, work started on the estimation of
exposure prior to November 1966. Only RDP measurements
were used in this estimate, thus excluding the early survey
data and most of the measurements made during the 1950s,

Table 15

which had been for radon only. Because of the paucity of
pre-1964 data, Eldorado used an annual averaging process.
Owing to the known variability of workplace conditions,
Eldorado decided to use the median, as opposed to arithmetic mean, to describe the central tendency of the exposures.
On the basis of a review of work histories for individuals
employed prior to November 1966, a system of 22 job categories was devised; classification into a particular category
was based on potential RDP exposure. Each person was
assigned to a category for each job held throughout his or
her employment. An effective RDP concentration (WL) was
calculated for each job category for each year based on thefraction of working time spent in each area (underground,office, etc.). These calculations [F7], recorded as single lifetime totals rather than separate annual totals, provided thebasis for the exposure estimates in Howe et al. [H18].
287. SENES [S12] used the raw radon and radon decay
product survey data from the Beaverlodge mine to assign
WL values to individual stopes. Production data were thenused to retain only those measurements that were taken whenwork was in progress in each area. The production data wereused further to “weight” the data by the number of man-shifts
worked in each area at the measured RDP concentration.
The effect of this re-evaluation is shown in figure IX, which
illustrates, for the case–control sample of 195 miners, revised
exposure estimates plotted against the original exposure estimates. The 45º line corresponds to the situation when the two
estimates of exposure are equivalent. Figure IX shows clearly
that the majority of employees had revised exposure estimatesthat are higher than the original estimates. Although a general correlation exists between the two methods of estimating
exposure, the revised estimates are substantially higher, up
to an order of magnitude for some employees. For example,
two employees originally in the 5–49 WLM category were­reclassified to the 250+ WLM category. Table 15 shows theextent of movement from one exposure category to anotherdue to the revision of exposure estimates.

Changes in exposure categories of Beaverlodge miners based on revised estimates of cumulative exposures [S12]Original exposure(WLM) category

Revised exposure (WLM) category

Number of
employees

<5

5–49

50–99

100–149

150–199

200–250

250+

<5

61
(78)

16
(21)

1
(1)

0
(0)

0
(0)

0
(0)

0
(0)

78

5–49

7
(13)

36
(67)

7
(13)

2
(4)

0
(0)

0
(0)

2
(4)

54

50–99

0
(0)

0
(0)

9
(60)

2
(13)

3
(20)

0
(0)

1
(7)

15

100–149

0
(0)

0
(0)

1
(8)

0
(0)

5
(42)

5
(42)

1
(8)

12

150–199

0
(0)

0
(0)

0
(0)

0
(0)

0
(0)

1
(33)

2
(67)

3

200–250

0
(0)

0
(0)

0
(0)

0
(0)

0
(0)

4
(80)

1
(20)

5

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

Original exposure(WLM) category

Revised exposure (WLM) category

255

Number of
employees

<5

5–49

50–99

100–149

150–199

200–250

250+

250+

0
(0)

0
(0)

0
(0)

0
(0)

0
(0)

0
(0)

5
(100)

5

Number of employees

68

52

18

4

8

10

12

172

Note: N umber in brackets is the percentage of employees in the original exposure category that are also in the revised exposure category. Revised exposures were estimatedfor 172 of the 195 employees in the case–control group. The remainder were not traceable in company records.

Figure IX. Comparison between revised and original­estimates of exposure for Beaverlodge uranium miners(note that both axes have logarithmic scales) [C17].

REVISED EXPOSURE ESTIMATE (WLM)

1000

100

10

1

1

10

100

1000

ORIGINAL EXPOSURE ESTIMATE (WLM)

288. The 1991 exposure estimates [S12] did not include
non-Beaverlodge mining exposure. Because of the remotelocation, the cost of recruiting at Beaverlodge was high;consequently, Eldorado’s policy was to recruit experienced
miners wherever possible. Inclusion of non-Beaverlodge
exposure would probably result in further movement of theminers into higher WLM categories.
289. Beaverlodge miners were also exposed to airborne
dust. Information on workplace dust levels (konimeter data)that had been recorded coincident with the WL measurements made at Beaverlodge was also entered into the newcomputer database [S12]. For active stopes, an analysis was
conducted to determine if a correlation existed between RDP­measurements and dust. The analysis revealed significant
correlations between the measured values of dust, the RDP
concentration and the calculated equilibrium factor. Dustlevels were significantly (99% confidence level) positively
correlated with radon and radon decay product concentrations and with equilibrium factors. However, the association
between the dust concentration and the equilibrium factor
was the strongest of the associations. A correlation coefficient of 0.31 was statistically significant at >99% level. Thissuggested an association of higher equilibrium factors withthe higher dust levels in stopes with work activity; this seems
reasonable, because high dust levels and high equilibrium
factors both result from low ventilation rates. For the years
where both RDP concentrations and dust levels were captured in the SENES database, the mean dust concentrations
by year and type of workplace were calculated. The data indicated high dust concentrations in stopes in 1954 and 1956,
with mean concentrations of more than 500 ppcc ­(particles
per cubic centimetre) in stopes. The mean dust concentrationwas <100 ppcc for the years 1963–1965, and was 133 ppcc
in 1966. The highest mean dust concentrations were in the­raising workplaces; this is reasonable, because raises are­generally more difficult to ventilate during development.
290. SENES [S12] performed a detailed review of WLM
exposures for a case–control group defined by Howe. Subsequently, Howe and Stager [H16] carried out an analysisof the case–control group using the revised exposures andobserved that the inverse dose-rate effect was no longer
present. The revised [S12] methodology for calculation of
RDP concentrations was based on estimating annual mean
concentrations, taking account of the time spent by an individual employee (employee duration) at specific workplaces.
Annual workplace mean RDP concentrations were weighted
averages of monthly RDP averages, and the weights were
estimates of employee duration based on monthly production statistics.
291. Annual mean RDP concentrations for individual
workplaces were agglomerated on the basis of the hier­archy of workplace, mine area and mine-wide estimates.
Annual mean concentrations for mine areas were calculated
as weighted means of the annual workplace means, where
the weights were representative of the estimated employee
duration in each workplace. Similarly, the mine-wide annual
mean concentrations were weighted means of the mine-area
annual mean RDP concentrations.
292. This methodology provided estimated annual mean
RDP concentrations for individual workplaces and averagemine-wide RDP concentrations. As previously noted, substantial differences in RDP concentration existed between
different areas of the mine; this was related both to worker
activity and the ventilation infrastructure development inthose areas. Even within these mine areas, substantial differences in RDP concentrations existed. Although the method
worked well for stoping workplaces, application of the256

UNSCEAR 2006 REPORT: VOLUME II

method to drifting, raising, travelways and shaft area workplace categories was not as successful.
293. Available bonus contract data were examined for thecase–control group of 195 miners. Bonus contracts wereavailable for the periods 1955–1957, May 1960 and 1963–
1965. These records were searched for information on the
195 miners in the case–control group, of whom 129 had
underground work experience. For 61 of these employees,
information was located on specific workplace, and exposure durations were extracted from the bonus contracts. Total
cumulative exposures were calculated for the 191 employeeswith occupations and exposure durations described in the
personnel files. The total RDP exposure over 4,149 workermonths underground was 7,298 WLM for the miners withbonus contract information, based on the use of mine-wide

average concentrations. When mine-area average concentrations were used instead of mine-wide average concentrations
only for months with bonus contract information, the estimated exposure to the group was 7,570 WLM. A third estimate of exposure, using the ratio of mine-area to mine-wide
average RDP concentrations over all months with bonus
contract information, was 8,120 WLM.
294. Figure X shows the distribution for these 61 individual employees of the ratios of exposures calculated bymine-area concentrations to those calculated by mine-wide
average concentrations [C17]. This distribution is indicative of the variability in exposure estimates due to RDP
concentration differences between the areas in the minefor those underground employees with no bonus contract
information.

Figure X. Distribution of ratios of exposures calculated by mine-area concentrations to those calculated by mine-wide­average concentrations for Beaverlodge employees for whom bonus contract information was available [C17].
40
Summary Statistics
35

Observations 61
Minimum 0.4
10th Percentile 0.630

PER CENTMedian 0.8
90th Percentile 1.925

Maximum 4.4
Arithmetic Mean 1.2

20Standard Deviation 0.8
15

10
5

1.3

1.5
1.7
1.9
2.1
2.3
2.5
2.7
2.9
3.1
3.3
3.5
3.7
3.9
4.1
4.3

0.7
0.9
1.1

0.3
0.5

0.1

0

RATIO

295. An increase in average exposure for the case–control
group was seen when mine-area-specific concentrations
rather than mine-wide average concentrations were used.
This suggested the presence of a bias towards low individual
exposure estimates when mine-wide average RDP concentrations were used.
296. For a given cohort, the total exposure calculated
by summing the individual exposures that are based on

workplace concentrations and durations should equal themean exposure calculated using the total exposure duration multiplied by the duration-weighted mine-wide mean
concentration, irrespective of the variation or uncertainty
in concentrations between workplaces. However, this is notnecessarily so for nested case–control groups, where the proportion of cases is higher than in the cohort group. This typeof differential uncertainty is likely to be present in many of
the other cohorts, since mine-wide average concentrations

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

were used for estimating the earlier (and generally higher)
exposures.
297. Substantial additional RDP exposures, both from nonBeaverlodge work and from other environmental sources(particularly radon in dwellings), along with exposure toother lung cancer risk factors such as arsenic and additionalradon exposure that might occur in gold mining, were considered as possible confounders. Of the 195 employees in
the case–control study who joined the Eldorado company
with recorded previous mining experience, 9 had worked
in gold mines and 8 had worked in Beaverlodge area uranium mines. Kusiak et al. [K12] reported RDP levels of theorder of 0.3 WL (or greater) in gold mines in Ontario during
the 1960s, along with exposure to other risk factors, including exposure to silica and arsenic. The mean duration ofrecorded mining experience was 6.9 years. It is of interest
to note that 5 of the employees selected from the nominal
roll, the restrictions on which were to have excluded persons
with work experience at other Eldorado facilities, actually
had recorded work experience at Eldorado’s Port Radium
underground uranium mine.
298. Using the revised exposure estimates, Chambers
et al. [C17] investigated how uncertainty in exposure might
affect the dose–response relationship in the Beaverlodge
miners. A file describing a cohort with characteristics similar to that previously studied for determining the dose–
response relationship was constructed on the basis of partial
information from the nominal roll of Eldorado Beaverlodge employees. This information included an employee
identification code and the duration of each occupation foreach time period worked at Beaverlodge. Birth year or ages
were not available, nor was the vital status of individual
employees. Ages at the start of employment were assignedon the basis of random sampling from a uniform distribution of ages between 20 and 40 years. The cohort was similar in size and was assumed to be similar in characteristics
to the cohort studied in the epidemiological study [H19].
RDP exposures and doses were assigned to undergroundemployment up to 1967 on the basis of the revised exposure estimates and the algorithm developed for estimating
exposures. This provided estimates of nominal, or minewide average, annual exposures for each employee plus anestimate of the range of the exposures or variability due
to differences in exposure rates between different areas ofthe mine. Exposures were calculated by multiplying theexposure for a given year by the number of months that theemployee worked in that year. Exposures for other occupational groups and for post-1967 time periods were basedon previous estimates of exposure rates and an assumed
variability between workplaces [C17].
299. Expected and simulated observed numbers of lungcancer deaths were calculated using a life table with reference age-specific total mortality and lung cancer mortality
rates. The probability of lung cancer during each year of­follow-up was calculated for every employee in the cohort
using both the reference lung cancer rates and the lung257

cancer rates based on the RDP exposure and the true reference (unexposed) lung cancer rate for the employee. The lungcancer status was assessed for each year of the follow-up on
the basis of the probability of lung cancer for that employee­during the year, conditional on that employee being alive at
the start of the year [C17].
300. The exposure rates for an individual were probabilistically drawn from log-normal distributions that reflected
the variability in exposure rate for that occupation and time
period of employment. The mean of this distribution was
equal to the mine-wide (or occupation) average conditions
and was retained as a nominal estimate. The true exposure rate reflected the variation in exposure rates within
the mine (or occupation). True reference lung cancer rates
were based on the nominal reference lung cancer rate but
included a modification based on individual employee variations in this reference rate. These variations could reflect
interindividual variability in background rates or varying
exposures to other lung cancer risk factors such as may
have existed in other types of mine, or variability in theextent of smoking [C17].
301. Employees were assigned to the exposure categories
used in the 1986 cohort analysis, and the simulated numberof lung cancer deaths was determined. The expected numberof lung cancer deaths was determined by summing theannual probabilities of lung cancer death during the followup. Relative risks (RRs) for the exposure category were
determined by dividing the simulated number of lung cancers by the expected number. Relative risk coefficients were
determined by dividing the excess relative risk (RR – 1) by
the average cumulative exposure in the exposure category.
The relative risk estimates and the relative risk coefficientswere then summarized by scenario to show the variability in
exposure response [C17].
302. Figure XI represents a summary of the simulated dose
response based on the simple relative risk model with a risk
coefficient of 2 per 100 WLM, no uncertainty in dose andno confounding factors. The three lines show the maximum,
median and minimum simulated values of the relative risk
coefficient, with the median relative risk coefficient following the 2 per 100 WLM value assumed to be “true” in thesimulation. Any observed dose response falling within thetwo outer lines would be consistent with the assumed modelfor the scenario. The separation between the outer lines
reflects solely the statistical variation in the outcome of arandom process, since for this scenario there is no uncertainty in exposure and no confounding factors. For example, between 3 and 7 lung cancer deaths would typically berealized if 5 lung cancer deaths were expected from the risk
model. The distance between the bounds is related to the statistical power in that either a large effect or a large study
population is required for statistical significance. A mortality
update, such as that reported in reference [H35], decreases
the distance between the upper and lower confidence bounds,
since the predicted number of lung cancers deaths is higher
and the relative variability is reduced [C17].

258

UNSCEAR 2006 REPORT: VOLUME II

Figure XI. Simulated exposure response for Beaverlodge miners [C17].
Shown are the maximum, median and minimum values of RR per unit exposure, based on a median RR of 2 per 100 WLM.
108Maximum
MedianMinimum

RR COEFFICIENT

6

?

x

x indicates original (1986) estimate for
exposure category multiplied by 0.65

4

x
2

x

x
x

x

0

-2
0

100

200

300

400

500

MEAN CUMULATIVE EXPOSURE (WLM)

4.

Epidemiological studies

303. The Eldorado epidemiological cohort had 19,370
work records for only 18,424 persons, meaning that up to
946 individuals worked at more than one Eldorado site [N6].
Many of these miners worked at Port Radium, which was
Eldorado’s first mine; therefore some miners in the Beaverlodge study are likely also to have received exposure whileat Port Radium.
304. The mortality study of Ontario miners [M19] had also
identified 1,430 former Eldorado employees from Saskatchewan who had worked in an Ontario mine; of this number,
726 had worked in Ontario uranium mines. Consequently,
Muller et al. excluded the Eldorado employees from theiranalysis of Ontario uranium miners because of the lack of
exposure data for the time when they worked in the Eldorado
mines. Conversely, the Beaverlodge study [H16] included
many of these individuals but failed to take account of their
non-Eldorado (Ontario) exposure.
305. Howe and Stager [H16] reported on a study ofBeaver­lodge miners that was part of a larger study of some
18,000 Eldorado employees. The Eldorado epidemiology
project has been followed in a series of papers published by
Eldorado [A1]. Owing to the size of the study population,

Howe et al. were not able to interview the Eldorado employees. The exposure reconstruction had to rely on work history
files from Eldorado. It was therefore possible that the work
histories of the study cohort were deficient with regard to
non-Eldorado employment. This observation is important,because miners could have worked in other mines for which
no radiation exposure data were available.
306. Howe [H35] reports an updated analysis of a cohortof 17,660 individuals known to have worked for Eldorado sometime in the period 1930–1999 [H35]. One of thesubcohorts of Eldorado employees includes undergroundminers employed by Eldorado at the Beaverlodge uranium mine in northern Saskatchewan. The study designfor the updated analysis was very similar to that used in
the original study. The nominal roll was that used in theoriginal study [H16] with the addition of workers who hadjoined the Beaverlodge operation between the cut-off ofthe original study (31 December 1980) and the final shutdown of the mine in 1982. A considerable effort was madeto improve the quality and quantity of the data that wereextracted from the nominal roll. This resulted in somedeletions and additions for the pre-1980 period [F19].
Exposure estimates and estimates of gamma ray doses
were accumulated for the cohort, partly from the original cohort records available at Eldorado supplemented

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

by records from the Canadian National Dose ­Registry.
Also, work histories and dose records for non-Eldorado
uranium mining employment were added. This included
miners who worked in Ontario mines as well as a few from
Newfoundland. Thus, if a Beaverlodge miner had exposure from RDPs from work at other locations (i.e. Port
Radium mine, Port Hope uranium processing plant, nonEldorado mines around Uranium City, Ontario mines and259

Newfoundland fluorspar mines), this exposure was added
to the miner’s total exposure. Exposures received up until
1999, if a miner continued to work, were also included.
The updated study adds a further 19 years of mortality
data for the Eldorado cohort and also includes lung cancer
incidence results for 31 years, i.e. 1969–1999 [H35]. The
basic characteristics of the updated cohort, including the
Beaverlodge uranium miners, are summarized in table 16.

Table 16 Basic characteristics of the updated Eldorado cohort [H35]
Characteristics

Number

Per centMales

16 236

91.9

Females

1 424

8.1

Port Hope

3 003

17.0

Port Radium

3 300

18.7

Beaverlodge

10 05056.9

Other sites

1 307

7.4

1900

414

2.3

1901–1910

1 028

5.8

1911–1920

1 803

10.2

1921–1930

4 030

22.8

1931–1940

3 913

22.2

1941–1950

2 790

15.8

1951–1960

3 118

17.7

1960+

564

3.2

Sex

Site

Birth year

Cohort/subcohort

Mean RDP exposure (WLM)

Standard deviation

Entire cohort

48.0

182.6

Port Hope

12.5

43.4

Port Radium

174.2

369.1

Beaverlodge

23.2

81.7

Other Eldorado sites

1.9

32.9

307. Two general types of comparison were used in theanalysis of the Eldorado cohort data. Firstly, observed andexpected values were used to estimate standardized mortality ratios (SMRs) and standardized incidence ratios (SIRs).
Expected values were derived from Canadian national population rates for mortality between 1950 and 1999 and forcancer incidence between 1969 and 1999. A second series ofcomparisons were based upon internal comparisons between
subgroups within the cohort, i.e. with no reference to an
external population.

308. The analysis of mortality rates showed that, while mortality from lung cancer was elevated, the cohort as a whole
and the various subcohorts had reduced risks relative to the
Canadian population for most of the other causes of death.
The analysis of mortality from lung cancer among men in thecohort with respect to RDP exposure was based on 639 lungcancer deaths. (This compares with previous analyses of
the Eldorado cohort where the total number of such deaths
was 122.) For the Beaverlodge underground miner subcohort, there were 198 lung cancers observed compared with

260

UNSCEAR 2006 REPORT: VOLUME II

120.7 expected. The SMR for lung cancer was estimated at 1.6(95% CI: 1, 1.9) and was statistically significant (p < 0.001).

exposure, exposure rate and age at risk, was also investigated. Using the same approach as the BEIR VI ­Committee[C20], Howe estimated parameters from the present study
for a “full interaction model”, which accounts for the influence of age at exposure, dose, dose rate and time since
exposure (see table 17). The BEIR VI Committee’s analysis
was based on 11 studies of underground miners, including
the previous analysis of the Port Radium and Beaverlodge
cohorts. In this study, Howe found that the addition of boththe time since exposure terms and the six exposure categories resulted in a statistically significant improvement infit, but that the addition of age at risk terms did not. He
suggested that these results may be regarded as essentiallyindependent of the data used by the BEIR VI Committee[H35]. Howe’s 95% confidence limits on estimates of timesince exposure (WLM5, WLM15 and WLM25) parameters
include the BEIR VI estimates.

309. Comparisons of the cancer incidence rates between
1969 and 1999 for the cohort with those for the general
Canadian population showed that the only cancer which isconsistently elevated is lung cancer. For cancer as a wholeand for specific cancers, incidence rates for the cohort weregenerally lower than those for the general population, which
was considered to be a manifestation of the healthy worker
effect. Howe also investigated mortality and cancer incidence
for diseases other than lung cancer and found no evidence ofany causal relationship between exposure to RDPs or gamma
exposure and an increased risk of any other diseases [H35].
310. The application of the BEIR VI type of risk model,
which allows for effect modification with time since

Table 17 Parameter estimates for full interaction model and comparison with BEIR VI model estimates for males in the­Eldorado cohort (1950–1999) [H35]
Parameter

Estimate

95% lower limit

95% upper limit

Estimate for BEIR VI

WLM 5

5.23

1.33

14.52

7.68

WLM 15

2.5

0.63

7.05

5.99

WLM 25

1.37

0.36

3.99

3.92

Rate(1)1

Rate(2)

1.02

0.39

2.67

0.49

Rate(3)

0.49

0.2

1.21

0.37

Rate(4)

0.35

0.12

1.01

0.32

Rate(5)

0.33

0.13

0.84

0.17

Rate(6)

0.16

0.06

0.44

0.11

Age(1)

1

Age(2)

1.94

0.77

4.89

0.57

Age(3)

1

0.37

2.72

0.29

Age(4)

0.05

06 266.67

0.09

Parameters as specified below:
WLM = total WLM (per 100 WLM) lagged by 5 years
WLM 5 = WLM 5–14 years previously (per 100 WLM)
WLM 15 = WLM 15–24 years previously (per 100 WLM)
WLM 25 = WLM 25 years+ previously (per 100 WLM)
Rate (2) = WL 0.5–1.0
Rate (3) = WL 1.0–3.0
Rate (4) = WL 3.0–5.0
Rate (5) = WL 5.0–15.0
Rate (6) = WL 15+
Age (2) = age at risk 55–64
Age (3) = age at risk 65–74
Age (4) = age at risk 75+

1

1

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

5.

Evaluation

311. The Eldorado Beaverlodge cohort was updated withthe revised dosimetry [S12, S14] to add a further 19 yearsof mortality data and to include cancer incidence results for
31 years, i.e. 1969–1999 [H35]. The study design for the
updated analysis was very similar to that used in the original
study [H16]. The nominal roll was that used in the original study with the addition of workers who had joined the
Beaver­lodge operation between the cut-off of the original
study (31 December 1980) and the final shutdown of themine in 1982. The updated analysis found little evidence of
departure from a simple linear exposure–response model.
312. For the Beaverlodge cohort, an ERR of 0.96 (95% CI:
0.56, 1.563; p < 0.001) per 100 WLM was reported [H35].
This can be compared with the previous Beaverlodge estimate of ERR of 3.25 per 100 WLM [H16]. The estimate of
ERR for the Beaverlodge cohort has decreased substantiallyin the new analysis. Howe suggests that this could in partbe accounted for by the time-dependent effect modifiers on
the ageing Beaverlodge cohort [H35]. Extensive work wasundertaken in this update to add the exposures received in
non-Beaverlodge mines to those received in the Beaverlodge mines. Thus the decreased ERR could also be partly
explained by the addition of these non-Beaverlodge mine
exposures.
F.

Germany: Wismut miners
1.

Introduction

313. The Erzgebirge (Ore Mountains) of Saxony ­(Germany)and Bohemia (Czech Republic) have a long ­history ofunderground mining. As early as the 12th century, silver­mining was performed, while later other metals, such asiron, ­bismuth, cobalt, nickel and tungsten, were mined. The­mining of uranium started at the beginning of the 19th century in the Schneeberg area. Miners often died of what wascalled Schneeberger lung disease, named after the town in
the Erzgebirge. By the end of the 19th century, this disease
was recognized as lung cancer.
314. Shortly after the Second World War, the Wismut­mining company carried out uranium mining in Saxony
and Thuringia in the former German Democratic Republic(GDR). According to Jacobi and Roth [J3, J4], large-scaleuranium mining started in the Erzgebirge of Saxony in 1946
and later was extended to the eastern parts of Thuringia.
The Wismut mines in Saxony were high in arsenic, while
the Thuringia mines were low in arsenic. Overall, Wismut
produced about 220,000 tons of uranium between 1946 and1990, making it the world’s third largest producer of ­uranium[K15]. Mining for uranium per se ended in 1990 following
the reunification of Germany.
315. Various papers discuss the potential exposure conditions in the uranium mines in the GDR [E3, E4, K5, K15,

261

K16, L3]. Enderle and Friedrich [E4] characterized the workplace situation in the post-war years (to 1955) as compulsory
labour, use of prisoners of war (almost 50% in 1947) and ahigh rate of illness and accidents. In August 1953, a treaty
was signed to convert Wismut from a Soviet enterprise to a
Soviet–GDR company, and this resulted in improvements inthe working conditions [E4]. From 1946 to about 1955, theunderground mine conditions in the Erzgebirge of Saxony
were characterized by dry drilling, no mechanical ventilation, very heavy manual work, the absence of industrial
health and safety standards, and very long working hours.
316. According to Jacobi and Roth [J3, J4], three time
periods can be distinguished: the years 1946–1954, whichwere referred to as the “wild years”; 1955–1970, during
which time, there was ongoing improvement in the conditions through the introduction of wet drilling and improved
ventilation; and the period after 1970, when individual exposures were recorded and compared with ICRP limits. Some
20,000 cases of silicosis and 7,000 cases of lung cancer arereported among the Wismut miners. Jacobi and Roth [J4]noted that radiation exposures in the years preceding 1970
can only be “roughly” estimated. For 1955, on the basis of theassumptions of a mean radon concentration of 120 kBq/m3
for drilling and ore exploration areas and 50 kBq/m3 for otherworksites, and the nominal time spent in workplaces, a nominal workplace value of about 80 kBq/m3 was estimated. On
the basis of a nominal equilibrium factor of 0.5, workplace
exposure was estimated at 160 WLM/a. Prior to 1960, theconcentrations of long-lived alpha activity in the air could
have been higher by a factor of 100 to 1,000 than in post1960 conditions. Finally, the mean external dose to miners in
the last 10–20 years of mining were about 5 mSv/a.
2.

Radon and radon decay products

317. The first radon measurements in workplaces in the
Wismut mines were carried out in 1955 [J3]. The mean ofmore than 2,000 measurements carried out in five mining
operations was 110 kBq/m3, which according to Jacobi [J3]showed good agreement with measurements carried out in
1937 and 1938 in the Schneeberg mines. This author [J3],on the basis of 1955 data, reported a mean annual miner
exposure of 150 WLM, with a range of 30–300 WLM. (Theconversion from Bq/m3 to WLM depends on the equilibrium
factor Feq. Assuming an Feq of approximately 0.4, exposure
to a radon concentration of 110,000 Bq/m3 for 2,000 h corresponds, very roughly, to an annual exposure of 150 WLM.)
318. As reported by Lehmann [L3], from 1946 to 1955
there was no dosimetric recording of radon by the Wismut
company. In 1955, radon gas monitoring commenced. Measurement of radon and its decay products was introduced in1966 in Saxony and in 1975 in Thuringia. Formerly, for purposes of compensation, an average annual exposure to radon
decay products of 150 WLM was assumed by the Wismut
company for underground workers for the period 1946–1954[B5]. This was used independently of conditions and led to262

UNSCEAR 2006 REPORT: VOLUME IIboth over- and underestimation of individual exposures to
RDPs in the early years. A working group of experts wasconvened in 1993 to develop a job–exposure matrix (JEM)
in order to improve the estimation of exposure. The resultsof their work were published in 1998 [L3]. Radon concentrations for 1946–1954 were estimated retrospectively on
the basis of the first available radon measurements in 1955.
These estimates took into account previous working conditions in the mines, mine architecture, historical measurements and data gathered by the Czech ore mining industry.
Based on these estimates and the measurements available

since 1955, the annual exposures to radon, RDPs, long-lived
radionuclides and external gamma radiation were evaluatedfor each year of employment between 1946 and 1989, eachmining facility and each place of work (underground, milling or processing, open-pit mining or surface mining). Thisevaluation (see figure XII) was performed for a reference
job for each place of work (there were fewer reference jobs
for underground than for above-ground workers), while theexposures received in other types of job (of which there weremore than 200) were derived using weighting factors for the
specific reference job (ranging from 1.0 to 0.0).

Figure XII. Annual exposure to radon decay products as estimated using the job–exposure matrix for the job of “hewer”, infive mining facilities typical for the Wismut cohort [K16].
400
Objekt 02 (Oberschlema)

ANNUAL RDP EXPOSURE (WLM)

350
300

Objekt 01 (Johanngeorgenstadt)

250
Objekt 03 (Schneeberg)

200

Objekt 09 (Aue)

150
100

Objekt 90 (Schmirchau)

50

88
19

85
19

82
19

79
19

76
19

73
19

70
19

67
19

64
19

61
19

58
19

55
19

52
19

49
19

19

46

0

CALENDAR YEAR

319. The improved estimates of the JEM showed that RDP
exposures in the early years depended strongly on the numberof old shafts in a given mine and on mining activity. In newly
established mining facilities, such as Objekt 09 (Aue) andObjekt 90 (Schmirchau), radon values in the very first yearsof operation were rather low, while in old re­opened mining facilities, such as Objekt 02 (Oberschlema), Objekt 03(Schneeberg) and Objekt 01 (Johanngeorgen­stadt), radon
concentrations were already high at the beginning of theoperating period. Generally the levels of radon increased
with uranium mining and the area of the worked vein to a
maximum in 1955–1956, and decreased later owing to theintroduction of different ventilation and sealing measures.
320. The JEM [L3] provided estimates of exposures not
only to radon and its decay products but also to long-lived
radionuclides and gamma radiation. Owing to improved
working conditions, the exposure to radon and its decay
products decreased, while gamma exposures still remained

relatively high. A JEM for arsenic, fine dust and quartz is
being developed on the basis of estimates given in reference[B22].
3.

Epidemiological studies

321. Several epidemiological studies of radiation exposuresamong the Wismut miners are under way. These include alarge cohort study [K5], a nested case–control study on lungcancer mortality [T11] and a cohort study among the offspring of miners [T11], among others [G17, K15, K16, K29,
K30, K31]. There is also an independent case–control study
on the incidence of lung cancer [B23, B24, B39]. About
400,000 people worked for Wismut between 1946 and 1990.
For about 130,000 of them, complete working histories,
including start and end of work, job specification and placesof work (with dates) are available. From these, a stratified
random sample of about 64,311 people was drawn [G17,ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

K15, K16, K30, K31]. In order to reflect the different mining
conditions in the Wismut company mines, the sample wasstratified by the date of first employment (1946–1954, 1955–
1970, 1971–1989), place of work and area of mining. Criteria
for inclusion in the cohort study were as follows: (a) mini­
mum duration of employment of at least 6 months; (b) dateof first employment between 1946 and 1989; (c) year of birth
after 1899; and (d) male. After collection and subsequentevaluation of the occupational data, which were extracted
from the original payrolls, a total of 5,150 individuals were
excluded from the initial cohort, because they did not meetthe criteria. The final cohort for the analysis thus consistedof 59,001 men [G17, K30, K31]. On the basis of year offirst employment, three subcohorts were defined to reflect
the different mining conditions: 1946–1954 (subcohort A),
1955–1970 (subcohort B) and 1971–1989 (subcohort C).

Table 18 Characteristics of the Wismut cohort study [G17, K30, K31]
Characteristics

Number of cohort members

Per cent

59 001

100.0

1946–1954

23 91740.5

1955–1969

17 95030.5

1970–1989

17 13429.0

1946–1954

2 720

4.6

1955–1974

19 593

33.2

1975–1984

12 963

21.9

1985+

23 725

40.2

Alive

39 255

66.5

Deceased

16 59828.1   Cause of death available

14 64688.2   Cause of death not available

1 952

11.8

3 148

5.3

<10

3 764

6.4

10–19

11 225

19.0

20–29

12 536

21.2

30–39

12 70421.5

40+

18 772

31.8

<1960

224

1.3

1960–1969

1 255

7.6

1970–1979

3 132

18.9

1980–1989

5 368

32.3

1990–1998

6 619

39.9

Year of start of employment

Year of end of employment

Vital status as of 31 December 1998

Lost to follow-up

263

Duration of follow-up in years

Year of death

264

UNSCEAR 2006 REPORT: VOLUME II

Characteristics

Number of cohort members

Per cent

Malignant cancers (C00–C99)

4 800

32.8

Circulatory diseases (I00–I99)

5 417

37.0

Respiratory diseases (J00–J99)

1 559

10.6

Digestive system (K00–K99)

815

5.6

1 284

8.8

771

5.3

Never

8 244

14.0

Ever

50 70786.0

Cause of deatha Injuries and poisoning (S00–S99 and T00–T99)
OthersExposed to radon

Cumulative exposure to radon in WLM
Mean (maximum)

241 (3 244)

Median18

a Codes from the International Classification of Diseases.

322. In the first mortality follow-up, the vital status for thecohort was determined as at 31 December 1998. At that time,
66.5% were alive, 28.1% had died and follow-up was notcomplete for 5.3%. The mean age of subjects alive at theend of 1998 was 54 years for the total cohort and 71, 59
and 40 years in subcohorts A, B and C, respectively. A totalof 2,388 lung cancer deaths occurred in the first follow-up
period (1946–1998), which comprised 1,801,626 personyears. The general characteristics of the Wismut cohort aresummarized in table 18 [K15, K16].
323. Data on smoking habits were available for about athird of the total cohort. This proportion was considerably
lower for subcohorts A (20%) and B (33%) than for sub­cohort C (64%), reflecting the fact that smoking habits wererecorded only after 1970. More than 50% of the miners
with known smoking habits were heavy smokers, while theproportion of non-smokers was about 26% [K16].
324. Within the cohort study, a nested case–control study
on lung cancer deaths was conducted that included individuals born after 1927. Two controls per case were matchedby date of birth. Controls could be either alive or deceased,
Questionnaires were sent either to next of kin or, in the caseof the controls, to the miners themselves, if they were still
alive. Information was gathered on smoking habits andoccupational exposures outside the Wismut facilities, whichmight be related to lung cancer. Next, data were abstracted
from the Wismut health archives in relation to smoking­habits, medical radiation exposures and jobs prior to Wismut
employment [G14, T11].
325. From the cohort, a subsample of 6,000 miners was
drawn as the basis for an offspring cohort study. The lifetime exposure of the 6,000 miners to radon and radon decay
products varied between 0 and >3,000 WLM. The first stageof the study, which is in progress, was to identify the children of miners to be included in the study and to investigate
their health status, life expectancy and causes of death. The
most important outcome variables are genetic anomalies,
infant mortality and childhood cancers. The offspring cohort
consists of 7,855 children.
326. Another case–control study on lung cancer incidence
among former Wismut employees was conducted between
1991 and 2001 [B23, B24, B39]. Patients with histologicallyconfirmed primary lung tumours were recruited from several study hospitals in Thuringia and Saxony. Controls wererandomly selected from the personnel files of the Wismut
company and were frequency-matched to the cases according to birth year in 5-year groups. Inclusion criteria for casesand controls were: male workers; employed underground at
the Wismut Company at some time between 1947 and 1990.
Occupational exposure to radon, its decay products, gamma
radiation and long-lived alpha emitters was estimated byusing the JEM described in reference [L3]. All subjects werepersonally interviewed about occupational and smoking history. Lung cancer risk was calculated by using the ERR modeland conditional logistic regression. A total of 505 cases and1,073 controls were included in the study. The cumulative
exposure from RDPs ranged from 1 to 2,911 WLM (an average of 552 WLM for the cases and 420 WLM for the controls). The exposure rate ranged from 0.1 to 31.4 WL (anaverage of 8.2 WL for the cases and 7.2 WL for the controls).
The odds ratios (OR; adjusted for smoking, year of birth andasbestos exposure) in the two highest categories, comparedwith the reference category of 50 WLM, were significantly
increased: ≤800–1,599 WLM, OR = 2.08 (95% CI: 1.40,
3.08); and 1,600–2,911 WLM, OR = 3.68 (95% CI: 1.92,
7.03). More than half of the study subjects had been exposed
more than 35 years earlier. Assuming a linear exposure–
response relationship, there was a significant increase in the

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

relative risk of 0.10 (95% CI: 0.05, 0.17) per 100 WLM after
adjusting for smoking and asbestos exposure. After correcting, in a sensitivity analysis, for the fact that the controls of
this study had a higher average exposure than the population of Wismut workers from which they had been recruited,the ERR increased to 0.24 per 100 WLM. For those stillsmoking, the increase in relative risk was lower (0.05 per
100 WLM), whereas it was higher (0.20 per 100 WLM)
among non-smokers and long-time ex-smokers. Lung cancer risk declined with time since exposure, except for thoseminers who had been exposed 45 or more years in the past.
No inverse dose-rate effect was observed.
327. More recently, the autopsy data in the archives of theCentral Institute of Pathology of the Wismut Company wereanalysed for 19,271 persons; these included 12,926 uranium
miners (WLM > 0, group 3), 1987 control cases of nonexposed Wismut workers (WLM = 0, group 2) and 4,358 control cases most likely never employed by Wismut (group 1)[W17]. The mean age at the start of exposure was 33.5 years,
mean duration of exposure 11.8 years and mean age at death62.7 years. The autopsy data investigated comprised about
152,300 person-years of uranium mining work with radon
exposure. Mean RDP exposure was about 725 WLM, with amaximum of more than 3,000 WLM. Mean exposure to longlived radionuclides (LRN) was about 7.3 kBq h m–3 (238U),
with a maximum of more than 50 kBq h m–3. On the basis of
the main cause of death or concurrent diseases, 8,882 casesof malignant tumours were found (6,403 for group 3,889 forgroup 2 and 1,581 for group 1). For primary malignant lungtumours, a higher incidence was found for exposed Wismut
workers (4,526 (35%) in group 3, 377 (19%) in group 2
and 2,472 (11%) in group 1). For primary malignant lungtumours, a significant correlation between OR and WLM
category was shown. Depending on the chosen category,
the OR for groups with high exposure increased to 10. In
categories of high LRN exposure, the OR increased to 8.4(95% CI: 6.19, 11.56). Below the 400–599 WLM category,no significant difference of OR related to the leading histomorphological tumour types could be determined. In higher
WLM categories, small cell carcinomas and squamous cell
carcinomas showed an OR nearly twice as high as that of
adenocarcinomas. For non-exposed Wismut workers, 6%
of non-smokers and about 19% of smokers showed lungtumours, and for exposed workers, 18% of non-smokers and
34% of smokers had developed lung tumours. The relative
share of small cell carcinomas and squamous cell carcinomas was higher for smokers than for non-smokers. In turn,the relative share of adenocarcinoma was lower in smokers
than in non-smokers. According to autopsy diagnosis, therelative share of silicosis cases increased with the WLM
category. This mirrored, at least for the mining sites in Saxony, the parallel exposures to quartz-containing dust andionizing radiation. The available data gave no indicationof a causal connection between silicosis as the cause ofdeath and primary malignant lung tumours for the W­ ismut
workers. Molecular biological investigations revealed noevidence for repeated mutations of codon 249 of the p53
tumour suppressor gene in the Wismut workers, as had

265

been shown for uranium workers of the Colorado Plateau
in reference [T1].
328. Pathological findings on 243 Wismut uranium miners with lung cancer, recruited between 1991 and 1995 into
the case–control study on lung cancer incidence [B24], arereported elsewhere [K14]. The frequencies of all tumour cell
types were found to be associated with increasing RDP exposure, but high radiation exposures tended to increase the relative proportion of small cell lung cancers and squamous cell
carcinomas [K14]. This effect was more pronounced among
those who had stopped smoking or had never smoked, andit seemed to be masked among those still smoking [K14].
The first evaluation of the pathology archive of the Wismut
company showed a shift from small cell lung carcinomas
as the predominant cell type in the first follow-up years to­squamous cell carcinomas in the later years [W4]. At present,
5,215 lung carcinoma cases have been identified among
former Wismut employees, showing extremely high proportions of small cell lung carcinomas (69%) in the early years(1957–1965), which declined to 34% up to 1990 [W4].
4.

Evaluation

329. A number of reports have presented information on thepredicted numbers of lung cancers in former Wismut miners[B5, J3, J4]. On the basis of a sample of 3,654 persons drawn
from the Wismut database, which resulted in a final data set
consisting of 2,282 men, and using 1985 death tables for theGerman Democratic Republic and various risk projection
models, one analysis [B5] predicted a further 1,700 to 4,800
additional lung cancer cases from 1995 onward. The peak
incidence was predicted to occur between 1985 and 1991.
Another study on pathological findings among Wismut
uranium miners was published in 2006 [T50].
330. The main strengths of the Wismut cohort study areits size, a wide range of exposure levels, a long duration of
exposure and a large number of cases of lung cancer andother diseases, as well as the availability of informationon dust and arsenic exposure. A joint analysis of 11 minercohort studies [L10] was based on a total of 60,606 exposedminers, including 2,674 with lung cancer, and a mean radon
exposure of 164 WLM. Data were combined from different cohorts of miners around the world. Heterogeneity withrespect to the quality of exposure assessment, the presenceof relevant covariates such as arsenic, dust and tobaccosmoke, as well as lifestyle and genetic factors, was likely
to be present in this combination of cohorts. The Wismuth
cohort provides a data set for analysis that is similar in size
(59,001 miners and 2,388 lung cancer cases) to the data set
used in the combined analysis, yet it is more homogeneouswith respect to data collection and estimation of exposure.
Therefore it represents a unique opportunity to verify the
results of the combined analysis in an independent data set.
The potential limitations of the Wismut cohort are the limited information on smoking and the limited validity of the
exposure assessment, particularly in the years before 1955.

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UNSCEAR 2006 REPORT: VOLUME II

331. A number of reports have already presented data on
the feasibility of epidemiological studies of Wismut miners
and have provided some early results [K5, K15, K16, T11].
Recent observations suggest a linear smoking-adjusted ERRof about 0.10 (95% CI: 0.05, 0.17) per 100 WLM [B24,
B39]. As noted in reference [T11], the results of future epidemiological studies will depend greatly on the quality ofthe exposure assessment and the information on vital status
and causes of death.
G.

Canada: Port Radium miner study
1.

Introduction

332. The Eldorado Port Radium mine has a long history
extending back to 1930, when a prospector by the name of
Gilbert LaBine identified pitchblende on the north shoreof the Great Bear Lake in Canada’s Northwest Territories.
Open-pit mining started shortly thereafter, and the first pitchblende was shipped in 1931. Underground operations wereat a high level of activity by 1932. Mining continued until
1940, when a decreasing demand for uranium ore led to themine being shut down. In 1942, the Port Radium mine wasreopened at the government’s request, and it continued in
production until 1960, when it was shut down. Annual average ore grades in the 1940s were 0.5–1.2%, with occasional
pockets of high-grade pitchblende being encountered.
333. Bloy [B11] noted that ventilation in the Port Radiummine was somewhat unusual. At the start of undergroundwork, the mine had only natural ventilation; this was greatly
reduced in winter, as all openings to the surface were kept
closed because of the extremely cold temperatures. Openings to the surface other than the shaft, such as raises fromstopes through the surface pillar, were covered and tightly
sealed in the winter. The only air entering the mine, in winterwas compressed air used to power the drills [T3]. Thompkins conducted a ventilation survey at the Port Radium minein 1945 [T4]. There was only a relatively small amount of air
circulating through the mine. This condition, together withthere being no definite routing of the air currents and thelack of dust-reducing features on mine machinery, resulted
in generally high dust concentrations in the mine. As part ofthe 1945 ventilation survey, Thompkins [T4] estimated thebalance of the air entering and leaving the Port Radium mineby natural means. Airflow from the surface entering throughthe 921 Raise and the 722 Raise was estimated at 455 and1,500 cfm,3 respectively. Compressed air entering the mine
was estimated to be 1,400 cfm. Therefore the total volume
of air entering the mine was approximately 3,350 cfm. The
volumes of air leaving through the shaft and the manway/
pipe raise were estimated as 2,350 and 850 cfm, respectively,
for a total of 3,200 cfm. Only 150 cfm of the air that entered
the mine remained unaccounted for [T4].
334. By 1944, the need for heated ventilation at the Port
Radium mine was recognized. A ventilation unit and steam
3

1 cfm = 4.72 × 10–4 m3 s–1.

plant for heating air going into the mine were installed in1946 and were in operation by early 1947.
335. By 1956, ore zones at the Port Radium mine hadreached a stage where it was necessary to relocate the minesurface ventilation unit and steam heating unit to accommodate active workings. Owing to the apparent short life
expectancy of the mine, driving a complex ventilation raisesystem was deemed impractical. As an alternative, old manways and workings were used to bring the air to active areas.
Air was forced into the mine down 136 Raise into 136 Stopeand down to the first level through the stope raises. From
there, the air was channelled to the lower levels through aseries of old raises. The air on the levels was controlled by aseries of vent doors [B14]; there were, however, substantial
air losses between the first and fifth levels as a result of air
escaping through old stopes filled with broken material. For
instance, at the first level, a loss of 7,000 cfm was noted.
Bloy [B14] concluded that it was virtually impossible toadequately ventilate the winze section with the existing setup, since air volumes were substantially reduced and any air
reaching the area was highly contaminated.
336. In 1956 and 1957, the main fan on the surface was
stepped up from 22,000 cfm to 35,000 cfm. Heating facilities were increased to handle the extra airflow in winter. By
then, a definite air route was established underground. However, when the installation was completed and the air volume
checked, only 6,000 cfm of the approximately 35,000 cfm of
air put into the mine actually reached the active workings, andthis was contaminated with radon and dust. The rest of the air
escaped through leaking doors, bulkheads and coarse backfill.
To solve this problem, the airways were lined with two typesof plastic: polyethylene plastic sheeting of about 4 mil thickness (0.1 mm) and a liquid spray plastic called cocoon [B12].
337. The plastic sheeting lined the raises from the first levelto the surface fan inlet. Although the sheeting worked very
well, it was difficult to handle and tore easily. By the timethese raises were completed, the liquid spray plastic method(cocoon) was being tried, and the results were so encouraging
that the system was used for all remaining airways [B12].
338. An auxiliary fan delivering approximately 11,000 cfm
was located on the 11th level to ventilate the winze area,
thus supplying this area with much better ventilation than
had been found in the 1956 survey. A raise system was also
completed in the winze area, which greatly improved theventilation [B13]. In addition to the cocoon method, three
additional auxiliary fans were added in the winter of 1957–
1958, resulting in a significant lowering of radon concentrations by the spring of 1958 [F5]. No ventilation data for theperiods after 1958 were found in the records of Cameco, thelater owner, or the literature review.
339. Kupsch [K6] provided a detailed history of the Eldorado company, including its Port Radium mine. Additional
interesting information on the history of the Port Radiummine was provided by McNiven [M6]. He reviewed theANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

operations at Port Radium from the reopening of the mine in1942 to its final shutdown in September 1960.
2. Radon and radon decay products
340. Port Radium was the first mine in Canada where radonand radon decay product sampling was performed. According
to Bloy [B11], the first radon samples at this mine were taken
in February 1945. The reason for the study was to investigate
the “suffocating gas” noticed by the miners. (The suffocating
gas turned out to be simply oxygen deficiency.) Radon levels
were found to range from 13,000 to 47,000 pCi/L.
341. Representatives from the Chalk River Nuclear Laboratories visited the Beaverlodge and Port Radium operations
and carried out preliminary surveys for radon in 1951 and1952 [S15, S30]. The surveys established that both Beaver­lodge and Port Radium had serious radon problems. By
1954, Chalk River Nuclear Laboratories had designed andbuilt new radon sampling equipment which they used inanother survey of the Port Radium mine [B11].
342. It was the practice in those years to perform konimeter dust sampling from time to time in parallel with the radonand radon decay product sampling. In addition, an external
gamma radiation survey of the Port Radium mine in 1952
showed gamma exposure levels in a few areas of 95 mR(approximately 0.95 mSv) per shift [B11].
343. The radon and radon decay product data were discussed by Frost [F6]. Initially, only radon was measured. By
the mid-1950s, both radon and radon decay product concentrations were measured (the Kusnetz method was used for thelatter). The equilibrium factor between radon and its decay
products was found to be approximately 30% in 1957. Frost[F5, F6] concluded that the equilibrium factor was likely tohave been higher in the early years of mining when there was
no forced ventilation.
344. Average RDP concentrations were estimated by Frost[F5] to be of the order of 77 WL in 1945, 38 WL in 1952,
10 WL in 1956 and 8 WL in 1957, on the basis of the available measurements and assuming an equilibrium factor of
30%. Unfortunately, these data and the corresponding estimates of miner exposure have not been published.
Table 19345. In a 1996 re-evaluation [S15] of RDP concentrations,radiation measurements recorded by Eldorado staff werecompared with the duration of employee exposure for eachradiation measurement. All but a few of the radiation measurements were solely for radon concentrations, as the threepaired measurements of radon and radon decay product concentrations in 1957 did not provide enough information toestimate equilibrium factors. Thus a modelled equilibrium
factor based on the ventilation rate was used to convert radonlevels to WL RDP concentrations.
346. WL estimates for individual workplaces were aggregated as a function of the workplace classifications to which
the 171 employees selected for this study were assigned byHowe [S15]. At each level of aggregation, efforts were madeto provide a consistent estimator of WL. Since there were no
individual measurements of RDP concentrations, individual
WLM exposures could not be calculated directly. An estimate of the exposures of individual employees was based on
the mine-wide average.
347. The duration-weighted arithmetic mean was chosen in this analysis as the estimator to characterize RDP
concentrations. This statistic facilitated the calculation ofa mine-wide average for a variety of employee classifications. This mine-wide average (mine index) provided a WL
value that would be the expected WL concentration over
all the individuals in the classification. From a preliminary
review of the data for Port Radium provided by Cameco,
the amount of downcast ventilation into the mine varied
substantially (by about a factor of 2) by season of the year,
especially from 1947–1957. This was because air needed
to be heated in the winter to prevent the upper portion of
the shafts from freezing. The limited capacity of the heating
plant required the ventilation volume to be reduced during
the colder ­winter months (mid-December to mid-March) to
avoid freezing the workings.
348. On the basis of the ventilation characteristics, theentire operational period of the mine was divided into threesubperiods: pre-1947, 1947–1955 and post-1955. Table 19
summarizes the available radon data for shaft stations.
Similar data are available for active stopes and other underground workplaces. A large winter/summer difference isevident.

Radon concentration in shaft stations of the Port Radium mine [S15]PeriodSeasonRadon concentration (pCi/L) a

Pre-1947

Summer

8 318 (6)Winter8318 bSummer

2 084 (10)Winter

4 760 (6)Summer

3 434 (16)Winter593 (4)

1947–1955

Post-1955a	Arithmetic mean of measured radon levels with number of observations given inround brackets; 1 pCi/L = 37 Bq/m3

267

b Indicates an estimated value, since measurements were not available.

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UNSCEAR 2006 REPORT: VOLUME II

3.

Exposure estimation

349. Information provided by Howe allowed a comparisonof exposure duration and exposure (WLM) for a 171-member
case–control group. There were significant differences in theestimates of months worked between the 1996 re-evaluationand the original epidemiological study [H15]. Estimates of
total WLM for the case–control group as a whole are not significantly different between the two approaches. However,as for exposure duration, individual differences in exposure(WLM) could be very large. To assess the implications of
this re-evaluation for the epidemiological study, it would benecessary to look at the differences in the individual miner
exposures between the two studies. It is not known from theevaluation whether the cases follow the general pattern or if
a difference exists between the cases and the controls (which
was the expectation of the authors).
4.

Epidemiological studies

350. Since early miners had a potential excess risk of lungcancer, Eldorado sponsored a pilot epidemiological study ofPort Radium workers. This study [G5] found an excess oflung cancers in miners who had 5 years or more of underground experience.
351. Consequently, Eldorado initiated a more detailed epidemiological study [A1] that included the radiation exposure
data, and that involved Statistics Canada and the National
Cancer Institute of Canada, which performed the actual
epidemiological analyses.
352. This study by Howe et al. [H18] investigated some
2,103 miners employed between 1942 and 1960. In thisgroup, 57 lung cancer deaths were observed comparedwith 24.73 expected. Employment records were not available before 1940 and hence exposures before that date were
not estimated. As a consequence, the exposures of the Port
Radium miners may well have been underestimated. Risk
coefficients estimated from the Port Radium analysis should
therefore be regarded as upper limits [H18].
353. Radon gas samples were collected for seven of theyears between 1945 and 1958, with between 9 and 71 samples per year and a total of 251 samples. The range of­concentrations was reported as 50–300,000 pCi/L. Howeet al. [H18] cite Frost [F5] as the source of their RDP exposure data. The Port Radium study, unlike the Beaverlodge
study by the same investigators, in at least some circumstances, used the annual average rather than the median as
representative of workplace RDP levels. While radon gas­samples were made as early as 1945, early data are sparseand the uncertainties in the exposures are likely to havebeen very large. Howe et al. [H18] indicated that weighted
average equilibrium factors were calculated on the basis of
the known labour distribution and type of workplace. The
highest factor used for many work groups was 0.5, although
it was discovered that there could have been substantiallyhigher values. Working time was based on a 40-hour weekand 48 weeks per year. Howe et al. [H19] acknowledged that
there were many potential sources of error in the procedurefor ­estimating RDP exposure.
354. The average exposure of the 2,103 miners was
183.3 WLM. About 42% of the person-years at risk were in
the <5 WLM exposure category. The ERR was estimated tobe 0.27 per 100 WLM and the excess absolute risk was estimated to be 3.1 cases per 106 person-years per working levelmonth [H18].
355. The ERR per 100 WLM for Port Radium from the
updated study [H35] is essentially unchanged from the previous assessment [H18], whereas the ERR for Beaverlodge
has decreased substantially. Howe suggests that this could be
accounted for in part by the early exposures of Port Radium
workers, for whom the time-dependent effect modifiers (timesince exposure and age at risk) may be of less importancethan for the younger Beaverlodge subcohort [H35].
356. More recently, the Eldorado cohort was updated to
add a further 19 years of mortality data (1950–1999) andto include cancer incidence results for 31 years (i.e. 1969–1999) [H35]. The study design for the updated analysis wasvery similar to that of the original study [H18]. The updated
estimate of the ERR for Port Radium miners was 0.37 (95%
CI: 0.23, –0.56; p < 0.0001) per 100 WLM, which may becompared with the ERR of 0.27 per 100 WLM from the1980 mortality analysis [H18]. As noted previously, the current estimate of ERR for Beaverlodge is 0.96 per 100 WLM[H35], compared with the previous estimate of 3.25 per
100 WLM [H16].
5.

Evaluation

357. During the early years of mining (pre-1947), WL values in the Port Radium mine were of the order of 60, andannual exposures of miners were likely to have been of theorder of 600–1,000 WLM. By the end of mining (around1959), the calculated WL values declined to 2–3, and hence
annual exposures were likely to have been of the order of
20–40 WLM. The estimates have large uncertainties, perhaps a factor of 10, for both pre-mechanical-ventilation andpost-mechanical-ventilation periods [M5]. Few data are
available for their estimation, and the quality of the sparse
data available is suspect.
358. Exposure estimates were particularly uncertain for anumber of work types where the proportion of time spent
underground was unknown. This included mechanics andelectricians who may have worked underground for extended
periods. The uncertainty in risk estimates if these employees
are included in an epidemiological analysis could be substantial, and exclusion of these workers should be considered.
359. In addition to the sparsity and limitations of radiation and ventilation data for the early years, there exist otherANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

recognized sources of uncertainty. Based on the sample
of 171 miners provided by Howe, many pre-Port-Radium
employment histories are incomplete. Many Port Radiumminers are likely to have worked in other mining environments (notably gold mining), which provided a further possible risk factor. Epidemiology for Port Radium miners only
included miners’ work after 1941, as work records prior tothat time were not available. Prior, unrecorded experience at
Port Radium may be a risk factor in some instances. The Port
Radium ores contained significant concentrations of otherelements, including arsenic, nickel and cobalt. The degreeof confounding arising from these elements in workplace
dust is not known. Overall, the Port Radium cohort provided
evidence of the risk of exposure to RDPs. The 1996 exposure re-evaluation [S15] represents the best available data
for epidemiological assessment of the Port Radium cohort.
However, the exposure uncertainties are very large, andquantitative estimates of dose–response relationships must
be viewed as having substantial uncertainty.
360. Notwithstanding the re-evaluation of miners’ exposures, the Port Radium data set provides a much weakerbasis for dose–response investigation than, for example, theBeaverlodge cohort, and is of lower reliability for use in aquantitative assessment of risk.
H.

French uranium miners
1.

Introduction

361. Uranium prospecting began in 1947 in France, andthe production of the first tonne of uranium occurred in1949. Extraction was located in four main mining divisions:
Crouzille (Limousin) and Forez from 1947, Vendée from1953 and Hérault from 1977. It continued up to 1999, when
the last mine closed.
362. The first radon measurements were taken in 1953. In
1956, forced ventilation was introduced in the mines, leading
to a sharp decrease in exposure levels, and systematic control of individual exposures began to be applied in the mines.
Individual exposure was assessed for each miner on the basisof ambient measurements. Monthly individual records of
RDP and gamma exposures were kept in the mining divisions. By 1958, a regulation decreed limits for internal andexternal exposures. From 1959 onward, systematic records
of uranium ore dust exposure were kept for each miner. After
1983, the system of assessment of exposure was replacedby the use of personal alpha and gamma dosimeters [Z10].
Pradel and Zettwoog [P8] and Bernhard et al. [B46] describe
the radiation protection practices in French uranium mines.

269

of 100–10,000 pCi/L in mine air (based on 40 samples).
Groundwater from mineralized zones was an importantsource of radon. The paper noted that, in one mine, a crosscut was driven into barren rock and air was blown throughthe cross-cut. The air came into contact with radon-richwater, and within a distance of 300 m exceeded the workplace concentration limit for radon, referred to as the tolerance level (50 pCi/L of air in French mines at that time). The
paper also noted high levels of radon in the smoke generated
by blasting (levels of as high as 50,000 pCi/L).
364. The 1974 paper by Pradel and Zettwoog [P9] showedthat RDP measurements started in 1955 with about 65 samples being taken per mine per year. Beginning in 1955, miners were required to wear an individual dosimeter to record
their exposure to gamma radiation. Monthly exposures to
RDPs and long-lived radioactive dust were calculated, andall results of exposure were reported on personal cards.
Annual and lifetime exposures for all three components ofradiation dose — RDPs, gamma radiation and long-lived­radioactive dust — were obtained by summing monthly exposures over time. From Tirmarche et al. [T9, figure 6], about
2,500 ­person-years of exposure in the 1947–1956 period had
to be reconstructed, about 11% of the total person-years.
365. Pradel and Zettwoog [P8] commented that, prior to1953, only 40 measurements of radon had been made. Large
numbers of radon measurements were made later, and thisenabled a close approximation to be made of the inhaledquantities of radon. The authors also discussed the possibility of relatively elevated exposures for short periods oftime when mining was taking place in high-grade-ore areas.
Data from Duport [D9] showed that, in the period 1956–
1982, there were typically more than 30 radon measurements per person-year. This is very high compared with thenumber available for the other miner cohorts used for risk
assessment.
366. After 1983, the system of exposure monitoring basedon area measurements was replaced by personal alpha andgamma dosimeters. The portable device is described by­Zettwoog [Z10]. It comprises an active dosimeter with aspectro­graphic head for the measurement of alpha radiation(ionograph track detection) and a thermoluminescent dosimeter for the measurement of external gamma radiation. It
allows the measurement of the number of 222Rn atoms inhaled,
the energy of the alpha particles emitted by the three shortlived decay products of radon and thoron (218Po, 214Po, 212Bi),
the alpha activity of the five long-lived alpha emitters (238U,
234
U, 230Th, 226Ra, 210Po) contained in ore dust (after decay ofthe short-lived products) and the gamma radiation dose.
3.

Exposure estimation

2. Radon and radon decay products
363. A 1955 paper by Jammet and Pradel [J7] provided insight into the early conditions in the French uranium mines. This paper reported radon concentrations

367. Information about RDP exposure in the initial cohortof French uranium miners is found in references [L10, T5, T6,
T7, T9]. This group includes 1,785 miners who began underground work between 1946 and 1972 and who were exposed

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UNSCEAR 2006 REPORT: VOLUME II

for at least 2 years. The mean cumulative RDP exposure wasrelatively low (70 WLM). The reports describe the reassessment of the RDP exposure of miners by an expert group, notably for the period 1947–1956, for which retrospective estimation was necessary. Working conditions changed dramaticallyin 1956, with a large reduction in the exposure of miners [T8].
Tirmarche et al. [T7, T8] provided a distribution of individual annual exposure (WLM). Prior to 1956, median annual
exposures were estimated to be of the order of 10–11 WLM.
After 1956, annual exposures were in the range 1–3 WLM
until 1980 and were <1 WLM thereafter, as illustrated in figure XIII. The figure also shows that exposures prior to 1956
had a 5-fold range of interindividual variability.
368. Rogel et al. [R10] and Tirmarche et al. [T30, T31]
provided an update of the exposures of the enlarged cohortof French uranium miners. This group included 5,098
males, employed as miners by CEA-COGEMA for at least
1 year between 1946 and 1990. Among the 4,134 miners
who were exposed to RDPs, the mean cumulative exposure was 36.5 WLM (with a range of individual exposuresof 0.1–960.1 WLM) over a mean duration of 11.5 years(with a range of 1–37 years). The mean annual exposure
was 23.9 WLM before 1956, whereas after 1956 it was only
1.5 WLM [R10].

56
55
54
53
32
31
30
29
28
27
26
25
24
23
22
21
20
19
18
17
16
15
14
13
12
11
10
9
8
7
6
5
4
3
2
1
0

Upper quartile
MedianLower quartile

46
48
50
52
54
56
58
60
62
64
66
68
70
72
74
76
78
80
82
84
19

ANNUAL RDP EXPOSURE (WLM)

Figure XIII. Distribution of annual exposures to radon
decay products by calendar year in the initial cohort ofFrench uranium miners [T7].

YEARS

4.

Epidemiological studies

369. The initial cohort included 1,785 uranium miners
who began underground work between 1946 and 1972 andwere exposed to RDPs for at least 2 years. Tirmarche et al.
gave a status report on the epidemiological follow-up ofthese uranium miners in a 1985 paper [T9]. A first analysis of this cohort, based on follow-up to December 1985,
was published in 1993 [T6]. Compared with national rates,
significant excesses of deaths from lung cancer (observed =45, expected = 21) and cancer of the larynx (observed = 17,
expected = 7) were observed.
370. For lung cancer only, a linear dose–response relationship was described with respect to the cumulative exposure to RDPs [T6]. The ERR coefficient was relatively low(0.35 WLM–1) in comparison with those found in other minerstudies, but as the number of lung cancer deaths was also
low (n = 45), the CIs of this coefficient included most of thevalues of the other studies, as well as estimates from international committees [L10]. The authors cautioned that, since
the mean age of the cohort was only 56 years, the cohort was
too young for full expression of the lung cancer risk.
371. Laurier et al. described the results obtained from theinitial cohort after extension of the follow-up to 1994 [L44].
The mean age at study exit was then 63 years. Causes of death
were obtained from the National Mortality Database, whichcollects information from all death certificates in France.
Compared with previous analyses, the use of the National
Mortality Database as the principal source of information on
the causes of death allowed a reduction in the potential bias
in the calculation of standardized mortality ratios (SMRs).
The analysis showed, however, that this had little impact
on the relationship between RDP exposure and lung cancer
risk. Compared with the earlier study [T6], the number of
person-years was increased by 25% (n = 56,372) and thenumber of deaths by 74% (n = 612). The analysis confirmed
the existence of an excess risk of death from lung cancer
among French uranium miners (85 observed deaths, SMR= 1.9; 95% CI: 1.5, 2.3), and an increase of this risk withcumulative RDP exposure (ERR = 0.6 (95% CI: 0.1, 1.2) per
100 WLM). An excess risk of laryngeal cancer, noted in the1993 paper by Tirmarche et al. [T6], was not confirmed in
the later study (12 observed deaths, SMR = 1.1; 95% CI: 0.6,
1.9) [L44, T5].
372. The French cohort of uranium miners was enlarged
by the inclusion of additional miners with lower radon
exposures, and the follow-up was extended to December
1994. The enlarged cohort is described in references [L40,
R10, T30, T31]. The study was limited to males who had
been employed as miners by CEA-COGEMA for at least
1 year between 1946 and 1990. The main characteristicsof the enlarged French cohort of miners are summarized intable 20. This cohort comprised 5,098 miners followed-up
from 1946 to 1994, with a total of 133,521 person-years.
The percentage of miners lost in the follow-up was 2.3%.
The average age at the end of the study was 55 years. A total

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

of 1,162 deaths were observed. The cohort included a groupof 964 non-exposed individuals (but who had the status of aminer and were working in the same mining divisions as the271

miners). Among the exposed miners, the mean cumulative
exposure was 36.5 WLM, accumulated over a mean durationof exposure of 11.5 years.

Table 20 Characteristics of the enlarged French cohort of uranium miners [R10]Period of follow-up: 1946–1994
Number of workers

5 098

Non-exposed miners

964

Number of person-years

133 521

Average cumulative exposure (WLM)a36.5

Average annual exposure (WLM)
Before 1956

23.9

1956 and later

1.5

Average duration of exposure (years)a11.5

Person-years by lagged cumulative exposure (WLM)
049 408

0–10

35 817

10–50

27 77850–10011 358100–200

6 213

>200

2 947

a Among 4,134 exposed miners.

373. The total number of lung cancer deaths observed
between 1946 and 1994 among the enlarged cohort ofFrench uranium miners was 125. On the basis of the reference rates for the general French male population, thenumber of lung cancer deaths expected was 83.1. The analy­
sis confirmed a significant excess of lung cancer deaths(SMR = 1.51; 95% CI: 1.25, 1.79). A significant excess
was also observed for all cancer mortality (SMR = 1.14;
95% CI: 1.03, 1.25), but this disappeared after exclusion of
the lung cancer deaths (SMR = 1.02; 95% CI: 0.90, 1.10).
Thus no significant excess of deaths from any other cancer
was observed [T30].
374. Rogel et al. [R10] reported on how factors such as
time since exposure and exposure rate modified the lungcancer risk in the enlarged cohort of French uranium miners. The statistical analyses were based on a linear relative risk model using Poisson regression to fit the models,maximum-likelihood methods to estimate parameters andlikelihood ratio tests for nested models. A linear exposure–
response relationship with an ERR of 0.8 (95% CI: 0.3, 1.4;
p < 0.001) per 100 WLM was found. No inverse exposurerate effect was observed in the extended French cohort.
The strongest modifier was the period of exposure. Analysis showed an ERR that was 10 times higher per WLM for
exposures received before 1956 than for exposures received

in 1956 and thereafter. This could be explained by a ­better
quality of exposure assessment after 1956. The ERR for
exposures after 1956 was 2.4 (95% CI: 1.1, 4.6, p < 0.0001)
per 100 WLM [R10, T30].

5.

Evaluation

375. The assessment of RDP exposure among French uranium miners is of good quality. Beginning in 1956, monthly
records of individual exposure were systematically kept in
the mining division. Furthermore, compared with the initial cohort involved in the international joint study in 1994
[L10], the size of the cohort was increased almost 3-fold,
and the follow-up was extended by 10 years. Also, it is worth
noting that, after 1956, the exposures of the French miners to
RDPs were of the same order of magnitude as those received
in some homes. Therefore the potential contribution of theFrench uranium miner cohort data to the estimation of therisk coefficient for lung cancer could be especially relevantto the estimation of the risk for populations exposed to RDPs
in their homes. A European project that includes the French,
Czech and German cohort studies is in progress [T27, T30].
The combined data, which are of good quality and relate tominers with low levels of exposure, will allow an analysis tobe undertaken with a large statistical power.

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UNSCEAR 2006 REPORT: VOLUME II

376. The quantification of the relationship between cumulative RDP exposure and risk of lung cancer mortality required
elaborate statistical methods. Different models were applied,
and modelling was performed independently by different
researchers [L10, R10, T6, T27]. All analyses confirmed anincrease of lung cancer mortality with cumulative exposure.
In the framework of a European programme, data from the
French and Czech miner cohorts were made available toresearchers involved in biologically based modelling, withthe aim of comparing the approaches and results of the different models [T30]. The different biologically based solutions gave a reasonably good fit of the data and confirmed alinear increase of risk with cumulative exposure [B27, H32].
The comparison of the different approaches provided very
interesting discussions regarding the biological validity ofthe models, the means for testing various hypotheses about
the processes of radiation carcinogenesis, the selection of thebest fitting model, and the comparison of these biologically
based models with the empirical approach that uses a statistical model for describing the data [B37, H33, L42].
377. Data on yearly gamma and long-lived radioactive ore
dust exposure are also available since 1956 and 1959, respectively, for each miner of the enlarged cohort. The information will allow all three components of radiation exposure —RDPs, gamma radiation and long-lived radioactive dust — tobe considered in an analysis of the dose–response relationship. In addition, a nested case–control study investigating
the joint effect of RDP exposure and smoking on lung cancer
risk among French uranium miners was published [L44]. It
confirms the existence of a significant effect of RDP exposure when smoking information is taken into account.
I.

Canada: Newfoundland fluorspar miners
1.

stage, pumps could not cope with the amount of water in the
open cut, and it was necessary to sink a shaft, which eventually went to a depth of 250 feet [A14]. This mine ceased
operations in 1942. In 1937, the St. Lawrence Corporation
started work on a vein called Iron Springs. Originally, mining there was also by the open-cut method, but by 1938, thework had moved underground, with the mine eventuallydescending to a depth of 970 feet. This mine was closed in
December 1956.
381. Standard underground mining procedures wereadopted with the first underground mine in 1936. Wet drilling (which resulted in reduced dust levels) was generally
adopted in 1942. Shrinkage stoping and cut-and-fill methods
were not practised until after 1964. The underground mines
were in general very wet. Ventilation was mostly provided bynatural means. Except in one case, supplementary ­blowers
were not used until 1946 [D5].
382. The report of the Royal Commission [A14] notedthat the St. Lawrence Corporation had 16 veins of fluorspar
on its mining properties, and at one time or another mostof these were mined by the company. Practically all miners employed by the St. Lawrence Corporation who worked
underground prior to 1960 had at some point also spent timeworking in the Iron Springs mine. According to the report of
the Royal Commission [A14], the working conditions in theBlack Duck mine were unpleasant. The Commission noted,that prior to 1942, drilling was done with a dry hammer, andthat dust and smoke were always such that the driller could
only be seen at close quarters. “He was always like a snowman and also had to shut off his machine to clear out his eyes
and nostrils.” Clearly, the mines were poorly ventilated, andany radon brought into the mines with mine water would
likely remain in the mine for a long period of time, resulting
in high radon/radon decay product equilibrium conditions.

Introduction

378. The 1969 report of the Royal Commission Respecting
Radiation, Compensation, and Safety at the Fluorspar Mines,
St. Lawrence, Newfoundland [A14] provided a wealth ofinteresting historical data, and is the source of much of thefollowing information.
379. For many years, St. Lawrence, Newfoundland, wasan isolated fishing community. Shortly after the First World
War, the community was devastated by the drop in price ofsalt-cured fish (the main source of income) and, in 1929, allof the fishing equipment was destroyed by a tidal wave. Mining eventually took over as the principal occupation [D4].
Fluorspar, which is used in the production of steel, aluminium and high-octane gasoline, is the only mineral resource
known to be of economic quality in the St. Lawrence area.
380. The earliest mine to start operation at St. Lawrence
was the Black Duck mine, which belonged to the St. Lawrence Corporation. This mine opened in March 1933. Originally, mining was by open-cut methods. By 1937, however,
the open cut had reached a depth of about 90 feet. At this383. Until 1942, there were three shifts working underground at the Iron Springs mine. Late in 1942, the mine
was put on two shifts. This change was most beneficial: thefour hours between shifts could be used for blowing out
the smoke with compressed air. There was practically noforced ventilation in any of the mines before 1960, naturalventilation having been the only source of ventilation untilthat time.
384. Mines were inspected annually by inspectors brought
in by the Newfoundland government; however, it was not
until 1951 that mining regulations were in place. Consequently, prior to this date, the inspectors had to rely solely
on persuasion to achieve improvements in the conditions in
the mines.
385. Morrison et al. [M16, M17] noted that, although morethan 40 fluorspar veins were located in the St. Lawrencearea, most of the ore produced came from only two mines,
namely the Iron Springs and the Director mine, mined by
the St. Lawrence Corporation and the Newfoundland Fluorspar Corporation, respectively. According to Morrison et al.

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

[M16, M17], the Iron Springs mine was believed to have had
the worst ventilation of any mine in the St. Lawrence area.
In contrast, the Director mine, from 1955 onward, employedsome forced ventilation. Both mines were extremely wet.
386. Morrison et al. [M16, M17] postulated that, since
operations were converted to underground mining proceduresin the mid-1930s, in the middle of the Depression, work wasdone with antiquated equipment, and it was possible thatthe health and safety conditions were poor as a result of thefinancial situation at that time. From the mid-1950s, it was
clear that miners in St. Lawrence were suffering from various
respiratory troubles, some of which had been diagnosed assilicosis. For this reason, the Newfoundland Department ofMines requested the federal Department of National Healthand Welfare to carry out a survey of dust conditions in the
St. Lawrence mines. Subsequently, the Industrial Hygiene
Division of the Department of National Health and Welfare(now Health Canada) conducted a dust survey in the period
1956–1957. By the end of 1957, it was clear that the minersof St. Lawrence were suffering from a respiratory ailmentthat was not caused by the excessive quantities of siliceousdust. The dust survey was therefore expanded into a broader
epidemiological study. By the end of 1959, Windish andSanderson had completed two brief radiation surveys in themine [W6, W7]. The results of these surveys established thatairborne radioactivity in the form of radon and radon decay
products was present in the two mines surveyed in excessof the maximum permissible concentration (indicated in theRoyal Commission report to be 1 WL). One of the suggestions made by Windish [W5] was that radon was carried into
the mine by mine water and then released to mine air.
387. As described in Aylward et al. [A14], the Occupational Health Division of the Department of National
Health and Welfare began a detailed clinical investigationof the St. Lawrence miners in August 1960, under the direction of Dr. A.J. de Villiers. Current epidemiological studies
have evolved from this foundation. It should be noted that,prior to 1959, there were no measurements of radon orradon decay product concentrations in the ­Newfoundland
fluorspar mines.
2. Radon and radon decay products
388. On the basis of work done described in the report of
the Royal Commission [A14], the source of radon was eventually identified as the water that poured into the mines [D5],the radon itself apparently originating from the host granite.
The report of the Royal Commission ([A14], table II) reported
levels of radon of 300–13,000 pCi/L in mine water.
389. Interestingly, high radon levels in water were notlimited to the mine. According to data of the Royal Commission [A14], measurements of municipal water supplies
in the St. Lawrence area revealed radon levels ranging from1,800 to 14,370 pCi/L. Radon levels in other water supplies
in the St. Lawrence area ranged from 6,000 to 12,000 pCi/L.

273

Radon levels in neighbouring communities close to St. Lawrence were found to range from 1 to 1,140 pCi/L.
390. Morrison et al. [M17] noted that in 1960, becauseof the high levels of radon identified, mechanical ventilation was introduced into all levels of the mine that were still
operating, and the RDP levels subsequently fell below thethen current limit of 1 WL. In 1978, mining operations in
St. Lawrence ceased, and the last fluorspar mine was closed.
By this time, 78 cases of lung cancer had already been
­identified [C12].
391. Surveys conducted by Windish and Little in 1959 and1960 collected 17 radon and 80 RDP readings (the Kusnetz
method was used for measuring the RDP concentrations)and several gamma radiation readings (reported in reference[A14]).
392. A retrospective study of early mining conditions andworking level exposures to RDPs was carried out by Corkill
and Dory of the Atomic Energy Control Board (AECB)
of Canada [C12]. On the basis of a detailed study of eachmine, including measurement data, mining records, ventilation data, interviews and simulation studies, every mine was
assigned a high, medium or low RDP exposure level for each
operating year (1933–1960). No RDP measurement data
were available prior to 1959. Before 1960, when additional
ventilation and control measures were introduced, miners
were likely to have been exposed to average RDP concentrations of 2.5–10 WL, depending upon the type and place ofwork. From 1960 onward, estimates of RDP exposures wereavailable for miners by calendar year [M14].
3.

Exposure estimation

393. Sources of RDP exposure data include: values estimated and assigned to each mine and each calendar year for
1933–1960 in the retrospective study by Corkill and Dory
[C12]; survey data by Windish and Little in 1959 and 1960
(as reported in reference [A14]); and personal exposure data
starting in 1960. The data (80 samples collected at 50 different locations) in Windish and Little’s survey formed the basisof de Villier’s epidemiological studies in 1964 and 1971 [D4,D5]. In the 1964 study, mortality among fluorspar miners
was compared with normal mortality in the same geographical region and with that of uranium miners.
394. The WLM method of exposure measurement for individuals was not used until de Villier’s 1971 paper [D5]. In
the 1971 study, analyses were performed on miners, drillers
and muckers. Pre-1960 exposures were included, since themortality analysis started in 1933; however, no explanationof how they were accounted for is given. The only exposure
data mentioned were those given in the Windish survey.
395. The retrospective study of Corkill and Dory [C12]and the study of Morrison et al. [M14] provided more comprehensive data for miners employed during the period

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UNSCEAR 2006 REPORT: VOLUME II

1933–1978, including mortality data from 1933 to 1981. In
the analysis by Morrison et al. [M16, M17], the cohort consisted of 1,772 miners employed either by the St. Lawrence
Fluorspar Company or Newfoundland Fluorspar Limi­ted.
Morrison et al. [M16, M17] described the databases, personal identifying information and occupational histories.
The exposure estimates in WLM used in this analysis were
calculated on the basis of year, mine and occupation for theperiod 1933–1960, based on data provided by the AECB
[C12].
396. According to Dory and Corkill [D6], from 1961
onward, estimates of RDP concentrations in WL and/or RDP
exposures in WLM were available from the miners’ individual files. The environmental conditions in the early years ofmining were reviewed. Exposure re-evaluations were carriedout on the basis of an assessment of the environmental conditions determined by a review of mine maps, inspectors’
reports, Royal Commission hearings, anecdotal information from former workers and the authors’ own experience.
Comparison was also made with the conditions in the minesin later years for which radiation measurements were available. Dory and Corkill [D6] also used computer modellingto simulate expected work environment conditions for eachmine for each year of its operation. Most importantly, theauthors recognized the uncertainty in the estimation procedure and presented a range of RDP concentrations for various workplaces taking into account the degree of wetness
and the degree of ventilation assumed to be present. Corkill
and Dory [C12] noted that the ranges they presented do not
represent extreme concentrations but are “average workplaceconcentrations for high, medium and low areas”.
397. Corkill and Dory [C12] considered it possible to place
job types within certain average concentration ranges. Their
exposure re-assessment substantially improved the estimation of RDP concentration (WL) and RDP exposure (WLM)
for the Newfoundland fluorspar miners. Nevertheless, as in
all attempts at reconstruction, there are large (and probably
irreducible) uncertainties in the actual conditions that existed
in the workplaces for the period 1933–1960.
398. Detailed occupational histories were obtained fromcompany records. The records for the 1933–1936 period
were reconstructed from census data, interviews with company officials and company report reviews [D4].
399. Workers’ occupational histories, by accumulated
hours of exposure and type and place of work, were prepared for all men on the employee list [D5]. Occupational
history was particularly important for fluorspar miners, since
epidemiological studies included surface workers as well as
underground miners.
400. Corkill and Dory [C12] assumed that the type of job
dictated where a man worked, and thus the concentration of
RDPs to which a worker was exposed. For example, development miners were assigned the high average concentrations, stope miners the medium average concentrations, andminers working in an established area near an air circuit the
low average concentrations. The high and low averages often
differed considerably (by a factor of from 2.5 to 10). To use
these averages, good knowledge of each worker’s duties wasdeveloped.
401. No mention of the effect of job mobility was made in
Corkill and Dory’s study. This might have been taken into
account, at least to a certain degree, since the payroll records
used to construct the occupational history were recorded
every two weeks.
402. No mention was made of whether previous hard rockmining experience had been taken into account. However,since St. Lawrence was an isolated fishing community, especially in the early years, miners were most likely to havebeen fishermen before they started work in the St Lawrence
mines.
403. A working month was considered to be 167 hours by
de Villiers and Windish [D4] and 170 hours by Morrison etal. [M16, M17] in the WLM calculations. Accumulations ofworking hours were likely to have been reasonable, since
they were compiled from payroll records (on the basis thatminers were paid on an hourly basis).
4.

Epidemiological studies

404. Early studies of the St. Lawrence fluorspar miners
included those of de Villiers and Windish [D4], Parsons
et al. [P1] and Wright and Couves [W10]. However, the firstattempt to examine the RDP exposure–response relationship was given in a paper published in 1971 by de Villiers
et al. [D5]. The number of hours worked underground wasused as a surrogate for actual exposure to RDPs. A plot oflung cancer deaths versus number of hours worked revealedan ­exponential relationship. To adjust for the variableradiation exposure of differing occupations, hours worked
were weighted according to occupation. Drifters and stope­workers were assigned an RDP concentration of 8 WL,
muckers, trammers, and chute operators were assigned alevel of 4 WL and shaftmen were assigned a level of 2 WL.
With the assignment of these weights, the exposure–response
relationship became linear. This procedure was followed byMorrison et al. [M18], who extended the mortality followup first to 1978 [M18] and then to 1981 [M14]. In the latterstudy (published in 1985), the expected number of deaths
was calculated from data for an internal control group of
unexposed surface workers, with an attempt to account for
cigarette smoking and latency period. These analyses, likeprevious ones, used a modified person-years approach.
405. In 1988, Morrison et al. [M16, M17] modelled theexposure–response relationship using an external control
group, and estimated the attributable and additive relativerisk coefficients. The radon exposure estimates developed by
Corkill and Dory [C12] were used. Attributable and relativerisk coefficients were examined by attained age, age when

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

first exposed and smoking status. In addition, lifetime riskof lung cancer mortality was assessed using both the relative and the attributable risk model. The cohort consisted of
1,772 miners employed by either the St. Lawrence Fluorspar
Company or Newfoundland Fluorspar Limited. Morrison
et al. [M16, M17] described the databases, personal identifying information and occupational histories. Estimatesof exposure in WLM were calculated on the basis of year,
mine and occupation for the period 1933–1960, and the­calculations provided by the AECB [D6].
406. A 1995 cohort study of the Newfoundland fluorspar
miners by Morrison and Villeneuve [M20] and Morrison
et al. [M27] examined the mortality experience (1950–
1990) of 1,744 underground miners and 321 millers or surface workers. As in the 1988 study, exposure estimates in
WLM by year, mine and occupation for 1933–1960 were
provided by the AECB [C12]. RDP exposure during the
5 years preceding lung cancer was assumed to be unrelated to lung cancer risk. Overall, 60,000 person-years of­follow-up were noted, with a mean cumulative exposurefor underground workers of 382.8 WLM over an average of
5.7 years of exposure. Smoking information was available
for 65% of the exposed cohort.
407. Values of relative risk (for exposures estimated on thebasis of the category of work), adjusted for attained age andperiod since exposure, increased with cumulative exposure
and were statistically significant for cumulative exposuresexceeding 200 WLM. On the basis of Poisson regression of
the ERRs and the exposure estimates, and through the useof a constrained intercept, the ERR was estimated to be 0.66
per 100 WLM, with a standard error of 0.17%. The effect ofcell killing at high WLM was not statistically significant at
the p = 0.05 level. The attributable risk coefficient for continuous exposure was estimated to be 6.3 deaths (standard
error of 0.74) per working level month per 106 person-years,
with a multiplicative correction estimated from the cohort.
408. The ERR per unit exposure increased with duration of
exposure, suggesting that those exposed over longer periodsof time had a greater risk than those exposed over a shorter
period at the same exposure level.
409. The ERR per unit exposure decreased with increased
duration of exposure in a previous 11-cohort study by Lubin
et al. [L10]. While this is biologically plausible, theseauthors presented an alternative explanation for the inverse
dose-rate effect, namely that the finding is an artefact resulting from a greater non-differential exposure misclassification at higher exposure rates than at lower exposure rates.
Because exposure rates were extrapolated for the periodprior to 1960, when they were high, these authors suggested
that there may be a much higher degree of miscalculation ofthese exposures than of exposures from 1960 onward. The
effect of non-­differential misclassification would be to bias
the risk estimates towards lower values, resulting in a greater
reduction in risk estimates at high dose rates than at low dose
rates. If the inverse exposure-rate effect is indeed the result

275

of bias, then it follows that the ERR would have been anunderestimate, since the significant random exposure misclassification that constituted the basis for the bias should
also bias the overall ERR towards lower values at higher
exposure rates.
410. The 1995 analysis [M20] found an ERR for lung cancer of 0.66 per 100 WLM, slightly lower than that observedby Lubin et al. [L10] in their analysis of the fluorspar cohort.
The difference with the more recent analysis was that it was
based on a later follow-up time (1990 versus 1984 for reference [L10]). However, the ERR of 0.66 per 100 WLM
was similar to that noted by Lubin for all 11 mining cohort­studies combined (0.49 per 100 WLM). Although statistically significant differences in ERR were detected between
smokers, non-smokers and former smokers, the joint effects
of exposure to RDPs and smoking could not be assessed.
411. A more recent report described an 11-year updated
analysis of the mortality experience (1950–2001) of the
Newfoundland fluorspar miners [V4]. The new study reports
on an analysis of 328 miners who worked exclusively on
the surface and 1,742 individuals exposed to RDPs fromworking underground. When compared with Newfoundland
males, the fluorspar miners had significantly increased numbers of deaths for lung cancer, silicosis, and accidents, poisoning and violence. In total, 206 lung cancer deaths wereidentified, 191 of which occurred among individuals who
had at some stage worked underground, the other 15 occurring among miners who had worked only on the surface.
412. Villeneuve et al. [V4] found a strong association
between cumulative exposure to RDPs (WLM) and lungcancer risk. Workers with estimated cumulative exposuresexceeding 2,100 WLM had relative risks more than 20-foldhigher than unexposed miners. After adjusting for age andcalendar period, the linear ERR among underground and surface miners (combined) was estimated to be 0.47 (95% CI:
0.28, 0.65) per 100 WLM. The relationship between cumulative exposure in WLM and lung cancer risk was modified bytime since last exposure, duration of exposure and exposurerate. In contrast, age at first exposure was not a statisticallysignificant determinant of lung cancer risk. After 35 years
since the time of last exposure, lung cancer mortality rates
among exposed miners dropped to levels experienced bythose who worked exclusively on the surface. Morphology
was available for 88 of the 191 lung cancer deaths among
those who worked underground. The histology included
squamous cell carcinomas (28), adenocarcinomas (8), small
cell carcinomas (7) and other carcinomas (45). Owing to thesmall number of cases, it was not possible to determine the
ERR per unit exposure by histological type.
413. Twenty-eight lung cancer deaths occurred among men
who started working after 1960 (when ventilation was introduced). There was no significant variation in the ERR per
unit exposure between those who started work before andafter 1960. However, the evaluation of cancer risk among
those who started mining after 1960 is based on younger

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UNSCEAR 2006 REPORT: VOLUME II

men. Thus few cancer deaths were identified, so the statistical power to detect an association in these workers was­limited [V4].
414. Some data on smoking were available for 1,107 of
the 2,070 miners (53%). There was no statistically significant difference in the ERR per unit exposure between those
who had smoked and those who had never smoked. However, strong associations between cumulative radon exposure and lung cancer risk were noted among individuals who
had smoked different numbers of cigarettes daily (p < 0.05).
Specifically, the ERR was 0.31, 0.46 and 0.94 per 100 WLM
among individuals who reported smoking <15, 15–<30 and
30 or more cigarettes per day, respectively. An evaluation of
the joint effect of exposure to RDPs and smoking (as measured by the number of cigarettes smoked daily) could notadequately discriminate between additive and multiplicativemodels. However, the data were suggestive of an intermediate relationship (between additive and multiplicative). The
evaluation of the joint effects of smoking status and radon
was severely limited by the small number of lung cancer deaths that occurred among miners who never smoked(n = 8) [V4].
5.

Evaluation

415. One possible advantage of the Newfoundland fluor­
spar cohort over most other studies of radiation-exposedmining populations is that the source of RDP in the fluorspar mines was from groundwater and not from radioactive
ore. Thus it was possible to exclude the effects of gamma
radiation, thoron and radioactive dust. The cohort was, however, exposed to silica. Another advantage is that fluorspar­miners were almost without exception local men with noprevious mining experience. Upon ceasing to mine fluorspar,
the people went back to non-mining professions. Unfortunately, all the exposures in the period of high exposure rates(1933–1960) were estimates only and are subject to large
uncertainty. While the exposure data are weak, the availability of smoking histories and the ability to investigate theeffect of changes in individual smoking histories over time
are strengths of the cohort. The study of Villeneuve et al.
[V4] confirms the strong association between cumulative
exposure in WLM and lung cancer incidence.
J.

Chinese miners

416. Uranium prospecting and uranium mining started
between 1955 and 1958 in China, and routine monitoring ofradon was carried out after 1959. The first uranium mine wasestablished in Hunan province in August 1958; very highlevels of radon were measured because of poor ventilation.
Radon levels decreased rapidly after 1960. Comprehensive
studies of radon exposure were completed in 1993 for 11 uranium prospecting teams and four uranium mines. A total of27,172 and 108,744 person-years were accumulated for theprospecting teams and the miners, respectively, during thefollow-up period 1971–1985. Over this period, the average
RDP concentrations were 0.3 WL for the prospecting teams
and 1.0 WL for the miners, resulting in average cumulative
exposures of about 80 WLM for each group. In total, there
were 28 lung cancers. ERRs of 1.19 per 100 WLM and 1.09
per 100 WLM were estimated for the prospecting teams andthe miners, respectively [S52].
417. There are some English language papers (e.g. [L11,L27, S26, S27, S50, S51, S52, S53, S54, S58, T32, Z6])
and many Chinese papers that discuss the lung cancer experiences of miners who worked in the Yunnan tin mines in
China. Qiao et al. [Q1] reported on an investigation of risk
factors and the early detection of lung cancer in a cohortof Chinese tin miners. They described a dynamic cohort
using an ongoing lung cancer screening programme amongtin miners exposed to arsenic and RDPs. The investigation
noted that about 6,000 tin miners are screened per year withsputum cytology, chest X-rays and personal interviews. The
authors calculated SMRs and 95% CIs. They also calculated
relative risk and 95% CIs for lung cancer risk factors from a
proportional hazards model. Exposures to RDPs and arsenic
were the predominant risk factors, but silicon and smoking
were also lung cancer risk factors in their cohort.
418. Chen and Chen [C36] reported a nested case–­control
study of 130 male lung cancer cases and 627 controls from acohort of 7,855 miners employed for at least 1 year between
1972 and 1974 in any of four tin mines in China. The­Maentel-Haenzel OR was used to measure the association
between lung cancer and various risk factors. Unlike Qiao et
al. [Q1], these authors did not find that silica exposure wasrelated to the risk of lung cancer. However, the authors did
find a strong association between risk of lung cancer andcumulative exposure to dust, cumulative exposure to arsenic
and duration of dust exposure. The most recent English language summary of this cohort was provided in a paper by
Shiquan et al. [S54], which indicated that the high incidenceof lung cancer in miners of the Yunnan Tin Corporation(YTC) in Gejiu, South China, had attracted attention since
the early 1970s. Underground monitoring of dust started in
1955 and of RDPs in 1972. Data collection from medical
examinations, chemical analysis of pulmonary tissues andanimal experiments started in 1975. The database for epidemiological studies was established in 1976, and reports fromcooperative studies with the National Cancer Institute (NCI)
in the United States began in 1998. Some of these epidemiological studies and the database provided by YTC were usedby the National Institutes of Health [L10] and BEIR VI [C20]as the basis of data for the Chinese miners in the joint analysis of 11 underground miner cohorts. Lung cancer cases fromthe YTC, one of the largest of the 11 cohorts, made up 36% of
the total cases in the NIH and BEIR VI combined analyses.
419. About 90% of the lung cancer cases at YTC had ahistory of working underground [S54]. Prior to 1950, theprincipal work involved men carrying ore on their backs
in small tunnels. That working style was gradually abolished after 1953 as mechanization started to be introduced.

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACESYTC miners represented a stable population without significant loss of follow-up. Mose miners started work before
1950, some before 1920. Most miners with lung cancer began
mining as children under the age of 14, but this group’s ageof death and risk of lung cancer showed no prominent differences from those of miners who started mining after theage of 15 or 20.

combination of tobacco smoke and arsenic, 15.8% to arsenic
alone, 11% to a combination of arsenic and RDPs, 9.2% to a
combination of tobacco smoke and RDPs, 8.7% to a combination of arsenic, tobacco smoke and RDPs, 5.5% to RDPs
alone and 8.7% to background gamma radiation.
K.

420. The first systematic monitoring for radon was done infour YTC mines in 1972. The highest level was at the L–mine,at 28.6 Bq/L on the average. Radon concentrations decreased
after the improvement of underground ventilation in 1974
[S54]. To estimate RDP exposure before 1950, 13 existingsmall tunnels were measured in around 1980, and showed
an average RDP concentration of 2.3 WL. Measurementsof 210Pb in rib bones obtained after operations on miners
with lung cancer provided additional data for retrospective
dose assessment [L27]. Estimates of cumulative exposure to
RDPs could be divided into three stages: 1950–1953, whenthe miners carried ore on their backs; 1953–1972, a periodof modern mining but lacking radon monitoring; and after
1972, with radon monitoring.
421. Before 1965, airborne dust concentrations underground were very high at about 27–60 mg/m3, because ofdry drilling. In 1965, the dust levels decreased to 6 mg/m3
after the introduction of wet drilling, [L27, S58]. Arsenic
and iron were present as relatively insoluble compounds in
YTC mines; the arsenic concentration in the rock was about
0.5–1%. Airborne arsenic concentrations measured in the
1970s were about 0.01 mg/m3 and were thought to have been
more than 10 times higher in earlier years.
422. A combined effect of exposure to RDPs and arsenic
on the aetiology of lung cancer among the YTC miners wasreported previously [S51, X1, Y1]. The data for arsenic(dust containing arsenic) and RDPs, the only occupational
carcinogens underground, were compared to identify theirrelative contribution in the aetiology of lung cancer. Since
cumulative exposures to RDPs and arsenic were highly correlated, both being related to the duration of undergroundwork, comparison of the relative contributions to risk from
RDPs and arsenic had to be approached from differences inlung cancer risks in miners working at different jobs (mining, tunnelling, auxiliary) and different mines (L, M and S,
the three largest YTC mines).
423. Sun et al. [S53] suggested that the arsenic adjustment
used by the NIH [L4] was unsuitable for use in risk projection. BEIR VI [C20] noted that adjustment for arsenic exposure was difficult because of the strong correlation between
RDP and arsenic exposures. Hazelton et al. [H7] noted ahigh risk from arsenic exposure and an interaction of arsenic
with other sources of exposure in the study of the YTC miners. These authors analysed the arsenic, radon, cigarettesmoke and pipe smoke exposures using the biologically
based two-stage clonal expansion model. They concluded
that, of 842 lung cancer deaths among YTC miners in Gejiu,
21.4% were attributable to tobacco smoke alone, 19.7% to a

277

Australia: Radium Hill uranium miners

424. In addition to the studies discussed earlier in this section, other miner groups exposed to radon have been discussed in the literature. For example, the BEIR VI report[C20] also discussed lung cancer in workers in the Radium
Hill uranium mine in Australia [W15]. Exposures in this mine
were estimated on the basis of 721 measurements of radon
concentrations; however, no data on RDP concentrations
were reported. Ventilation data were used to estimate mean
residence time in the mine and subsequently the exposure ofminers in WLM. Overall exposures of this cohort were very
low, with a mean exposure of 7 WLM. The authors reported
an excess of lung cancer (4 lung cancer deaths) in miners
whose estimated cumulative exposures exceeded 40 WLM.
Thirty-six per cent of the cohort could not be traced beyond
employment at the mine. Overall, while supporting the findings of an association between RDP exposure and increased
lung cancer, this study provides no useable information for
the evaluation of an exposure–response relationship.
L. Overall evaluation of miner studies
425. All of the miner studies reviewed in this section
involve retrospective evaluation of exposures to RDPs. In
some cases, such as the Newfoundland fluorspar miners andthe Port Radium miners, almost all of the exposures were
estimated, whereas for the Czech and Wismut cohorts (starting after about 1971), relatively less speculation was needed.
Studies of miners also differ by type of mine (e.g. uranium,
iron, tin, fluorspar), exposure rate (i.e. WL in the workplace),
size of study (e.g. number of subjects, number of lung cancers) and other factors. All of the miner studies describedin previous sections confirm the risk of lung cancer fromexposure to RDPs. However, not all of the studies are of thesame “quality”. It is evident that the studies summarized intable 21 vary widely with respect to factors that affect thedetermination of the exposure–response relationship andfactors that modify that relationship, including, for example,
the number of excess lung cancers, the quality of the exposure data (both the range of exposures and the uncertainty in
exposures) and confounders such as smoking and exposure
to arsenic. A qualitative overall evaluation of such considerations is provided in table 21.
426. Table 2l is a summary of some of the key features ofthe various miner studies discussed in this section. The tablealso provides the average ERR per unit exposure estimatedfor a simple linear ERR model for each of the studies. The
ERR per unit exposure ranges over approximately a factorof 5. On the basis of an analysis of 11 miner cohorts, Lubin

278

UNSCEAR 2006 REPORT: VOLUME II

et al. [L10] reported an ERR of 0.49 (95% CI: 0.2, 1.0)
per 100 WLM.

using a generic inverse variance method and assuming that
random effects were present between studies. The results areillustrated in figure XIV, which shows a combined ERR of
0.59 (95% CI: 0.35, 1.0) per 100 WLM, comparable to thatreported by Lubin et al. [L10].

427. In a manner similar to that adopted by Lubin et al.
[L10], the estimated ERRs reported in table 21 were combined

Figure XIV. Estimates of ERR per unit exposure from studies of miners.
Combined ERR of 0.59 (95% Cl: 0.35, 1.0) per 100 WLM; 95% CI developed with random effects model.
Chinese tin [L10]
Port Radium [H35]
Colorado [H17, L10]
Newfoundland fluorspar [V4]
Ontario [K12, K13, L10]
French [R10, F30]
Swedish [R2, L10]
Beaverlodge [H35]
Czech [B9]Combined

0.1

0.2

0.5

1

2

5

10

EXCESS RELATIVE RISK (100 WLM)

-1428. It is of interest to compare the risks based on minerstudies with those from the more recent pooled residential
radon studies. Assuming for purposes of illustration a nominal indoor equilibrium factor of 0.4, 35 years of exposure
to indoor radon at 100 Bq/m3 for 7,000 h/a, an exposureof about 19.5 WLM can be estimated. If an ERR of 0.59
(95% CI: 0.35, 1.0) per 100 WLM based on miner studies
is assumed (see para. 427), an ERR of about 0.12 (95% CI:
0.04, 0.2) per 100 Bq/m3 can be estimated, a value which

compares remarkably well with values from pooled residential radon studies, as discussed in the next section. The
miner studies show that the ERR decreases with increasingtime since exposure. This is an important consideration in
evaluating potential risks from lifetime exposure. Finally, as
evidenced by the updates of the Port Radium and Beaverlodge cohorts [H35], continued follow-up of miner studies
is important, because the ERR and other outcomes of theepidemiological analyses may change as the cohort ages.

282

915

79

279

2 328

230

125

206

936

Ontario uranium­minersc[K12, K13, L10]

Czech uranium miners[T39]

Swedish iron miners[L10, R2]aBeaverlodge uranium
miners [H35]

Wismut uranium
miners [K27, K28, K29]

Port Radium uranium
miners [H35]

French miners[R10, T30]

Newfoundland
fluorspar miners [V4]]Chinese tin miners
[L10]

649

–

83.1

142.7

Not available

217.8

44.7

240.8

221

74

Expected cases

277.4

378

36.5

174.2

242

23.2

80.6

70

30.8

807.2

Mean cumulative
exposure
(WLM)

10

>35

26

>44

30.5

33

25.7

23.2

17.8

24.6

Mean follow-up time(years)135 357

70 894133 521111 222

1 801 626

285 964

32 452261 428319 701

75 032

Person-years

Exposure uncertainty;
smoking; arsenic; age
at exposure

Exposure uncertainty;
smoking; high dustlevelsExposure uncertainty
prior to 1956; smoking

0.16 (0.1, 0.2)

0.47(0.28, 0.65)

0.8 (0.3, 1.4)

0.37 (0.23, 0.59)

Exposure uncertaintyb;
smoking

0.96 (0.56, 1.56)

Exposure uncertaintyb;
smoking0.21 (0.18, 0.24)

0.95 (0.1, 4.1)

Exposure uncertaintyb;
smoking

Exposure uncertainty(especially prior to1966); smoking;arsenic; asbestos

1.6 (1.2, 2.2)

0.89 (0.5, 1.5)

0.42 (0.3, 0.7)

Average ERR per unit
exposure(per 100 WLM)
(95% CI in brackets)

Exposure uncertainty inearly years; smoking

Other hard rock mining;smoking

Exposure uncertainty;
other hard rock mining;smoking

Main potential
confounders

Low to medium

Low to medium

High

Low

Potentially high

High

Lowa

High

High

MediumOverall evaluation(subjective)

a Exposures in the Malmberget iron ore miner study have been updated, and miner doses have increased by about 50% from those used in reference [R2]; however, at this time the epidemiological study has not been updated.
b Uncertainty remaining after dose re-evaluation.
c The Ontario miner study is being updated and will more than double the number of person-years.

327

Colorado Plateau
uranium miners[H17, L10]

Observed cases

Excess relative risk of lung cancer for exposure to radon in mines

Study

Table 21

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES
279

V.

EPIDEMIOLOGICAL STUDIES OF RESIDENTIAL EXPOSURESA.

Introduction

therefore a matter that requires careful consideration, ideallyon a subject-by-subject basis. The pooled studies [D17, D21,
K1, K26, L26], based on individual data, are more informative than the previous meta-analyses.

429. Over the past twenty years or so, there has been agreat deal of interest in the risks arising from exposure toradon and its decay products. It is clear from studies of miners that exposure to radon and radon decay products causeslung cancer, as described in section IV (e.g. [C20, D14,
D16, I2, I5, L10, N11, U2, U5]). Data from animal experiments such as described in section III also demonstrate that
exposure to radon and its decay products causes lung cancer.
Until recently, data from studies of underground miners, inuranium mines and other mines, formed the basis for estimating risks from exposure to RDPs and for investigating
the exposure–response relationship, as, for example, wascarried out by BEIR VI in their pooled analysis of 11 minercohorts [C20]. Risks from residential RDP exposure were
estimated by extrapolation from miner studies. Now, however, there are more than 20 case–control studies of residential radon exposure and lung cancer. While individual
studies have limited power, pooled analyses of European[D17, D21], North American [K1, K26] and Chinese [L26]
residential radon exposure studies provide a clear demonstration of the risks of lung cancer from residential radon
exposure and a direct basis for estimating risk in dwellings
from such exposure. This section provides an overview of
the case–control residential radon exposure studies and, in
addition, a short commentary on the relevance of geographical correlation (“ecological”) studies. Further informationon the epidemiology of radon exposure and on “ecological”
studies is provided in annex A, “Epidemiological studies ofradiation and cancer”.
B.

431. Several ways to address the problems resulting fromsuch exposure assessment errors are discussed in the literature, including the use of special films placed on glass artefacts to measure the long-lived RDPs directly [L31]. One ofthe primary motivations for the glass-based measurements
was to take account of systematic increases in residential
radon concentration due, for example, to efforts to increase
the energy efficiency of homes by providing better insulation and decreasing air leakage. However, United Kingdom
data [L45] suggest that, in the United Kingdom at least, any
increase has not been substantial. Another motivation for theglass-based measurements was to eliminate the difficulties
caused by missing measurements in the residential studies,for example in cases where a house was demolished. Much
work in improving the retrospective assessment of radon
exposure was conducted using glass-based detectors andother approaches (e.g. [F9, L31, M33, P15]). Bochicchio
reviewed the use of nuclear track detectors in the context ofresidential radon epidemiology and also discussed the various sources of uncertainty in the retrospective estimation ofresidential exposures. Bochicchio suggests that 210Po alphaactivity on the surfaces of glass objects might be a bettersurrogate for past exposures than contemporary radon measurements. However, he noted that such data are confoundedby aerosols from cigarettes, which reduce the ratio of 210Po
surface activity to radon concentration [B42].
432. An early case–control study of domestic exposure to
RDPs and lung cancer in Port Hope, Ontario, was carriedout to determine if there was an excess of lung cancer in Port
Hope residents attributable to exposure to elevated RDP levels [L41]. Since smoking is the major cause of lung cancer,
the study controlled for smoking. “Cases” were defined as
any person who developed or died from lung cancer in the
period 1969–1979 and who had lived at least 7 years in Port
Hope. The 7-year residence period was selected because
this was the shortest possible time between exposure andthe occurrence of lung cancer. Twenty-seven cases met thecriteria. There were two controls for each case. Estimates of
exposure were developed by adding the cumulative exposures estimated for each house occupied by a case or control
since 1933. The statistical analysis failed to demonstrate anincreased risk of lung cancer from elevated domestic radon
exposure, but did identify a very strong risk of lung cancerfrom cigarette smoking.

Case–control studies of residential radon

430. Many case–control studies published prior to about
2000 are well described by Lubin and Boice [L4], BEIR
VI [C20], NCRP SC65 [N11] and the UNSCEAR 2000
Report [U2]. Lubin et al. [L8] discussed how errors in
exposure assessment can affect the interpretation of results.
These authors showed that information from seven case–control residential radon studies supported a wide rangeof risks, ranging from no excess risk to excess risks larger
than those predicted using data from miner studies. Theseauthors ­discussed various sources of error in the estimationof residential exposure, including various potential sourcesof measurement error. In addition, a number of authors (e.g.
[L4, P17]) have reported meta-analyses that are based on
published relative risks. However, such studies are not able
to correct for smoking, a key determinant of lung cancer and
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433. High concentrations of indoor thoron were observedin the Loess Plateau region of China (e.g. [S67, T35, W13,
W20, Y8]). The assessment of the risk from RDPs is known
to be affected by the presence of thoron and its decay products [P19, T17]. It was thought that the detectors used in
the study by Wang et al. [W13] might have been affected by
the presence of thoron [S10, T17] and thus that the measured radon concentrations might have been overestimated.
A re-assessment of the exposures from radon and thoron
decay products is currently under way. Moreover, the screentype diffusion battery (SDB) measurements in undergrounddwellings indicated the presence of ultrafine particles ofaround 10 nm [Y8]. Current evaluations are based on exposures; however, since the dose conversion factor for thesesmall particles is high, a modified contribution to dose mightneed to be considered in the future.
434. Several studies investigated residential radon exposureand lung cancer in China. A study carried out by Blot et al.
[B25] in Shenyang City, during 1988 and 1989, included
308 females with lung cancer and 356 female controls. The
median radon level measured in the homes of both cases andcontrols was 2.3 pCi/L (85.2 Bq/m3). The median durationof residence was 24 years. No link between radon exposureand an excess risk of lung cancer was found, irrespective ofsmoking status (other than a non-significant trend in heavy
smokers). A study by Wang et al. [W13] included 1,659 cases
and 768 controls who lived in an area of Gansu province.
Prior to 1976, many of the subjects had lived in undergrounddwellings (99%), although many had since moved to aboveground houses. The mean radon levels were quite high forboth cases and controls, at 230.4 Bq/m3 and 222.2 Bq/m3,
respectively. Using a linear model, the authors estimated an
excess odds ratio (EOR) of 0.19 (95% CI: 0.05, 0.47) at an
exposure of 100 Bq/m3. If adjustments were made for uncertainty in exposure, the EOR increased by about 50%.
435. Tokonami et al. [T35] and Sun et al. [S63] carried out
a radon and thoron survey in the Loess Plateau region ofChina. Their study area was located near Gansu province.
Since the geological features seemed to be almost the sameas in Gansu province, the characteristics of radon and thoron
concentrations were also likely to be similar. The radon concentration was lower than that in the study by Wang et al.
[W13], but the thoron concentration was higher. Because
thoron was underestimated in the past, reassessment of risks
due to radon exposure may need to take the presence ofthoron into account.
436. Field et al. described a residential radon study offemales in Iowa, United States [F10]. The Iowa radonstudy included 413 cases and 614 controls. The medianresidency was 33 years for cases and 31 years for controls.
Some 357 cases and 200 controls were “ever smokers”(i.e. people who had at some time smoked); an additional
104 cases and 200 controls were former smokers. The radon
exposure assessment involved a number of components,
including 1 year of on-site radon measurements, regional
outdoor radon measurements and linkage of the subject’s

mobility with exposure to radon indoors and outdoors. Outdoors, radon concentrations varied from 7.4 to 56 Bq/m3,
the ­latter being comparable to the United States national
average indoor radon level of 48 Bq/m3 (see figure I). An
average of four radon detectors were placed in each home.
The measurements exhibited approximately a log-normal
distribution. The majority of basement measurements and asignificant fraction of the measurements on the ground level
and the upper level exceeded 148 Bq/m3 (the United States
Environmental Protection Agency action level of 4 pCi/L).
The authors calculated cumulative exposures to RDPs
that occurred 5–19 years prior to diagnosis for the cases,or prior to time of interview for the controls, as 8.6 WLM
and 7.9 WLM, respectively. The odds ratios (ORs) for lungcancer in women who had smoked at least 100 cigarettes
or for at least 6 months in their lifetime relative to women
who had never smoked was 13.2 (95% CI: 9.5, 18.3). The
authors also found a significant positive trend between lungcancer and RDP exposure. The authors estimated risks fora cumulative 15-year radon exposure of 11 WLM (taken by
the authors as equivalent to an average radon concentrationof 4 pCi/L) for all cases. After adjustment for age, active
smoking and education, the authors estimated an EOR of
0.24 (95% CI: –0.05, 0.50) or (95% CI: 0.004, 1.81), whenradon exposure was treated as a continuous or a categorical variable, respectively. A subsequent paper by Field et al.
[F11] reported a slightly different CI for the EOR of 0.24;
this was 95% CI: 0.05, 0.92, calculated treating radon
exposure as a continuous variable. Field et al. observed astatistically significant trend for large cell carcinoma anda “suggestive” trend for squamous cell carcinoma (categorical p for the trend of 0.06). However, the linear excess odds
between different histological types were not significant.
437. Tomasek et al. [T10] reported a residential radonstudy of 12,000 people living in central Bohemia witha total of 173 lung cancers and a follow-up period of
1961–1995. This follow-up period was later extended to1999 [T29]. A total of 210 lung cancers were observed.
The study area in central Bohemia is mostly granitoid andhas radon levels considerably higher than other areas of
the Czech Republic. Exposure estimates were based on
measurements of the equilibrium-equivalent concentrations (EECs) of radon (i.e. the RDP concentrations) madein most (80%) of the homes in the study area. Typically,
two detectors were installed for 1 year in the two mostoccupied rooms. To compare their results with those ofother studies, the authors established a conversion factor
on the basis of 652 simultaneous measurements of EEC
and radon. Where necessary, mean values for a community
were used to replace missing data. The mean radon concentration was estimated at 509 Bq/m3, with 10% of the homeshaving indoor radon levels in excess of 1,000 Bq/m3. The
authors used a linear relative risk model, taking into accountthe exposures received between 5–34 years previously andestimating expected cases from national ­mortality data. The
authors estimated an ERR of 0.087 (90% CI: 0.017, 0.208)
per 100 Bq/m3. The ERR did not change substantially after
adjustment for smoking.

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438. Barros-Dios et al. [B26] described a populationbased case–control study in an area of north-west Spain
considered to be radon-prone. The study covered 163 cases(151 men and 12 women) and 241 controls (219 men and22 women). Radon concentrations were measured usingalpha track detectors placed in the homes for a minimumof 90 days and a median of 150 days. The mean radonlevels were 141.4 Bq/m3 for the cases and 114.0 Bq/m3 forthe controls. Overall, 22% of the homes had radon levelsof above 148 Bq/m3. A multiple logistic regression analysis assessed the risk of lung cancer, taking into account a
number of possible confounding factors, including smoking,family history of lung cancer, type of dwelling construction
and hours per day spent at home. A total of 145 (91.8%) of
the cases and 129 (54.7%) of the controls were smokers. The
authors reported ORs by quartiles of the radon distribution,and observed a greater than 2-fold increase in the risk oflung cancer for exposures to radon of above 37 Bq/m3. The
authors noted that the risk of lung cancer in smokers was
46 times higher than in non-smokers exposed to radon levelsof below 37 Bq/m3.
439. Lagarde et al. [L31, L32] reported on a Swedish residential radon study based on an existing database of persons who had never smoked that had been developed to
study environmental and occupational exposures to agentsother than radon. The database was augmented with measurements of radon concentrations made with alpha track
detectors that had been placed in the bedrooms and livingrooms of residences for 3 months. The database was also
supplemented with data from a nationwide Swedish case–control study for which the radon measurements were similar. On average, about 25 years of the 32-year residential
history of subjects was covered by measurements. Covariates included environmental tobacco smoke and history inoccupations with a risk of lung cancer. A total of 436 cases
and 1,649 controls (all of whom had never smoked) were
included in the risk assessment. The excess relative risk perunit of time-weighted residential radon concentration was
estimated using conditional logistic regression and a linear
relative risk model. The authors found a trend in relative risk
with increasing radon exposure, comparable to that found in
the nationwide Swedish study. The trend of increased risk
with increasing radon levels was limited to subjects exposed
to environmental tobacco smoke at home. For those who had
never smoked, a relative risk of 1.10 (95% CI: 0.96, 1.38)
was estimated, and for those exposed to passive smoking, therelative risk was estimated at 1.29 (95% CI: 0.97, 2.24).
440. Sobue et al. [S34] reported a case–control study ofresidential radon levels and risk of lung cancer in Misasa,
Japan. The case series consisted of 28 lung cancer deaths(26 males and 2 females) between 1976 and 1996, and 36(33 males and 3 females) controls chosen randomly fromresidents. Radon levels were measured using alpha track
detectors in the most frequented areas of the subjects’ homes
for a period of 1 year. The average radon level in the study
area was about 50 Bq/m3. The residential radon value measured over the year was used as a surrogate for cumulative

283

radon exposure over the 20 years of the study. None of theORs calculated using regression was statistically significant.
The authors attributed this in part to the small sample size.
441. The risks of lung cancer from residential exposure toradon in Devon and Cornwall in the south-west of the United
Kingdom were reported by Darby et al. [D15]. The study
looked at 982 cases of lung cancer and 3,185 controls, allunder 75 years of age. Detailed information on the demographics of the study population, smoking characteristics andresidency was provided. The investigators attempted to measure radon levels at all of the addresses where the subjects had
lived in the preceding 30 years. Two alpha track radon detectors produced by the United Kingdom National RadiologicalProtection Board were installed (one was placed in the living
area and the other in the bedroom) for a period of 6 months.
For residences in which radon had been measured, the investigators calculated a weighted average radon concentration
assuming that residents spent 45% of their time in the living
area and 55% of their time in the bedroom. A time-weighted
average radon concentration was estimated for the 30-year
period of interest using the times spent at each address as
the weights. Estimates of the lung cancer ERR per unit concentration in air were obtained using linear logistic regression on the assumption that the subjects’ radon exposure was acontinuous variable. The regressions also included interactions
with the factors noted earlier and other variables. The meanseasonally adjusted radon level in the 9,448 residences was
58 Bq/m3, with a maximum of 3,549 Bq/m3. The ERR for lungcancer risk was 0.08 (95% CI: –0.03, 0.20) per 100 Bq/m3 after
adjustments for age, sex, smoking status, county of residence
and social class. After further adjustment for the uncertainty
in estimates of exposure, the authors reported an ERR of 0.12
(95% CI: –0.05, 0.33) per 100 Bq/m3. The observed variations
among different tumour types were no larger than would be
expected by chance.
442. A case–control study for the period 1990–1996 inwestern Germany was carried out by Kreienbrock et al.
[K18]. Detailed demographics and information on potential
confounding factors (smoking and occupational asbestos
exposure) were collected, as was detailed information on residences occupied in the previous 35 years. Radon measurements were made over 1 year using nuclear track detectors
exposed in the living rooms and bedrooms of participants’current and previous homes. (Radon exposure was quantified in two ways: a time-weighted average of the living roomand bedroom radon concentrations in the last residence;
and an estimation of the time-weighted average cumulative
radon exposure in the 5–15 years prior to the interview date.)
An attempt was made to correct for home alterations andchanges to home ventilation. There were 1,449 cases and2,297 controls for the entire study area, with a subgroup in aradon-prone area consisting of 365 cases and 595 controls.
In the overall study area, women constituted 235 of the cases
and 432 of the controls. Among the men, 2% of the cases and
23% of the controls had never smoked. Among the women,
31% of the cases and 60% of the controls had never smoked.
Occupational exposure to asbestos was identified for men

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only (30.6% of the cases and 19.8% of the controls). Rate
ratios and 95% CIs were calculated using logistic regression.
All ORs were adjusted for age, sex, smoking and occupational asbestos exposure. The authors noted two results. In
the entire study area, no rate ratios were significantly different from unity, while in the radon-prone areas, for an
increase in radon concentration of 100 Bq/m3, an ERR of
0.13 (95% CI: –0.12, 0.46) was obtained for the exposure
assessment based on the last residence only, and an ERR of
0.09 (95% CI: –0.14, 0.38) was obtained for the assessment
based on cumulative exposure in the 5–15 years prior to theinterview date. After conducting sensitivity analyses, theauthors attributed the absence of an observable risk in theradon-prone areas to inaccuracy of radon exposure assessment. The inaccuracy increases if the variation in exposure
within the study population is low, which was the case in theentire study area.
443. Kreuzer et al. [K17] reported on a study of residential radon concentrations and lung cancer in Saxony andThuringia, which are areas in eastern Germany with naturally elevated radon levels. The study included 1,192 cases
and 1,640 controls. Alpha track radon detectors were placedfor 1 year in the living room and bedrooms of the subjects’ homes. Radon exposure was calculated as the timeweighted average of radon concentrations in each room.
The authors also calculated a time-weighted average radon
concentration for the whole of the study period, which was
5–35 years prior to the interview date. Mean radon concentrations were 76 Bq/m3 among cases and 74 Bq/m3 among
controls. Detailed information on demographics and potential confounding factors was obtained using a standardized
questionnaire similar to the one used in the study by Kreienbrock et al. [K18]. Approximately 12% of cases and 14%
of controls were women. Among the men, only 2% of thecases and 26% of the controls had never smoked. Amongthe women, 51% of the cases and 77% of the controls had
never smoked. ORs and 95% CIs were calculated using conditional logistic regression (for example, smoking and work
history with asbestos exposure were quantified). The OR for
smokers compared with those who had never smoked was 18(95% CI: 12, 29) for men and 2.8 (95% CI: 1.70, 4.8) among
women. For men, about 30% of cases and 28% of controls
had occupational asbestos exposure. Asbestos exposure in
women was negligible. ORs adjusted for age, sex, smoking
and occupational asbestos exposure, as well as 95% CIs were
calculated using conditional logistic regression. Overall, an
ERR of 0.08 (95% CI: –0.03, 0.20) per 100 Bq/m3 and 0.09
(95% CI: –0.06, 0.27) per 100 Bq/m3 was found for subjects with complete measurements for all 30 years. Smoking
acted as a negative confounder, and there was a moderateincrease in lung cancer, which for small cell cancers was
pronounced.
444. Conrady et al. [C9, C32] conducted a residential radonstudy among females living in the Schneeberg and Schlema
areas of Saxony in eastern Germany. This study looked at all
female lung cancer cases in the study area between 1 January 1952 and 31 December 1989. The final group included

72 cases and 288 controls. About 78% of cases and 94% of
controls were non-smokers. Radon measurements were madeusing alpha track detectors. For 24 houses in Schneeberg, theauthors compared the radon data with measurements of 210Po
in glass samples, which were used for backward extrapolation of the radon concentrations, and concluded that indoor
exposure conditions had been “stable” over the study period.
They also noted a large variation in indoor radon levels by
week and by season, and cautioned that, while 1-year measurements will yield “quite constant” values, shorter-term
measurements could be misleading. The authors reported
radon levels of 730 Bq/m3 in case homes in Schneeberg,
about 540 Bq/m3 in “register” controls and about 290 Bq/m3
in “hospital” controls. Logistic regression was used in therisk analysis. Elevated odds ratios (OR = 4.35; 95% CI: 1.47,
12.90 and OR = 1.94; 95% CI: 0.59, 6.33) were observedin the two highest exposure categories (radon concentrations of 1,000–1,500 Bq/m3 and >1,500 Bq/m3) comparedwith the reference category of 50 Bq/m3. No elevation inrisk was evident in the lower exposure categories, but thismight be attributed to the low statistical power. When data
were restricted to non-smokers and lung cancer cases withhistological confirmation (38 cases, 172 controls), the risk
remained elevated in the highest exposure category, with aradon concentration of >1500 Bq/m3, showing a borderline
significant trend only among small cell lung cancers.
445. Pisa et al. [P6] carried out a residential radon study
in an alpine valley in Italy where residential radon levels
averaged 132 Bq/m3, compared with the Italian national
average of 77 Bq/m3. “Ecological” evidence suggested
the possibility of a weak association between lung cancer
and residential radon. Measurements of residential radonlevels were carried out over the course of 1 year using alphatrack detectors in the bedrooms of the most recent residence
for each subject. Between 1 January 1987 and 31 December 1993, 224 residents in the study area died from lungcancer. Of these, interviews were completed for 138 cases(122 men and 16 women) and 291 controls matched for sex
and year of birth. Lifetime smoking, dietary variables andoccupational history were all considered in the statisticalanalysis, in which multiple unconditional logistic regression, with radon treated as a continuous variable, was used
to estimate the OR and its 95% CI. The OR showed a strong
association with smoking in both males and females. No
association was observed between lung cancer and exposure to potential occupational carcinogens. An association
between radon and lung cancer was seen in men only. Men
who lived in homes with radon levels of 40–199 Bq/m3(mean 80.4 Bq/m3) showed an OR, adjusted for age, sex andsmoking, that was approximately a factor of 2 greater thanthat of men who lived in homes with radon levels of below
40 Bq/m3. The authors reported an OR of 1.4 (95% CI: 0.3,
6.6) per 100 Bq/m3 for the group as a whole.
446. Oberaigner et al. [O2, S59] carried out a residential
radon study in the highly radon-prone district of Imst in Tyrol,
Austria. Lung cancer deaths and a sample of deaths withcauses other than lung cancer during the years 1970–1992

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(causes of death highly related to smoking were excluded)
were matched by age, sex and year of death. The next of kinof the cases and the controls were interviewed for residential
history, smoking and other risk factors. Radon concentrations were measured in the last residence for approximately
1 year by means of alpha track detectors placed in the bedrooms and the living rooms. The study included 194 cases
and 198 controls. The percentage of men was 88% among
both cases and controls. Measurements covered 68% (cases)
and 75% (controls) for the period 5–35 years before death.
The mean radon concentrations were 266 Bq/m3 among thecases and 123 Bq/m3 among the controls. ORs were estimated using conditional regression models adjusted forsmoking and occupation. The ERR at an increased radon
exposure of 100 Bq/m3 was 0.25 (95% CI: 0.08, 0.43).
447. Wichmann et al. [S61, W19] reported the results of apooled analysis of the two German radon studies. These twocase–control studies were performed during 1990–1997 ineastern and in western Germany [K17, K18], with identical
study design. The original data were extended and pooled,
and included a total of 2,963 incidences of lung cancer and
4,232 population controls. Radon measurements were carriedout over the course of 1 year in houses occupied during the
5–35 years prior to the interview date. Conditional logistic
and linear relative risk regression was used for the analysis.
Measurements showed an average radon exposure of 61 Bq/
m3. The smoking- and asbestos-adjusted ORs were 0.97 (95%
CI: 0.85, 1.11) for radon concentrations of 50–80 Bq/m3,
1.06 (95% CI: 0.87, 1.30) for radon concentrations of
80–140 Bq/m3 and 1.40 (95% CI: 1.03, 1.89) for radon concentrations of above 140 Bq/m3, compared with the reference
category with radon concentrations of <50 Bq/m3. The linear increase in the OR was 0.10 (95% CI: –0.02, 0.30) per
100 Bq/m3 for all subjects and 0.14 (95% CI: –0.03, 0.55)
per 100 Bq/m3 for less mobile subjects who had lived in only
one home in the previous 5–35 years. The risk coefficients
generally were higher when measurement error in the radon
concentrations was reduced by restricting the population to
those for whom good measurements had been made. With
respect to histo­pathology, the risk for small cell carcinoma
was much higher than for other subtypes.
448. Bochicchio et al. [B43] reported on a case–control
study of lung cancer and residential radon in Lazio, central
Italy, characterized by high levels of indoor radon and by
Mediterranean climate and diet. All subjects — 384 cases
and 404 controls, aged 35–90 years — were recruited in thehospital. Detailed information regarding smoking, diet andother risk factors were collected by direct interview. Residential history during the 30-year period ending 5 years before
enrolment was ascertained. In each dwelling, radon detectors
were placed in both the main bedroom and the living room
for two consecutive 6-month periods. A quality assurance
programme was set up for radon measurements [B41]. ORs
and 95% CIs for time-weighted radon concentrations werecomputed using both categorical and continuous unconditional logistic regression analysis and adjusting for smoking,diet and other variables. Approximately 89% and 91% of the285cases and the controls, respectively, were concluded to havegood radon exposure data. The adjusted ORs were 1.30 (1.03,
1.64), 1.48 (1.08, 2.02), 1.49 (0.82, 2.71) and 2.89 (0.45,
18.6) for radon concentrations in the range 50–99, 100–
199, 200–399 and above 400 Bq/m3, respectively, comparedwith the reference category with radon concentrations of
0–49 Bq/m3 (OR = 1; 0.56, 1.79). The adjusted odds ratio riskfor 100 Bq/m3 was 0.14 (–0.11, 0.46) for all subjects, 0.24(–0.09, 0.70) for subjects with complete radon measurements
and 0.30 (–0.08, 0.82) for subjects who had lived in no morethan one or two dwellings. There was a tendency towards
higher risk among subjects with low to medium consumptionof dietary antioxidants (EOR = 0.32; –0.19, 1.16). In conclusion, both ­categorical and continuous analyses clearly support an association between residential radon concentrationand lung cancer. Moreover, subjects with a presumed lower
uncertainty in the concentration assessment showed a higher
risk. Finally, this is the first study indicating that dietary­antioxidants may act as an effect modifier for radon.
449. Baysson et al. reported on indoor radon concentrationand lung cancer in France [B33]. The study took place from1992 to 1998 in French districts with elevated concentrationsof radon. The study included 486 lung cancers cases recruitedfrom university hospitals, and 984 controls. Subjects were
eligible if they had lived in the study area for at least 25of the previous 35 years. Full residential histories covering
the previous 30 years were obtained, with specific information collected for the dwellings that had been occupied
for more than 1 year. Radon concentrations were measured
using track etch detectors placed for a 6-month period in thecurrent and the former residences. The authors estimated atime-weighted average radon concentration for each subject
for the period of 5–30 years prior to the interviews. Finally,
the authors considered the effect of smoking and occupational exposure to carcinogens. The ORs and 95% CIs were
calculated using logistic regression with adjustments for sex,age, region, smoking and occupational exposure. Adjusted
odds ratio risks for 100 Bq/m3 were estimated for all subjects
as OR = 1.04 (95% CI: 0.99, 1.11) and for subjects withcomplete measurements (850 subjects) as OR = 1.07 (95%
CI: 1.0, 1.14). The French study supports a small excess riskof lung cancer arising from exposure to residential radon.
C. Ecological studies of residential radon450. The limitations of ecological epidemiological analyses are well discussed in references [C20, L8, L28, U2, U5],
as well as in annex A, “Epidemiological studies of radiation and cancer”. Nonetheless, a few short comments on
ecological versus case–control studies are appropriate here.
Radon studies that rely on averages over geographical areas
are especially vulnerable to biases that are not present inresults based on individual-level data such as those used incase–control or cohort studies. This is because radon levels
are highly variable even within limited geographical areas.
In addition, smoking is the major cause of lung cancer, andrisk is highly dependent on individual smoking habits. The

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distribution of smoking habits across the population in aparticular geographical area is an important confounding
factor.
451. Stidley and Samet [S49] reviewed 15 ecological studies of residential radon exposure from a number of countries.
A positive association between radon concentration and lungcancer was observed in 7 of the studies, no association wasobserved in 6 and statistically significant inverse relationships were observed in 2. The authors discussed methodological issues associated with the ecological studies and suggested that ecological studies should be given little weight inassessing the potential risks of residential radon. In a subsequent paper [S3], the authors showed by simulation that evenmodest levels of error in exposure or misspecification of therisk model could introduce significant biases into the resultsof ecological studies. As discussed in section B, the pooled
analysis of case–control studies in Europe [D17] and North
America [K1] have gone to considerable effort to correct for
uncertainty in radon concentrations. Further discussion ofthis limitation of ecological studies is provided in annex A.
452. Cohen used ecological epidemiological studies toinvestigate the linear no-threshold theory (LNT) by looking

at whether (county average) lung cancer rates in UnitedStates counties decreased with increasing (county average) radon concentration. Cohen’s work, involving 275,000
measurements in all 50 states [C23, C24, C25, C26, C27,
C28], generated a great deal of discussion, including references [A21, F8, F12, G15, G16, L24, L28, L29, L30, S48].
The focus of the discussion was the use of ecological rather
than analytical studies to investigate the potential risks
from residential radon. Puskin [P10] provided a plausible
explanation of Cohen’s observation of a negative association (see figure XV) between lung cancer and residential
radon concentration. However, Cohen points out that, while
an ecological study such as that given in reference [C24]cannot determine a risk versus exposure relationship, it cantest the LNT theory. Puskin [P10] found that the inverse
association between lung cancer and residential radon was
also seen in other smoking-related cancers not related toradon exposure. The result suggested that Cohen’s observations could largely be explained by a negative correlation
between smoking and radon exposure [P10]. On the other
hand, Cohen [C37] argued that Puskin’s observation ofsimilar dependence on radon exposure for lung cancer andfor other smoking-related cancers is not affected by data on
smoking prevalences.

Figure XV. Risk estimates of lung cancer from exposure to radon (adapted from reference [L4]).
Shown are the summary relative risks from meta-analysis of eight indoor radon studies and from the pooled analysis of underground minerstudies, restricted to RDP exposures of less than 50 WLM [L19], together with the estimated linear relative risk from the correlation study by
Cohen [C24]. Note that references [L4] and [C20] both show Cohen’s data extrapolated to beyond where they were actually analysed, about
200 Bq/m3. The figure has been adjusted to reflect this.
Relative risks from:
Indoor studies (case-control)
Miner studies (cohort)5

Ecological study by Cohen

RELATIVE RISK

Log-linear fit to indoor
data with 95% CI

1Relative risk = 1

0.3
0

100

200

300

RADON CONCENTRATION (Bq/m�)400

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

453. Heath et al. [H29] discussed exposure to residential
radon and lung cancer risk, as well as providing commentary on Cohen’s county-based studies [C24, C25, C35]. To
understand Cohen’s work, the authors obtained the data sets
on which his study was based and carried out an independent
analysis of lung cancer mortality on a subset of the data (lungcancer mortality for 1970–1979). Figure XVI (adapted fromreference [H29]) shows the results of the analysis of lung cancer mortality in relation to (ecological) radon levels but uncorrected for smoking (solid circles). The same figure shows thepattern of smoking frequencies superimposed (solid squares).
The smoking frequencies shown in figure XVI paralleled thepattern of lung cancer except at the lowest radon levels, and287

this led the authors to suggest that confounding by smoking
is particularly important at low levels of radon. Cohen argued
that figure XVI is meaningless because the effects of smoking
prevalences were fully taken into account in his analyses, andthese smoking prevalences were derived from three independent sources, all giving the same results. He also argued that noremotely plausible correlations between radon concentrations
and smoking prevalences or intensities of smoking could substantially change his results. Nonetheless, Heath et al. [H29]suggested that Cohen’s ecological studies are of limited value
in assessing the risk from residential radon exposure becauseof their reliance on grouped data, coupled with confoundingby cigarette smoking.

100

100

90

90

80

80

70

70

60

60

50

50

40

40

30

30

20

20

10

10

0

SMOKING PREVALENCE (%)

MORTALITY RATE IN MALES DUE TO LUNG CANCER (10-5 a-1)

Figure XVI. Average annual lung cancer mortality per 100,000 males, 1970–1979 (), and average percentage of cigarettesmokers among males () within counties, grouped by average county radon level.
Confidence intervals are expressed as the standard deviation of the distribution for each county group (adapted from [H29]).

0
0

50

100

150

200

250

AVERAGE RADON CONCENTRATION (Bq/m3)

454. Large ecological studies such as Cohen’s [C24, C25],
which show an inverse relationship between county averages of radon exposure and lung cancer mortality, continue
to generate controversy, despite the results of the numerous
analytical (case–control) studies such as those discussed in
the preceding section. As discussed in references [S3, S49]
and in annex A, ecological studies have a number of methodological challenges, notably their inability to adjust formobility and smoking habits. The pooled studies involved
control of both individual exposures and smoking habits,

and provide a methodologically sound basis for estimating
the risks from residential radon.
D. Overall evaluation of residential radon studies
455. Baysson and Tirmarche [B29] provided an overview
of the case–control studies of residential radon concentrationand lung cancer carried out since 1990. They suggested thatthe results of these studies indicated a positive association

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UNSCEAR 2006 REPORT: VOLUME II

between lung cancer risk and residential radon concentration
with, an EOR of 0.06–0.09 per 100 Bq/m3. Table 22 and figure XVII summarize much of the currently available data on
the risks per unit concentration in air from residential case–control studies. In considering the residential case–control
studies summarized above, it is important to understandthat some studies have greater statistical power than others,
owing to factors such as study size and data quality. The key
residential case–control studies were included in the pooled
analyses of Darby et al. [D17, D21] and Krewski et al. [K1,K26], as previously discussed. Uncertainty in exposure
assessment is a factor in residential radon studies, as are theeffects of smoking (both active and passive). However, when
the results of individual studies are pooled and analysed in aconsistent format, there is a coherent trend of increasing riskof lung cancer with increased radon exposure.
456. Bochicchio [B42] reviewed the application of solidstate nuclear detectors in residential radon studies. He argued
that exposure uncertainty is generally non-­differential andtherefore not only adds to the uncertainty in estimatedrisks but also introduces a bias towards underestimationof risks. Among the factors discussed by Bochicchio are:seasonal variability, which can vary from house to housedepending on the source of the radon; the characteristicsof the house and of people’s living habits; and the difficulties in retrospective dosimetry. Most of the radon studies
reviewed by ­Bochicchio covered periods of up to 35 years,
during which there may have been changes to the house,

the house ventilation system and residents’ habits, all ofwhich can affect both the levels of radon and the levels of
RDP exposure. Further discussion of the effects of measurement error in assessing radon exposure is provided in references [H37, H41]. As noted previously, the pooled studies in Europe [D17, D21], North America [K1, K26] andChina [L26] have attempted to account for measurement
error ­uncertainty (uncertainty in the measurement of radon
concentration and exposure) by restricting the assessment
to individuals who had lived in only one or two homes. In
addition, as noted earlier, Darby et al. attempted a statisticalcorrection for measurement uncertainty [D21].
457. Bochicchio et al. also provided a convenient summaryof residential case–control studies [B44]. The OR per unit
concentration in air and 95% CIs for each study are shownin figure XVII. Also shown in figure XVII are the results ofseveral meta-analyses and three pooled analyses in Europe,
North America and China. These include the Lubin and Boicemeta-analysis of the residential studies [L4], Lubin’s analysis
of North American and Chinese residential studies [L36], themeta-analysis of 17 case–control studies by Pavia et al. [P17],
the Wichmann et al. pooled study of residential exposure in
Germany [W19], two pooled studies of 13 case–control studies in Europe by Darby et al. [D17, D21], and the pooled study
by Krewski et al. [K1, K26] of 47 case–control ­studies in North
America. All of the studies included in the meta-analyses by
Wichmann et al. [W19] and Lubin and Boice [L4] are also
included in the pooled analyses [D17, K1].

Table 22 Indoor radon case–control studies (adapted from reference [B29])
Reference

RegionPopulation

Number of cases(controls)Radonmeasurements

OR a

95% CIYear

Author

1990

Schoenberg et al. [S42]

New Jersey,
United States

Women

480 (442)

1 year

1.49

0.89, 1.89

1990

Blot et al. [B25]

Shenyang, China

Women

308 (356)

1 year0.95

Undefined 1.08 g

1992

Pershagen et al. [P16]

Stockholm,
SwedenWomen

201 (378)

1 year

1.16

0.89, 1.92

1994

Pershagen et al. [P11]

SwedenBoth sexes

1 281 (2 576)

3 months

1.10

1.01, 1.22

1994

Letourneau et al. [L25]

Winnipeg, Canada

Both sexes

738 (738)

1 year0.98

0.87, 1.27

1994

Alavanja et al. [A19]

Missouri,
United States

Women,
non-smokers

538 (1183)

1 year

1.08

0.95, 1.24

1996

Auvinen et al. [A5]

FinlandBoth sexes

517 (517)

1 year

1.11

0.94, 1.31

1996

Ruosteenoja et al. [R3]Southern FinlandMen

164 (331)

2 months

1.80

0.90, 3.50

1997

Lagarde et al. [L1]

SwedenBoth sexes

1 281 (2 576)

1.17b

1.03, 1.37

1998

Darby et al. [D15]South-west
United Kingdom

Both sexes

982 (3 185)

6 months

1.08
1.12b

0.97, 1.20
0.95, 1.33

1999

Alavanja et al. [A18]

Missouri,
United States

Women

247 (299)372 (471)

1 year

0.85c
1.63d

0.73, 1.00
1.07, 2.93

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

Reference

RegionPopulation

289

Number of cases(controls)Radonmeasurements

OR a

95% CIYear

Author

2000

Field et al. [F11]

Iowa,
United States

Women

413 (614)

1 year

1.24

0.95, 1.92

2001

Kreienbrock et al. [K18]Western Germany

Both sexes

1 449 (2 297)

1 year

0.97 e1.09 f0.82, 1.14
0.86, 1.38

2001

Pisa et al. [P6]

Trentino, Italy–

138 (291)

1 year1.40

0.3, 6.6

2001

Lagarde et al. [L32]

Sweden436 (1649)

3 months

1.10

0.96, 1.38

2001

Tomasek et al. [T10,
T29]

Pluton,Czech Republic

210

1 year

1.087

1.017, 1.208

2002

Wang et al. [W13]Gansu, China

Both sexes

768 (1 659)

1 year

1.19

1.05, 1.47

2002

Barros-Dios et al. [B26]

Spain

–

163 (241)

150 days

2.48

1.29, 6.79

2002

Lagarde et al. [L31]

SwedenNon-smokers

110 (231)

3 months

1.33c
1.75d

0.88, 3.0
0.96, 5.30

2002

Oberaigner et al. [O2];Schaffrath et al. [S59]

Tyrol, Austria

Both sexes

194 (198)

1 year

1.25

1.08, 1.43

2003

Kreuzer et al. [K17]

Eastern Germany

Both sexes

1 192 (1 640)

1 year

1.08

0.97, 1.20

2004

Wichmann et al. [S61,W19]

Eastern andwestern Germany

Both sexes

2 963 (4 232)

1 year

1.1

0.98, 1.3

2004

Baysson et al. [B33]

FranceBoth sexes

486 (984)

6 months

1.04

0.99, 1.11

2005

Bochiccio et al. [B43]

Lazio, Italy

Both sexes

384 (404)

6 + 6 months

1.14

0.89, 1.46

Non-smokers–

a OR at 100 Bq/m3 (calculated).
b Analysis including measurement error.
c Analysis based on air monitors.
d Analysis based on surface monitors.

e Entire study period, 5–15 years.
f Radon-prone areas.
g	As reported in reference [B29], Blot et al. [B25] found no association between
lung cancer and radon.

458. Figure XVII is a graphical presentation of the data
for the individual studies shown in table 22. In addition,
the results of pooled analyses are also shown in the figure along with an estimate of the ERR from the combinedpooled studies (see table 22 and previous paragraph). As
discussed earlier, corrections for measurement uncertainty increase the predicted ERR somewhat, as shown intable 22.

United States. The meta-analysis included those studies
with 200 or more cases and long-term measurements ofindoor radon concentrations. The pooled study contained
in total 4,263 lung cancer cases and 6,612 controls. Using
published data from each of the eight studies, the authors
carried out regression analyses. The risks (OR) and 95%
CIs for each of the studies are also shown in figure XVII.
As was shown also in figure XIV, such results are generally
consistent with extrapolations from miner studies based on
the results of BEIR IV and BEIR VI [C19, C20]. As notedby Lubin and Boice, the CIs for the individual studies arelarge and include a relative risk of unity, which is consistent
with the possibility that there is no effect from exposure toresidential radon. Overall, however, the authors found thatthe excess risk from the combined data was significantly
different from zero. For residential exposure to 150 Bq/m3,
the authors estimated a combined OR of 1.1 (95% CI: 1.0,
1.3). Overall, Lubin and Boice concluded that the risk fromresidential radon is unlikely to be larger than that predictedon the basis of the miner data, that the negative exposure
response seen in some studies is likely to have been due toexposure misclassification or uncontrolled confounding factors, and that their results are consistent with a small effecton lung cancer from residential radon [L4].

459. Lubin [L36] discussed several case–control studies ofresidential radon in North America and two in China. He
acknowledged the inherent difficulties in establishing an
association between lung cancer risks and residential radon
concentration as a consequence of the overall expected risk
from radon and the large uncertainties. He noted the latest
studies in dosimetry, and suggested that pooling may permit
more detailed assessment of the risk. This has been demonstrated, for example, in the pooled residential case–control
studies in Europe [D17, D21], North America [K1, K26] andChina [L26].
460. Lubin and Boice [L4] undertook a meta-analysis
of eight case–control residential radon studies carried out
in China, Finland, Sweden, the United Kingdom and the290

UNSCEAR 2006 REPORT: VOLUME II

Figure XVII. Risk estimates from residential radon studies [B29].
Shown are the summary relative risks for exposure at a radon concentration of 100 Bq/m3 (except for references [L4, P17] at 150 Bq/m3) andthe corresponding 95% confidence intervals.
INDIVIDUAL STUDIES
Alavanja et al. (1999) [A18]
Blot et al. (1990) [B25]
Kreienbrock et al. (2001) [K18]
Letourneau et al. (1994) [L25]
Baysson et al. (2004) [B33]
Darby et al. (1998) [D15]Non-smokers

Alavanja et al. (1994) [A19]
Kreuzer et al. (2003) [K17]
Tomasek et al. (2001) [T10]
Wichmann et al. (2005) [W19]
Lagarde et al. (2001) [L32]
Pershagen et al. (1992) [P16]
Auvinen et al. (1996) [A5]Pershagen et al. (1994) [P11]
Lagarde et al. (1997) [L1]Non-smokers

Wang et al. (2002) [W13]
Field et al. (2000) [F11]
Oberaigner et al. (2002) [O2]
Lagarde et al. (2002) [L31]

Non-smokers

Pisa et al. (2001) [P6]Schoenberg et al. (1990) [S42]
Ruosteenoja et al. (1996) [R3]
POOLED ANALYSES
Lubin & Boice (1997) [L4]
Pavia et al. [P17]
German [W19]
Chinese� [L26]
North American� [K1, K26]
European� [D21]
0.1

0.2

0.5

1

2

RELATIVE RISK (at 100 Bq/m�)�Includes [B25, W13]
�Includes [A18, A19, F11, L25, S42, S76]
�Includes [A5, B26, B33, B43, L32, O2, P11, P16, R3, T29, W19]510

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

291

risk of lung cancer according to the time-weighted averageobserved residential radon concentration, after stratification
by study, age, region of residence and smoking habits. The
relative risks and 95% CIs for categories of radon concentration are shown in table 23. When the analysis was repeatedwith only those exposed below a radon concentration of
200 Bq/m3, the dose–response relationship remained statistically significant (p = 0.04).

461. Pavia et al. [P17] reported the results of a meta­analysis of residential exposures to radon gas and lung cancer from 17 case–control studies. Their analysis suggested
an association between residential radon concentration andlung cancer. A weighted log-linear regression analysis wasused to develop estimates of pooled ORs for exposure to aradon concentration of 100 Bq/m3, reported as OR = 1.15
(95% CI: 1.07, 1.24). The authors indicated that not all studies were adjusted for smoking and that their results cannot
exclude the possibility that their inability to fully adjust
for smoking (or other confounders) could account for theincreased risk seen in their study.

Figure XVIII. Relative risk of lung cancer versus observed­residential radon concentration.
The estimated ­linear ­relationship RR = 1 + 0.00084χ (solid line),with 95% ­confidence limits (dashed lines). The relative risk is equal to
1 at 0 Bq/m3 (adapted from figure 2 of reference [D21]).

462. Darby et al. [D17] reported a pooled analysis of
13 European case–control studies of the risk of lung cancer from residential radon. This analysis is of great interest,
especially since heterogeneity among the results of the various individual studies disappeared once the data from the
13 studies were put into a common format and analysed ina consistent manner. The study included 7,148 lung cancer
cases and 14,208 controls. The mean radon level measured
using long-term alpha track-etch detection in the houses of
the control group was 97 Bq/m3, and in the houses of lungcancer cases was 104 Bq/m3. Radon exposures during theprevious 5–34 years were considered in the analysis. The
authors investigated the association between radon concentration and lung cancer using two models. In one model therisk of lung cancer was proportional to (1 + βχ), where χ isthe measured radon level and β is the proportional risk factor per unit increase in radon. The second model subdivided
cases and controls by categories of radon exposure. The
authors noted that the dose–response relationship appeared
to be linear with no threshold and did not depend on smoking status. Before correcting for random uncertainties inmeasuring radon concentrations, the authors reported anincreased excess odds ratio (EOR) of about 0.08 (95% CI:
0.03, 0.16) per 100 Bq/m3. Figure XVIII shows the relative

RELATIVE RISK

3.00

2.00

1.00

0.00

0

200

400

600

800

1 000 1 200 1 400

OBSERVED RADON CONCENTRATION (Bq/m�)

Table 23 Relative risk of lung cancer according to time-weighted average observed residential radon concentration(adapted from table 16 of reference [D21])
Observed radon concetrationa(Bq/m3)

Number of cases

Number of controls

Mean observed radon
concentration

Relative risk
(95% CI)

<25

566

1 474

17

1.00 (0.87, 1.15)

25–49

1 999

3 905

39

1.06 (0.98, 1.15)

50–99

2 618

5 033

71

1.03 (0.96, 1.10)

100–199

1 296

2 247

136

1.20 (1.08, 1.32)

200–299

434

936

273

1.18 (0.99, 1.42)

400–799

169

498

542

1.43 (1.06, 1.92)

≥800

66

115

1 204

2.02 (1.24, 3.31)

Total

7 148

14 208a	Observed radon concentration for each address in the 30-year period ending 5 years prior to the index date weighted according to the length of time that the person lived atthat address.

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UNSCEAR 2006 REPORT: VOLUME II

463. When the analysis was restricted to people who had
lived in at most two residences during the previous 30 years,
the excess (EOR) increased to 0.094 (95% CI: 0.034, 0.175)
per 100 Bq/m3. Finally, after correction for random uncertainties in the assessment of radon concentrations, the dose–
response relationship remained linear, but the relative risk
doubled to 0.16 (95% CI: 0.05, 0.31) at 100 Bq/m3. Darbyet al. also discussed the combined effect of smoking and residential radon exposure on the absolute risk of lung cancer,
and indicated, using the same relative risk factor of 0.16 forlifetime exposure (taken as 75 years) to a radon concentration of 100 Bq/m3, that the risks of lung cancer in lifelong
non-smokers and cigarette smokers would be about 0.47%
and 11.6%, respectively. Expressed differently, almost all of
the risk accrues to the population of smokers.
464. Krewski et al. [K1, K26] reported a pooled analysisof residential radon exposure and lung cancer risk in sevencase–control studies in North America. The combined study
included some 3,662 cases and 4,966 controls. Residential radon levels were determined using long-term alphatrack detection spanning 12 months. The analysis focusedon ­exposures 5–30 years prior to the interview date. Data
were analysed using conditional likelihood regression andthe ­linear model OR (χ) = 1 + βχ, where χ is the cumulative radon exposure in the previous 5–30 years. The authors
stated that the EORs for individual studies ranged from 0.01
(95% CI: <0.00, 0.42) per 100 Bq/m3 in a Missouri study[A18] to 0.56 (95% CI: 0.22, 2.97) per 100 Bq/m3 in a New
Jersey study [S42]. The authors also investigated potential
modifying effects of smoking and demographic factors, andnoted that, while there was no apparent heterogeneity in
EOR by sex or education level, there was some suggestionof a decreasing radon-associated risk with increasing age.
The authors also indicated that they had found no significant
differences in EOR with measures of smoking status. Overall, an EOR = 0.11 (95% CI: 0.00, 0.28) per 100 Bq/m3 wasestimated. The authors also indicated that analyses restricted
to subsets of the data with “presumed more accurate radon
dosimetry” resulted in increased estimates of the EOR, of
0.18 (95% CI: 0.02, 0.43) per 100 Bq/m3.
465. Smoking is a potential confounding factor in both
the residential and the miner studies. However, there is far
more information about smoking histories available in theresidential studies. Furthermore, several analyses show thatthe information collected on smoking habits gives estimatesof the risks of lung cancer that are very much in line withthose of other studies of lung cancer and smoking [C20,
C39, P20].
466. Becker [B34] reviewed the radon experience of miners, residential radon exposure and “the therapeutic use” ofradon. Becker noted that BEIR VI [C20] indicated that smoking has a much greater risk of lung cancer than does exposure to radon. He then argued that the correction of the miner
data for smoking is complicated by the under-reporting ofactual smoking, and hence that the uncertainties associatedwith retrospective analysis of smoking greatly confound theanalysis of radon risk in miners. Similar arguments weremade about the residential radon studies, noting that uncertainties in smoking “by far dominate” the uncertainties associated with the retrospective analysis of exposure to radon.
467. Conclusion. The main studies of residential radon are
the pooled analyses of European [D17, D21], North American [K1, K26] and Chinese [L26] residential case–control
studies. These studies indicate a significant association
between the risk of lung cancer and exposure to residential
radon. The studies also examined the effect of restricting
analyses to those who had lived in at most two residences. In
addition, Darby et al. [D17, D21] also carried out analyses
that adjusted for exposure uncertainty. Both the European
and the North American studies have looked at the estimatedrelative risk from radon for individuals with different smoking habits and demonstrated not only that there is no significant heterogeneity, but that the risk estimates on the relative
scale are very similar for individuals in different smoking
categories.
468. Table 24 shows the ERR per unit residential radon
concentration from three pooled analyses of case–control
studies. The pooled analyses also reported the ERR for
analyses restricted to individuals who had lived in only one
or two residences and hence, whose radon exposures arepresumed to be more precisely known than those of individuals who changed residences many times. The ERR
estimates from the restricted analyses were higher than the
ERR estimates from the primary analyses. The analysis byDarby et al. used a regression model correction for exposure uncertainty, which approximately doubled the ERR
obtained from analysis of the primary data [D21].

Table 24 ERR per unit radon concentration in air (per 100 Bq/m3) and 95% confidence intervals from combined residential
radon studies
Study

Primary analysis

Restricted analysis

Exposures adjusted
for uncertainty

European [D17, D21]

0.084 (0.03, 0.158)

0.094a (0.034, 0.175)

0.16c (0.05, 0.31)

North American [K1, K26]

0.11 (0.00, 0.28)

0.18a (0.02, 0.43)

Chinese [L26]0.133 (0.01, 0.36)

0.319b (0.07, 0.91)

Combined

0.093 (0.04, 0.15)

0.11 (0.05, 0.19)

a Only one or two residences and at least 20 years of coverage.
b Only one residence with complete coverage.

c Correction for measurement uncertainty.

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

469. An analysis was conducted to (approximately) combine the risk estimates from these three pooled analyses
using weighting by the inverse of the variance in the ERR
estimate. The ERR estimate for the combined primary analyses was 0.093 per 100 Bq/m3, and for the restricted analyses,
the combined ERR estimate was 0.11 per 100 Bq/m3. The
European study [D21] provided 72% of the weight (of information) from the primary analyses and 82% of the weight(of information) from the restricted analyses.
470. The European study provided an ERR estimate of 0.16
per 100 Bq/m3 of residential radon, based on a correction formeasurement uncertainty that was about twice as high as the
ERR from the primary analyses [D21]. The measurement

293

error correction of Darby et al. [D21] requires a number ofassumptions concerning the distribution describing the longterm average residential radon levels within a geographicalarea and the variability associated with repeated measurements of radon concentrations in the same dwelling. However, as discussed by Darby et al. [D21], there is also information to support many of the assumptions concerning, forexample, the magnitude of year-to-year variability. Nonetheless, at this time it seems reasonable to adopt the estimate
corrected for random uncertainties in the assessment of radon
concentrations from Darby et al. [D21], namely an ERR of
0.16 (95% CI: 0.05, 0.31) per 100 Bq/m3, as an appropriate,
if possibly conservative, estimate of the (lifetime) risk fromresidential radon.

VI.

EFFECTS OF RADON ON organs and tissues other than THE lungA.

Dosimetric considerations

much higher dose than any other organ. Of the other organs
and tissues, those with a high fat content receive somewhat
higher doses from radon gas. Red bone marrow, thought tobe the tissue in which childhood leukaemias originate, does
not receive doses that are large compared with those to othertissues. Nevertheless, the calculated doses are high enough tosuggest that radon may be responsible for a small proportion ofchildhood leukaemias [H38]. It is possible that alpha particles
from RDPs irradiate the cells in which skin cancers originate
and thus induce skin cancer. However, the location of thesesensitive cells is not known with certainty, and it is possiblethat they are too deep to receive a significant dose. If they areirradiated, it is likely that the doses would be larger in the caseof children than in adults. However, the evidence so far available is inconclusive.

471. It is generally recognized that the main hazard frominhaled RDPs is from irradiation of the lung. However, insome circumstances, irradiation of the stomach from ingestion of water containing dissolved radon gas may need to beconsidered. One important factor is the length of time thatingested radon remains in the stomach [I14]. Various estimates of dose to the stomach are within a factor of about 10[N9], which is quite a small variation considering the uncertainties associated with such estimates. Calculations alsosuggest that decay products deposited on the skin may becapable of irradiating the sensitive basal cells [S4].
472. In addition to the dose to the lung, Jacobi and Eisfeld[J2] and Harley and Robbins [H38] calculated the dose toorgans other than the lung, such as kidney, bone marrow andskin. Kendall and Smith [K21] applied ICRP dose models toestimate effective doses to organs and tissues from radon andits decay products, including doses arising via inhalation,
external exposure of the skin and ingestion. The aim was toprovide a self-consistent summary that would allow the various hazards to be compared and put into context. The largest
dose overall from inhaled radon and its decay products wasto the respiratory tract; doses to other organs were usuallyat least an order of magnitude smaller (see table 25, adaptedfrom reference [K21, table 2]). In particular, doses to tissues
with a relatively high fat content (such as red bone marrow),while somewhat higher than those to most other tissues, did
not appear to be high enough to present a particular problem.
The conclusion was that the conventional focus on risk oflung cancer from inhaled RDPs was appropriate

475. Robbins and Harley [R5] suggested that maternal
ingestion of radon in water can deliver a dose to the foetus.
Radon is transported by the blood and diffuses throughout thebody, including the placenta. Radon and its short-lived decay
products can thus reach an embryo/foetus. There is an interval during pregnancy when the foetus is at the highest risk ofsevere effects from radiation exposure. For the early embryo,there are two important factors to consider: the very small sizeof the embryo yields a small target for alpha particle hits, butconversely, alpha particle damage to DNA may have major
consequences. The dosimetric calculations for the developing
embryo and foetus used the maternal and the foetal placental
blood supply at different points in time [R5]. These calculations relied on available published data for these, and for foetal
weights. The accumulation of radon in various compartments
following the ingestion of 100 Bq of radon dissolved in water
was estimated, and the dose determined as a function of time,on the basis of the pharmacokinetic model developed. The ratioof the weight of blood in the embryo/newborn infant to its total
weight was assumed to be a constant, as was blood flow to theplacenta at 115 mL kg–1 min–1. These parameters need verification, as they are critical not only for such calculations but forother basic toxicological calculations. The clearance half-timesof radon for the various tissues in the mother were based on
published human data [H24, H25]. For an average intake of
0.6 L of raw tap water per day, containing a radon concentration of 100 Bq/L, the calculated total equivalent dose to the
foetus over the term of pregnancy was 250 μSv. The highest
calculated equivalent dose of 3 μSv/week occurred between
weeks 6 and 16 [R5]. The foetal doses estimated in reference[R5] are considerably higher than those reported by Kendall
and Smith [K19], who have noted that the foetus has little fat
until late in gestation. The difference between these estimatesoriginated from the assumptions made in the models.

473. Kendall and Smith [K21] also considered the dose to
the foetus. For RDPs, they adopted the foetal discrimination
factors of the ICRP. The ICRP makes no recommendation
for radon gas. Kendall and Smith noted that, for many radionuclides, the dose to the foetus is similar to that to maternal
muscle. Arguing that the fat content of the foetus is low, they
assumed that maternal muscle provides a reasonable surrogate. An alternative approach is discussed below.
474. According to Kendall and Smith [K19], the general
pattern of doses to different tissues for inhalation and ingestion of radon and its decay products by children is similar tothat in adults. Both for inhalation and for ingestion, the organ
of intake receives much higher doses than any other organ. In
the case of inhalation, the largest doses are to the lung and theextrathoracic part of the respiratory tract (the nose, pharynx
and larynx). In the case of ingestion, the stomach receives a
295

Po

3.3 10-11

3.0 10-11

2.8 10-11

3.9 10-11

3.4 10-11

5.8 10-11

2.8 10-11

8.7 10-11

2.8 10-11

2.7 10-11

2.8 10-11

2.8 101.4 10-9

1.3 10StomachSmall intestine

Colon

RBM

Bone surface

Liver

BreastKidneyGonadsBrain

Bladder

MuscleCED

Foetus-11

6.8 10-11

6.8 10-11

6.8 10-11

5.8 10-9

6.8 10-11

6.8 10-11

1.0 10-10

6.8 102.1 10-9

5.0 10-10

1.4 10-10

4.8 10-9

1.4 10-10

1.5 10-10

2.1 10-10

1.4 10-11-11

2.6 10

6.8 10-11

3.2 10-10

5.3 108.2 10-11

1.5 10-10

5.6 10-9

7.5 10-11

1.7 10-10

2.5 10-9

9.4 10-11

1.7 10-10-11

6.3 10-8

2.7 10-8-10

3.9 10-8Bi

214

1.4 10-8Pb

214

4.1 10-11

3.6 10-9

1.0 10-10

1.4 10-10

1.1 10-10

1.0 10-10

3.6 10-9

1.0 10-10

2.9 10-10

1.0 10-9

1.9 10-10

1.1 10-10

1.2 10-10

1.3 10-10

3.0 10-8

2.5 10-8

Mixture

25

0.06

5.30

0.15

0.21

0.15

0.15

5.20

0.15

0.43

1.48

0.28

0.16

0.17

0.19

44.5

35.8

Annual dose (mSv)
at 200 Bq/m3

Po

1.6 10-12

3.3 10-9

3.0 10-12

2.9 10-12

2.9 10-12

2.9 10-12

9.5 10-12

3.0 10-12

6.3 10-12

3.6 10-12

4.2 10-12

3.9 10-12

6.9 10-12

1.2 10-11

6.7 10-9

2.7 10-8

218

6.2 10-12

1.6 10-8

1.6 10-11

2.2 10-11

1.5 10-11

1.5 10-11

5.0 10-10

1.6 10-11

5.6 10-11

2.5 10-10

3.7 10-11

2.4 10-11

4.9 10-11

7.7 10-11

4.8 10-8

1.3 10-7Pb

214

2.9 10-12

1.4 10-8

7.4 10-12

1.0 10-11

7.1 10-12

7.1 10-12

5.9 10-10

7.8 10-12

7.6 10-12

7.3 10-12

7.4 10-12

9.9 10-12

1.8 10-11

4.9 10-11

9.3 10-8

1.1 10-7Bi

214

4.9 10-12

1.3 10-8

1.1 10-11

1.5 10-11

1.1 10-12

1.1 10-11

3.7 10-10

1.2 10-11

3.2 10-11

1.2 10-10

2.2 10-11

1.7 10-11

3.2 10-11

5.6 10-11

4.9 10-8

1.1 10-7

MixtureDose per unit intake (Sv/Bq)

Dose per unit intake (Sv/Bq)

25

0.01

19.7

0.02

0.02

0.02

0.02

0.54

0.02

0.05

0.17

0.03

0.02

0.05

0.08

70.9

159

Annual dose (mSv) at
200 Bq/m3

0.04

0.28

0.05

0.05

0.06

0.05

0.05

0.42

0.09

0.03

0.65

0.05

0.06

0.06

1.2

Annual dose (mSv)
at 200 Bq/m3

Rn

222

Note: The activity of decay products in comparison with radon is taken to be 218Po:214Pb:214Bi in the ratio 0.9:0.45:0.225; F = 0.41. 90% of decay products are taken to be attached to aerosols, 10% unattached. Annual volume breathed taken to be
7,300 m3. Ext.thor. is the extrathoracic part of the respiratory tract. RBM is red bone marrow. CED is committed effective dose. Kidney is shown as the organ receiving the next highest dose after the respiratory tract. The dose to the foetus fromradon gas taken to be that to muscle over 9 months. Dose to skin is taken from reference [N9]. It does not depend on type.

Skin

4.4 10-9

Ext.thor.

-11

1.1 10-8

Lung

218

Type MType F

Table 25 Summary of dose coefficients from inhaled radon decay products together with annual doses from decay products and radon at 200 Bq/m3 (see reference [K21]for discussion)
296
UNSCEAR 2006 REPORT: VOLUME IIANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

476. Kendall and Smith [K19] also investigated differences between doses to adults and those to children aged 1
and 10 years. They calculated both dose coefficients, i.e. theequivalent dose resulting from an intake of unit activity ofthe material in question, and also annual equivalent doses.
The former tend to be higher in children than in adults. The
differences between the annual doses to children and adults
are smaller than the differences in the dose coefficients,
because children breathe less air and ingest less water thanadults.
B. Epidemiology of cancers other than lung477. Annex A, “Epidemiological studies of radiationand cancer”, provides a discussion of the epidemiologicalmethods as well as information complementary to thatprovided in this annex.
478. The currently available epidemiological evidence
indicates that risks other than lung cancer from exposure toradon and its decay products are likely to be small. A numberof studies of residential radon and non-lung cancer are available. Laurier et al. [L18] reviewed 19 ecological studies,
8 residential case–control studies and 6 miner cohort studies published between 1997 and 2000 in order to examine a
possible association between radon exposure and leukaemia.
While the ecological studies suggested a possible positive
association, the case–control studies and the miner cohort
studies did not. Overall, the authors concluded that the available data (i.e. to 2000) did not provide any evidence of an
association between radon exposure and leukaemia [L18].
479. A series of papers [E6, L38, T28] followed the review
by Laurier et al. [L18]. Eatough [E6] noted that potential
leukaemia risk to miners was investigated by looking fortrends in relative risks with increasing exposure to RDPs.
However, studies of the dosimetry for radon gas and RDPs
suggest that the doses to the red bone marrow are not appreciable [H38]. Tomasek [T28] reported a study of Czech miners in which an excess of leukaemia related to duration of
exposure was observed, and went on to suggest that durationof exposure in the mine environment is likely to be a surrogate for exposure to uranium dust, which he considered to be
the predominant dose contributor. Laurier et al. [L38] commented that results published since their review [L18] did
not modify the conclusions that available epidemiological
data do not demonstrate an association between leukaemia
and exposure to radon. They also noted that if such a relationship exists, the association would be slight and of little
significance for residential exposure.
480. An ecological study of cancer incidence and radonlevels in the south-west of the United Kingdom looked at
14 major cancer sites, using data for the South-Western Cancer registry [E5]. Average radon levels for residences were
sorted into 10 categories, from low (<40 Bq/m3) to very high(>230 Bq/m3), and age standardized cancer rates were calculated for each category. Incidence rates for lung cancer were297

similar across all radon categories. Except for non-melanoma
skin cancer, the authors found no significant positive correlation with radon. Overall, the authors found no significant
difference in the corrected survival rates for any cancer site
between the low- and high-radon areas.
481. A research letter by Law et al. [L20] commented
on residential radon exposure and leukaemia, referring
to an earlier study [K24] of acute leukaemia in the southwest, north and north-west of England of residents aged
16–69 years. Radon measurements (of about 6 months’
duration) were made in the living rooms and bedrooms of
1,881 houses (about 78% of the homes of the subjects, 76%
for the cases and 80% for the controls). No association was
found between acute leukaemia and radon concentration; theauthors concluded that their study did not support exposure
to residential radon as a causal factor in leukaemia in the
United Kingdom.
482. Thorne et al. [T24] investigated possible associations
between residential radon exposure and paediatric cancers in
Devon and Cornwall, United Kingdom. This study compared
the incidence of childhood cancers in postal sectors with low
radon concentrations (<100 Bq/m3, average 57 Bq/m3), andhigh radon concentrations (>100 Bq/m3, average 183 Bq/m3)
in a total of 238 postal sectors. The authors found no significant difference in cancer incidence rate between low- andhigh-exposure sectors.
483. Steinbuch et al. [S46] reported an investigation of residential radon exposure and risk of childhood acute myeloid
leukaemia (AML). Alpha track detectors were placed in thehouses of 173 cases and 254 controls for 1 year. Overall,there was no association between residential radon concentration and the risk of AML. Lubin et al. [L12] reported anage-matched study of childhood acute lymphocytic leukaemia (ALL) and residential radon exposures of children in the
United States. Radon levels for the 505 cases and 443 controls were estimated for 97% of the exposure period. The
mean radon level was lower for cases (65.4 Bq/m3) than forcontrols (79.1 Bq/m3). No association between ALL andradon exposure was found.
484. Evrard et al. evaluated the ecological association
between indoor radon concentrations and acute leukaemia
incidence among children under 15 years of age in France[E7]. The study considered the whole country, divided into
348 geographical units. Incidence data included 4,015 casesof acute leukaemia registered by the French National Registry of Childhood Leukaemia and Lymphoma between 1990
and 1998. Exposure was based on a national campaign of
13,240 indoor radon measurements. A positive ecologicalassociation was observed between indoor radon concentration and childhood leukaemia incidence, on the borderlineof statistical significance (p = 0.053). A significant association was observed for AML (p = 0.004) but not for ALL(p = 0.49). Consideration of exposure to terrestrial andcosmic radiation did not modify the observed association
between radon exposure and incidence of AML [E8].

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UNSCEAR 2006 REPORT: VOLUME II

485. In a 1993 paper, Tomasek et al. [T51] reported ona study of RDP exposure and cancers other than lung cancer in a cohort of uranium miners in western Bohemia.
These authors investigated site-specific cancer mortality in
4,320 miners who had been followed-up for an average of
25 years. An average exposure of 219 WLM was reported.
Data on smoking habits and alcohol consumption were notavailable. An analysis of rates of observed to expected numbers of deaths (O/E) showed that overall, the risk of deathfrom non-lung cancers was slightly greater than the naturalrates but not significantly so. A significant excess of non-lungcancer mortality in men who started to mine when they wereyounger than 25 years of age was found, but the increase was
not related to cumulative RDP exposure. Overall, the authors
concluded that there is no significant risk of any cancer otherthan lung cancer, although further investigation is needed ofthe effect of RDP exposure on cancers of the gall bladderand extrahepatic bile duct and on multiple myeloma.
486. Rericha et al. [R13] reported an investigation of theincidence of leukaemia, lymphoma and multiple myeloma,
a total of 177 cases, in Czech uranium miners, using a retro­
spective case–cohort design in a study of 23,043 uranium
miners. The authors found no apparent association between
RDP exposure and either non-Hodgkin’s lymphoma or multiple myeloma, but did find an association between RDP
exposure and an increased risk of leukaemia (chronic lymphocytic leukaemia, CLL), which was not previously thought
to be radiogenic. The association is based on a comparisonof the relative risk (RR) of CLL in miners who had an RDP
exposure of 110 WLM with those who had an exposure of
3 WLM. An RR of 1.98 (95% CI: 1.10, 3.59; p = 0.016) was
reported.
487. Darby et al. [D22] reported on RDP exposure and nonlung cancer in a group of Swedish iron ore miners. Theseauthors observed that, when the mortality from all cancers
other than lung in miners was compared with that expected
for the northernmost county of Sweden, the O/E ratios werehigher in men with exposures of >100 WLM than in men
with lower exposures, but the trend was not significant. As in
the western Bohemian study [T51], excesses were seen forcancer of the gall bladder and extrahepatic bile duct and for
multiple myeloma; however, the increases were not statistically significant.
488. Darby et al. [D12] carried out a collaborative analysisof 11 miner studies to look for risks of cancer other than lungcancer. The miner populations included 10 of the 11 minercohorts covered by the joint analysis by Lubin et al. [L10](the Radium Hill study [W15] was excluded since follow-up
was incomplete). The 11th miner cohort added in the analyses by Darby et al. was a cohort of Cornish tin miners [H29].
Overall, the study [D12] included some 64,209 men who had
worked for an average of 6.4 years, accumulated an ­average
exposure of 155 WLM and been followed for an averageof 16.9 years. External mortality data were available for 10of the studies (but not for the Chinese study of Xuan et al.
[X1]). In total, some 1,179 non-lung cancer deaths wereobserved, which was close to the expected number (O/E =1.01; 95% CI: 0.95, 1.07). Among the 28 individual cancer
types studied, the only sites of statistically significant excess
cancers were the stomach and the liver; however, mortality
from these cancers was not related to cumulative RDP exposure in WLM and was thus unlikely to have been caused byradon. Among the leukaemias, the O/E ratio for acute myeloid leukaemia was larger than for other leukaemias, but was
not statistically significant. Overall, the authors concluded
that exposure to high concentrations of radon in the air isunlikely to result in an increased risk of cancer mortality
other than from lung cancer [D12].
489. Möhner et al. [M44] reported on a study of leukaemia in German miners. This case–control study included
377 cases and 980 individually matched controls. Exposures
were based on a job–exposure matrix that included exposure to radon and its decay products as well as exposure
to external gamma radiation and to long-lived radioactive
dust. Using logistic regression methods and taking study
power into account, the authors concluded that a “casual
relationship between radon progeny and risk of leukaemia
can largely be excluded”. The study did, however, suggest
an elevated risk of leukaemia for cumulative exposures of>400 WLM.
490. Kreuzer et al. [K29] reported on a study of the risk oflung cancer and other cancers in the German uranium minercohort. The study, based on external comparisons, showeda statistically significant excess of lung cancer risk and atrend of increasing risk with increasing cumulative RDP
exposure. The study also found an excess, although not statistically significant, mortality from liver and lung cancer.
The authors indicated that the excess is unrelated to cumulative RDP exposure and noted that, in the early years ofmining, the Wismut mining company provided miners withalcohol and cigarettes free of charge. Tirmarche et al. [T9]reported an excess of larynx cancer in French miners; however, ­Villeneuve et al. [V4] found no significant relationshipbetween RDP exposure and other cancer sites.
491. The update of data for the Eldorado miners [H35]involved 17,660 workers. The update extended the mortality analysis by almost 20 years and added 30 years of new
information on cancer incidence. A total of 5,332 deaths
occurred between 1950 and 1999, and 2,355 workers developed at least one cancer between 1969 and 1999. Mortality
and cancer incidence were compared with those of the general Canadian population. Lung cancer was the only cancersite with an excess for both mortality and cancer incidence.
In the internal analysis, there was no meaningful evidence ofany causal relationship between RDP exposure and increased
risk of any cancer other than lung cancer.
C. Effects other than cancer
492. A full discussion of this subject is provided in annex B,“Epidemiological evaluation of cardiovascular disease andANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

other non-cancer diseases following radiation exposure”.
This section provides a few additional observations fromepidemiological studies of miners.
493. Villeneuve and Morrison [V2] investigated the mortality from coronary heart disease (CHD). In this study, thecohort consisted of 1,743 underground miners and 321 millers or surface workers. Men in this cohort had a mean exposure to RDPs of 378.6 WLM over an average of 5.7 years.
As for other analyses of these miners, exposures were basedon the analysis of Corkill and Dory [C12]. Smoking datafrom a 1993 survey were used to update data from previous surveys in 1960, 1966, 1970 and 1978. Smoking status(current, former or never smokers) was determined for 59%
of the cohort. Finally, data on the mortality experience ofthe cohort were available up to 1990. Multivariate Poisson
regression analysis was used to estimate the relative risk of
CHD from RDP exposure, with adjustments for attained age,duration of exposure and smoking status. Unlike previous
studies of this cohort, the analyses [V2] used internal comparisons to control for bias potentially introduced throughthe healthy worker effect. This study found that workers
with high cumulative exposures to RDPs (over 1,000 WLM)
had an elevated risk of CHD (RR = 1.5; 95% CI: 0.77, 2.75)
compared with those with no exposure. However, no statistically significant trend of increasing risk with increasing
exposure was found, nor was there a statistically significant
interaction between cumulative RDP exposure and smoking
status. The authors found that smoking status was a significant predictor of CHD.
494. Villeneuve et al. [V6] further explored the relationship between mortality from CHD and RDP exposureusing both external and internal cohort comparisons in theirmost recent update of this cohort. There was no association
between cumulative exposure in WLM and CHD mortality.
A reduced CHD mortality rate was observed relative to the
population of Newfoundland males (SMR = 0.86; 95% CI:
0.74, 0.98); the authors attributed this reduction to a “healthy
worker effect”.
495. Kreuzer et al. [K30] reported an investigation of themortality from cardiovascular diseases in a cohort of German

299

uranium miners. The cohort included 590,001 male subjects
who were employed for at least 6 months between 1946 and1989 at the former Wismut uranium company in eastern­Germany. As for other studies of Wismut miners, exposures
to RDPs, long-lived radionuclides and external gamma radiation were estimated using a detailed job–­exposure matrix.
As of 31 December 1998, the cohort included (about)
16,598 deceased miners, with 5,417 deaths from cardiovascular diseases. Linear Poisson regression models were used
to estimate the ERR per unit of cumulative radiation exposure after adjusting for attained age and calendar period.
The study found no trend in risk of circulatory diseases withincreasing cumulative exposure to RDPs (ERR was 0.0006(95% CI: -0.004, 0.006) per 100 WLM), external gamma
radiation (ERR was -0.26 (95% CI: -0.6, 0.05) per Sv) or
long-lived radionuclides (ERR was -0.2 (95% CI: -0.5, 0.06)
per 100 kBq h m–3. The authors concluded that their results
did not support an association between cardiovascular­disease mortality and exposure to radiation among uranium
miners.
496. More recent data are also available from Canada.
Howe [H35], in the update of the Eldorado cohort study,reported that for most of the causes of death, the cohort as awhole as well as various subcohorts had reduced risks relative to the population. In particular, the authors stated that,
although an increase of hypertensive disease was observed(significant in only one of the subcohorts examined), thestudy indicated a statistically significant deficit in males ofcardiovascular diseases such as stroke and ischaemic heartdisease. The authors go on to suggest that this is probably aconsequence of the healthy worker effect, since heart disease
would be likely to prevent people working in a strenuous
physical occupation such as mining. Villeneuve et al. [V6]
found that there was no apparent trend between cumulative
exposure to radon and the relative risk of death from CHD(p = 0.63). Furthermore, this finding was unchanged after
adjusting for lifetime smoking status, which was available
for approximately 54% of the cohort. Additionally, cumulative exposure to radon was found to be unrelated to diseasesof the circulatory system, acute myocardial infarction andcerebrovascular disease. All of these findings are consistent
with those of Kreuzer et al. [K30].

VII.

Implications for risk assessment

497. Epidemiological studies of underground miners provide
the current basis for estimating the risk of exposure to radonand its decay products. Studies of smoking and non-smoking
miners show that exposure to RDPs carries an enhanced riskof lung cancer. However, miners were not exposed to RDPs
alone but rather to RDPs in association with many otheragents, among them silica (quartz), various metals, including arsenic, and diesel exhaust. Traditionally, the risks fromdomestic radon exposure have been developed by extrapolation from the data for miners; however, the risks from residential radon can now be estimated directly using the results ofpooled residential case–control studies (see section V).
A.

with the BEIR IV [C19] estimate of 140 per 106 personyears per working level month for male non-smokers. The
approach in reference [S10] was necessarily limited because
the authors did not have access to data on individual miners
and could only utilize published relative risks within categories of cumulative exposure (J h m–3 or WLM); similarly,
no attempt could be made to evaluate patterns of error fordifferent mining periods within a study. For the ERR model,
the authors estimated that the most likely range for the ERR
per unit exposure para­meter for this group of studies was
0.009–0.00005 (J h m–3)–1 (0.5–1.5% WLM‑1).
Figure XIX. Distribution of absolute risk for Colorado
miners with no exposure prior to 1950 [S10].

General studies of exposure uncertainty

498. There is a great deal of interest in dosimetry-based
estimates of risk from exposure to RDPs (see section II).
There is a discrepancy between the dosimetry-based and theminer-epidemiology-based approaches, with differences in
estimates of a factor of 2–3. The miner epidemiology data
provide the basis for investigating biological mechanisms andmodifiers such as time since exposure and age at exposure.
As discussed in section IV, there is considerable uncertainty
in the exposures of miners, with uncertainty increasing thefurther we go back in time. In view of the ongoing importance of miner studies in understanding the risk from exposure to radon and RDPs at work and in the home, it is useful
to consider the effects of uncertainty in miner exposures on
the epidemiological analyses performed using miner data.

PROBABILITY DENSITY

Mean β1 = 85.16 x 10-6

499. A 1989 report considered the effects of uncertainty in
exposure estimates for the studies then available of miners in
Ontario, Beaverlodge and Port Radium (Canada), Czechoslovakia, Sweden, and the Colorado Plateau (United States)[S10]. Several subgroups of the Colorado Plateau cohort
were also studied. The focus of the study was on adjusting
the exposure–response parameter in the simple linear modelfor excess relative and excess absolute risk. The authors useda Bayesian error-in-variable relative risk approach to analytically assess the effect of uncertainty on the exposure–
response relationship. Figure XIX shows the posterior probability density factor for the absolute risk coefficient for
Colorado miners with no mining exposure prior to 1950. At
that time (i.e. follow-up to 31 December 1982), the cohort
with no exposure before 1950 included 168 miners withlung cancer. The mean lifetime risk was estimated at about
85 cases per 106 person-years per working level month,with a range of about 60–150 cases per 106 person-years per
working level month [S10]. The estimate may be compared

95% Confidence
interval

-100

-80

-60

-40

-20

0

20

40

60

80

100

120

140

160

RISK, β1 (106 WLM)-1

500. A subsequent paper by Chambers et al. [C15] reviewed
the factors affecting exposure estimation for seven groups ofuranium miners. These included the temporal and spatial variability of radon and RDP levels in the workplace, changing
mining methods and ventilation practices, and uncertainties
about the miners’ work histories. Bayesian methods were used
to develop posterior probability density functions of absoluteand excess relative risk coefficients for each of the cohorts.
Figure XX shows the posterior probability density function
for the ERR. The authors noted that the estimated RDP exposures in WLM are uncertain in all the miner studies. For thereasons given in reference [S10], the studies of United Statesminers [H17, P14, S10], Ontario miners [M19, M23], and
301

302

UNSCEAR 2006 REPORT: VOLUME II

Czechoslovak miners [S25] were considered at the time toprovide the strongest basis for risk estimation. The strengthsof these groups are that they are all large, well-traced cohorts
for which considerable information existed on which to baserisk estimates for the period of interest. To the best of theauthors’ knowledge, there was no systematic bias with regard
to the magnitude of the group exposures estimated in thesestudies. However, the uncertainties in exposures of individual
miners could be very large, particularly for exposures associated with the early years of mining. Of these three groups,
the United States miners were exposed to the highest RDP­concentrations, the Czechoslovak miners were exposed tointermediate levels, and the Ontario miners were exposed to
the lowest levels. While figure XX showed the posterior probability density function for all cohorts, figure XXI illustrates
the posterior probability density function developed by combining the three studies considered “best”, together with thatfrom combining all the studies (with and without a notionalcorrection for bias in the Swedish and the Beaverlodges­ tudies). Overall, Chambers et al. [C15] estimated the ERR
to have a 95% uncertainty range of 0.5–1.5 per 100 WLM,
comparable to the estimates of Lubin et al. [L10].

Figure XX. Excess relative risk (posterior) probability
density functions with unadjusted variances [C15].

Figure XXI. Combined excess relative risk (posterior)probability density functions [C15].

PROBABILITY DENSITYPROBABILITY DENSITY

United States
Port
Radium
Newfoundland
Czechoslovakia
Ontario
SwedenBeaverlodge
-1

0

1

2

3

501. The pooled analysis by Lubin et al. [L10] considered
the impact of uncertainties in the estimates of RDP exposure
only in the most general way and only within the context
of modification of the exposure–response relationship by
exposure rate and exposure duration. Exposure uncertainty
was considered greatest in the earliest years of mining, i.e.
those years in which exposure rates were highest. According to Lubin et al. [L10], exposure uncertainty would therefore have tended to attenuate the effects of high exposuresand potentially induce an inverse exposure-rate pattern[L22]. They estimated an ERR and associated 95% CIs as
depicted graphically in figure XXII. Lubin et al. [L22] report
a combined estimate of ERR of 0.49 (95% CI: 0.2, 1.0) per
100 WLM, where the joint 95% CI is based on a randomeffects model.

4

EXCESS RELATIVE RISK (% PER WLM)

502. BEIR VI [C20] also investigated the role of uncertainty
in estimating lung cancer risk, and developed an extensive
table summarizing the sources of uncertainty in estimating
lifetime risk from residential exposure to radon. The BEIR VI
estimates of lung cancer risk were based on analyses of thedata from miner epidemiological studies. The BEIR VI Committee acknowledged that there were uncertainties in exposures to RDPs and other potential factors, such as exposure
to cigarette smoke and arsenic. No systematic bias in the estimates of miner exposure to RDPs was identified. BEIR VI
also suggested that random errors might result in an underestimate of the slope (risk) of the exposure–response relationship. In addition to extensive qualitative discussion, theBEIR VI Committee applied quantitative methods for uncertainty analysis, acknowledging that their analysis should be­considered illustrative, “not to replace the Committee’s newcomprehensive qualitative analysis”, since not all sources of
uncertainty could be identified and characterized. The BEIR VI

Bias corrected

All studies

Best three

-1

0

1

2

3

EXCESS RELATIVE RISK (% PER WLM)

Committee’s “preferred” uncertainty limit was derived using
a simple constant relative risk model fitted to the miner data
at cumulative exposures of below 50 WLM. On this basis,
BEIR VI estimated an ERR of 1.17 per 100 WLM, with a
95% uncertainty interval from 0.2 to 22.5 per 100 WLM.
503. Notwithstanding the different data and methods thathave been used to develop combined estimates of ERR fromthe studies of miners, there is a remarkable consistency inthe ranges of ERR per 100 WLM that have been estimated.
As discussed in the preceding paragraphs, the 95% CI for the
ERR was estimated by Chambers et al. [C15] as 0.5–1.5 per
100 WLM; by Lubin et al. [L10] as 0.2–1.0 per 100 WLM;
and by BEIR VI [C20] as 0.2–22.5 per 100 WLM, withcentral estimates of ERR ranging from 0.49 to 1.17 per
100 WLM. On the basis of the most recent results of minerstudies as described earlier in this annex (see para. 427), an
estimate of a combined ERR of 0.59 (95% CI: 0.35, 1.0)

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per 100 WLM is consistent with previous estimates but withsomewhat narrow 95% CIs. By multiplying a notional lifetime risk of lung cancer of, say, 8%, and an average ERR of
Figure XXII.

303

0.59 (95% CI: 0.25, 1.0) per 100 WLM, a nominal absolute
lifetime risk of about 4.7 × 10–4 (95% CI: 2.8 × 10–4, 8 × 10–4)
for 100 WLM can be calculated.

Estimates of excess relative risk of lung cancer per WLM (adapted from [L10]).
Yunnan tin
Czech uranium
Colorado uranium
Ontario uranium
Newfoundland fluorspar
Swedish iron
New Mexico uranium
Beaverlodge uranium
Port Radium uranium
Radium Hill uranium
French uranium
Combined
0.0005
0.002
0.001

0.005

0.02

0.05

0.01

0.2
0.1

EXCESS RELATIVE RISK (WLM-1)

B.

Biologically based models

504. Several investigators reported analyses of experimental animal data and epidemiological data with multistage
models to study mechanisms of carcinogenesis and assess
risks from the inhalation of RDPs (e.g. [B20, B28, B29, C20,
C29, C30, C31, C33, H26, K22, L16, L23, L33, L34]). The
main difference between the approaches was in the choice of
exposure–response model to be used for each of the cellular
processes in the biologically based model. Age-dependent
cancer incidence data do not yet provide the basis for determining the most appropriate model, as evidenced by resultsof a model comparison using lung cancer data for radonexposed rats [H26]. The BEIR VI Committee [C20], as noted
earlier, acknowledged the importance of biologically basedmodels and indicated that, when biological mechanisms are
better understood, such models might become the preferredapproach to assessing the risks. Similar views are expressedby Krewski et al. [K22], who applied the two-stage clonal
expansion (TSCE) model to two cohorts important to assessing the risks from RDP exposure, namely the Colorado
Plateau miners and the Chinese tin miners.
505. An important project, within the framework of theEuropean Union project FIGH-CT1999-00013, looked ata multistage model analysis of the data from the French
and Czech uranium miner cohorts [T30]. One of the issues
being investigated was how risks can be transferred between

populations with different baseline lung cancer incidence
rates. The combined cohort includes 5,098 miners and
125 lung cancer deaths in the French cohort and 5,002 minersand 449 lung cancer deaths in the Czech cohort. The corresponding mean total cumulative exposure levels were 37 and57 WLM, respectively, for the two cohorts. A two-­mutation
carcinogenesis model with clonal expansion of cells in theintermediate stage was fitted to the individual miner data.
Linear exposure-effect relationships were used for the twomutational steps. The authors found that the f­itted linear effectof radon on the first mutational step was an order of magnitude larger than for the second mutational step. Although thebaseline lung cancer risk in the Czech miner cohort was considerably higher than that for the French ­miners, both data
sets could be described with the same parameter values forthe relative effect of exposure to RDPs on the mutation rates.
The authors argued that the uniform description of the effectof RDP exposure for two miner cohorts with distinctly different baseline lung cancer risks (0.09 for the French and 0.23
for the Czech miners) demonstrated the possibility of using
the model for risk transfer across populations. In addition,
the biologically based model implicitly describes age andexposure-rate effects, and thereby allows for extrapolation
to lifetime exposures to low radon concentrations. Lifetime
risks were calculated for a 75-year continuous exposure to
1 WLM/a (~256 Bq/m3). The lifetime ERR calculated fromthe model solution in references [B27, B28] was 1.1 for the
combined cohorts.

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506. Heidenreich et al. [H32] provided further analysis of
the French and Czech cohorts using the biologically based
TSCE model together with an analysis of the Chinese andColorado cohorts. The model allows an action of radiation
on initiation, promotion and transformation in cancer induction. While all four studies indicate a highly significant actionof radiation on promotion, the action on initiation is not significant in the French cohort and is barely significant in the
Colorado miner cohort. No action on transformation is foundin the Colorado miners, while the other data sets indicate aborder­line significance. The doubling exposure rate for initiation is about 3.5 WLM/a in the new data sets, while it is higher
in the historic data sets. For transformation, the doubling rate
is about 20 WLM/a for the new data sets, while again thehistoric data give higher estimates. The action of radiation on
promotion is different in the four data sets. The larger powerof the French and Czech cohorts requires less extrapolation
when the risk at very low exposures is estimated.
507. A 1990 study [M11] reanalysed the data for the Colorado Plateau uranium miners in order to investigate RDP
exposure, cigarette smoking and lung cancer using a twomutation model as the biological basis for their assessment.
The authors concluded that exposures to both RDPs andcigarette smoke affect the first mutational step and the rateof cell division but that the second mutational step was independent of RDP and cigarette smoke exposures. The authors
also concluded that the age-specific risks arising from joint
exposure to RDPs and cigarette smoke are more than additive
but less than multiplicative. An inverse dose-rate effect wasobserved, in that fractionally lower RDP exposure resulted
in higher lung cancer risks. The authors also obtained thesame estimates of lung cancer risk from exposure to residential radon for both smokers and non-smokers. This observation is consistent with observations from the European [D17]and North American [K1] pooled residential radon studies ofno significant heterogeneity across categories of smoking.
These studies showed an ERR of 0.2 per 100 WLM for nonsmokers at age 70 exposed at a rate of 0.2 WLM/a.
508. Luebeck et al. [L16] provided further discussion basedon an analysis of the data for the Colorado Plateau uranium
miner cohort using a TSCE model. Exposure to RDPs was
suggested to affect both the rate of initiation of intermediate
cells in the pathway to cancer and their rate of proliferation.
However, the effect of radon on the rate of initiation was
not statistically significant. The results of reference [L16]
showed that, depending on total radon exposure, the lifetime
ERR per unit exposure first increased with duration of exposure, reached a maximum and then declined. Non-smoking
miners who were exposed to RDPs for 10 years were foundto have approximately the same risk (as measured by lifetime ERR per unit exposure) as a non-smoking individual
who spent 10–20 years in a residence with very low levelsof radon. These authors stated that if the inverse dose-rate
effect observed in “miners at much higher total doses wereextrapolated naively to durations (and exposure rates) more
typical for homes, the risk (lifetime ERR per unit exposure)would be grossly overestimated.”

509. More recent studies such as those reported in references [B27, B28, H32, H36, K22, M11, T30] suggest thatmechanistic or biologically based models, which allow theopportunity to investigate mechanisms of carcinogenesis,
are likely to find increasing application. Tirmarche et al.
[T30], for example, reported estimates of final and second
mutation rates based on an analysis of the French and Czech
miner cohorts. One observation from their analysis is thatan inverse dose-rate effect cannot be excluded for exposure
rates of >30 WLM/a.
510. Harley et al. [H4, H36] applied a biologically basedmodel derived from an evaluation of the number of nuclei traversed by an alpha particle that will cycle as a function of time
following unit exposure to RDPs. The model was developedin two steps. First, the number of basal cell nuclei traversed byan alpha particle following an exposure of 1 WLM was determined from biological data. Secondly, the number of nuclei
traversed and the measured cycling rates in normal bronchial
epithelium were used to calculate the dividing population as
a function of time after exposure. The cycling rates decreased
with a half-time of about 15 years. The authors compared
their model fit with the combined excess risk data from thejoint analysis of 11 underground mining cohorts [L10], andsuggested that their model may explain why the tumour risk
decreases with time since exposure and with attained age, inthat both could reflect the reduction in cycling frequency ofcell nuclei repopulating the basal stem cells.
511. Little [L21] fitted multistage cancer models withclonal expansion to the Colorado miner data, allowing for up
to three mutational steps. Both radiation and smoking wereallowed to affect the mutation rates as well as cell proliferation in the intermediate stages. The best fit of the data was
obtained for a three-mutation model in which the first andsecond mutation rates increased with RDP exposure, and thefirst mutation rate increased with smoking rate, in a strongly
non-linear way. This three-mutation model was slightly
superior to a two-stage model in which the first mutation
rate depended in a non-linear fashion on radon and smoking,in combination with a reduction of intermediate cell death/differentiation rate with radon exposure.
512. The preceding examples show that current multistage
cancer models with clonal expansion are not specific enoughto determine the dose–response relationships for the cellular
processes in the model from a fit to cancer incidence data.
This was shown in model intercomparison on a large data
set of radon-exposed rats [H21]. An alternative approach tothis problem of specificity was applied in the two-mutation
model fit of Leenhouts to the Colorado uranium miner data
[L23]. He used a stepwise fitting approach in combination
with biologically motivated mutation equations. The radiation effect in this model was limited to the mutation rates,
and no effect of radiation on proliferation rates was assumed.
Using a stepwise approach, Leenhouts fitted the backgroundparameters to the lung cancer incidence of non-smokers, thesmoking coefficients to the lung cancer incidence of smokers, and the radiation parameters to the miner data. From

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

the model solution, the following risks were calculated atage 75 for a lifelong exposure to 0.1 WLM/a: EAR of 0.008
for non-smokers, increasing to 0.0014 for smokers, which­corresponds to ERR values of 0.19 and 0.009, respectively.
C.

Risk projection

513. The ICRP [I3] recommends that radon risk assessment be based on epidemiological studies of miners. The
BEIR VI Committee also base their risk estimates on an
analysis of pooled miner data [C20]. Both reports predate
the pooled analyses of residential case–control radon studies. As discussed below, the pressent report suggests that therisk of lung cancer from domestic exposure to radon can be
estimated using the results of the pooled residential case–control radon studies [D17, D21, K1, K26, L26].
514. It is important to develop reliable risk estimates oflung cancer due to radon exposure in the workplace or in
the home. Many published evaluations (e.g. [C20, U2, U17,W18]) are available, as described in previous sections.
Whether there is a risk from residential radon exposure has
been widely debated in the literature (e.g. [B34, C24, C25,
C34, F8, F12, H29, P10]). However, notwithstanding thewide range of results from residential case–control studies and the important effects of confounding by smoking
and other factors, overall the pooled European [D17, D21],
North American [K1, K26] and Chinese [L26] case–control
residential radon studies clearly demonstrate an association
between risk of lung cancer and residential radon exposure.
There is a remarkable coherence both among the pooled
residential studies and the downward extrapolation of radonrisk estimates from miner studies.
515. Uncertainties associated with downward extrapolation from miner studies include, but are not limited to: uncertainties in the reconstruction of miner exposures; possible
exposure to other carcinogens; the high but uncertain levelof smoking among miners; and the fact that exposures inmines historically were at relatively high levels comparedwith levels in homes and in present-day mines. Uncertainties associated with residential radon studies include, among
other factors: lack of contemporary radon measurements inresidences (e.g. while it may be possible to measure radonlevels today in homes previously occupied by subjects in residential studies, there will be large uncertainties associatedwith assumptions about changes to home ventilation and in
the habits of subjects over time); uncertainties about smoking; and low statistical power.
516. Many studies of lung cancer risks in miners exposed
to RDPs (see section IV) and several joint analyses of
11 miner cohorts have been published [C20, L4, L10].
Tirmarche et al. [T30] discussed risk estimates among thegeneral population and reported a comparison of different
models used to estimate lifetime risk. These authors estimated a lifetime ERR in the range 0.08–2.31, according to
the exposure scenario. These results are in the envelope of305

the results in BEIR VI. However, there are wide variationsin these studies in size, duration of follow-up, reliability of
exposure estimation, availability of smoking data and otherfactors. For example, it is questionable whether the following cohorts should be used for risk assessment: (a) Radium
Hill cohort, because of limited follow-up and the low quality of the exposure data; (b) Port Radium, Newfoundland
fluorspar and (without further re-evaluation using therevised dosimetry) Swedish iron miner cohorts, becauseof the low quality of the exposure data; and (c) Chinese
tin mining cohort, because of extreme confounding fromarsenic exposure. In addition, the French cohort data [C20,
L4, L10] are being updated, and the new Wismut cohort
described in section V, is being developed. The current
European project in which a joint analysis is being performed of the extended French cohort and the Czech cohort
is an important step, as these two cohorts are both of highquality. In addition, the Eldorado cohort, consisting ofminers at the Port Radium and Beaverlodge uranium mines
and the Port Hope uranium processing facility, has been
updated [H35]. One important aspect of this update is theextensive review of work histories and the use of updated
exposure algorithms, both of which significantly improve
the quality of the epidemiological studies carried out on
this cohort.
517. The BEIR VI model was developed from a pooled
analysis of 11 underground mining cohorts and takes
account of the reduction of relative risk with increasing timesince exposure, adjusting for attained age and exposure rate.
Estimates of lifetime risks developed with the model also
incorporated sex and smoking status. BEIR VI developed
two models, an exposure-age concentration model and an
exposure-age duration model. The general structure of the
BEIR VI model is illustrated below:
ERR = β [ θ5-14 w5-14 + θ5-14 w5-14 + θ25+ w25+ ] θage γzwhereβ=	slope of the exposure–risk relationship for the assumed reference categories of the modifying
factors.

θ5-14 θ15-24, θ25+

=	weighting factor for time periods
5–14, 15–24 and >25 years postexposure (values of θ5-14, etc., areprovided in reference [C20] andrange from 0.31 to 0.81).

w5-14, w15-24, w25+ =	radon exposures in the time windows 5–14, 15–24 and >25 years.
θage=	parameter to describe the decline
in ERR with increasing age.
BEIR VI [C20] reported values of
0.13–1.00.

γz=	parameter to describe the exposure-rate effect. BEIR VI [C20]reported values of 1.0–10.2.

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518. For exposure conditions in modern Saskatchewan
uranium mines, the power to detect any excess risk arisingfrom workplace exposure to radon is likely to be very small.
A study carried out for the Canadian Nuclear Safety Commission [S55] investigated the feasibility of conducting epidemiological studies of underground miners working under
today’s exposure conditions. To carry out this assessment,
two hypothetical cohorts were considered: a retrospective cohort of miners employed in Saskatchewan uranium
mines between 1975 and 2000, and a prospective cohortthat included miners employed from 2000 to 2030. The
cohorts were developed using demographic data provided
by the mining companies, dose data from the National Dose
Registry, and reference baseline cancer and mortality rates
from all of Canada and Saskatchewan. More than 50% ofmodern Saskatchewan uranium miners smoke. Adjustments
were made for different smoking prevalences in miners andthe reference population. In 2000, the average undergroundminer in northern Saskatchewan was exposed at work toabout 0.11 WLM from RDPs (with an upper 95% CI of about
0.43 WLM). At the same time, annual exposure to RDPs
at home was estimated to be about 0.4 WLM/a (ranging
upward to about 10 WLM/a). The feasibility analysis simulated incremental risks of lung cancer from radon exposureusing a relative risk model and an ERR of 0.89% WLM–1
based on the Ontario cohort, since it has provided the largest
and best exposure data of the Canadian uranium miner studies. Sensitivity analyses were carried out with respect to bothreference risk and exposures. Two statistical measures were
estimated using probabilistic simulation, standardized mortality ratio (SMR) and regression analysis, which produced
an estimate of excess lung cancer risks for a working levelmonth. Both cohorts were modelled to 2030. SMRs were
distributed on an average of 1.01, with most (80%) falling in
the range 0.93–1.08. Similarly, the slope estimated from theregression analysis had a mean of 1.01, with 80% of the slope
estimated to be in the range 0.083–0.105 WLM–1. This quite
large range reflected the statistical uncertainty estimated in
the cohorts. The results of the regression analyses showed
little probability (power) of detecting the predicted excess
risks in the cohorts, because the probability of the lower confidence level on the slopes exceeding zero is only about 3%(i.e. much less than the objective of 80%) for most combinations of scenario, cohort and follow-up period. Moreover,
exposure at home accounted for about 98% of the total for
modern Saskatchewan miners.

biological and physical mechanisms of carcinogenesis, given
the new information that has emerged from the pooling ofresidential radon case–control studies, the reconciliation of
the dosimetric and epidemiological evaluations is now muchless important for estimating risks from residential radon.
Experimental studies will continue to play an important rolein understanding the mechanisms and risks arising fromexposure to radon at work and at home, and biologically
based models will play an increasingly important role in the
future; however, at this time, the epidemiological studies ofminers and the residential case–control studies provide thestrongest basis for risk assessment.

519. There is now a great deal of information available
concerning the risks from exposure to radon and its decay
products. The studies of underground miners exposed tohigh levels of radon in the past have formed the principal
source of information on the risks, and serve as the basis fordeveloping exposure–response models and evaluating modifiers of effects such as time since exposure and age at exposure. A great deal of study has also been given to dosimetricevaluations, both as a means for transferring risk estimatesfrom a miner population to other circumstances and also intheir role as a source of ab initio risk estimates. While dosimetric evaluations remain important for understanding the− Information on all causes of death and cancer
incidence other than lung cancer.

520. In the past, radon risk estimates for residential exposures were based on downward extrapolation of evidence
from studies of miners who were exposed at higher exposures
for shorter times. Pooled analyses of European, North American and Chinese residential case–control studies provide
strong evidence supporting the evidence from miner studies
that exposure to high levels of radon and its decay products inhomes leads to an increased incidence of lung cancer.
521. Both the miner and the residential studies have advantages and disadvantages, some of which are briefly summarized below. The advantages of miner studies discussed in
section IV include:− Relatively high (relative to domestic) cumulative
exposures and exposure rates, which allowed thedevelopment of dose (exposure)–response relationships, at least at high cumulative exposures.
− The ability to examine factors that modify thesimple linear dose–effect relationship (time since
exposure, age at exposure, exposure rate).
− Information on risks over a lifetime, not just awindow of 30 years or so. Miner studies will
require continued follow-up to realize this potential to the full.
− Exposure estimates based on contemporary measurements (albeit incomplete, most often area
measurements rather than individual measurements and subject to uncertainty). While someevaluations of the effects of the uncertainties inthese measurements on the resulting risk estimates were carried out, more remains to be donein this area.

522. Many residential case–control studies (section V),
have been published. Individually, they have limited statistical power, and meta-analysis has suggested that theresults of the studies are inconsistent. However, more recent
pooled analyses, especially those of European [D17, D21]
and North American [K1, K26] studies, combined the dataon all the individuals in a number of residential studies andhave greater power than the individual component studies.
The pooled residential radon studies provide strong, direct

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

evidence of risk from residential radon. The pooled residential studies have certain advantages over the miner studies:− Exposures received under aerosol conditions similar to those of interest.
− Exposures received at concentrations similar to
those of interest (thus reducing the need to extrapolate from the relatively high rates of exposure inmines).
− Reduced confounding from possible exposure tooccupational carcinogens such as arsenic.
− Detailed individual smoking histories for study
participants.
− Detailed individual exposure data based on measurements in the homes where the individuals had
lived. Nonetheless, exposure uncertainties arealso part of residential studies. For example, themeasurements were usually made some time after
the period over which the risk was to be assessed,
and in some instances, there were alterations to
the home between the time when the resident
lived there and the time when the measurement
was made. Data are available regarding the uncertainties in the assessment of residential radon

307

exposures, and the European pooled analysis, inparticular, carried out calculations quantifying theeffect of this uncertainty on the risk estimates.
− Data for both men and women for a wide variety
of ages.
523. Notwithstanding the strengths and weaknesses of risk
estimates from studies of miners and of residential radon,
there is now a remarkable coherence between the risk estimates developed from epidemiological studies of miners andpooled residential case–control radon studies. While both theminer studies and the residential case–control radon studies are subject to various limitations arising from exposure
uncertainty and confounding by smoking, for example, bothtypes of study are suitable for risk estimation. The minerstudies provide a strong basis for evaluating risks from RDP
exposure to people at work and at home, and for investigating the effects of modifiers to the exposure–response relationship [S2]. However, the results of the pooled residential
studies now provide a direct method of estimating risks topeople at home without the need for (downward) extrapolation from miner studies. The measurement-adjusted risk
coefficients reported, for example from the European pooling study, provide an appropriate basis for estimating risks topeople at home.

VIII. Overall Conclusions
524. This annex, “Sources-to-effects assessment for radonin homes and workplaces”, discusses: potential sources of
exposure of workers and the public to radon; issues of current interest in the dosimetry of radon and its decay products; information from animal experiments and experiments
at the cellular and subcellular levels that are important inunderstanding the mechanisms of carcinogenesis; epidemiological studies of miner and residential exposure to radon;
and approaches to risk projection.

to underground miners (derived from updated studies ofcohorts of uranium miners) suggest that the two approaches
are less different than initially thought. The Committee
therefore continues to recommend a radon dose conversion factor of 9 nSv (Bq h m–3)–1 to evaluate the effective
dose from radon inhalation. The dose conversion convention recommended in ICRP Publication 65 [I2] is approximately 30% lower than this factor but the difference is notconsidered significant.

525. During daily life, everyone is exposed to radon, an
inert radioactive gas that occurs naturally and is present
everywhere in the atmosphere. The levels of radon indoors
vary widely both within countries and between countries,
with (nominal) geometric mean concentrations of radon in
air indoors ranging from less than 10 Bq/m3 in the Middle
East to up to around 100 Bq/m3 in a number of European
countries.

529. Studies of miners exposed to radon and its decay
products provide a direct basis for assessing lung cancer risk. The United States National Research Council’s
6th Committee on the Biological Effects of Ionizing Radiation (BEIR VI) [C20] reported an excess relative risk fromexposure to radon that was equivalent to 1.8% (95% CI:
0.3%, 35%) (MBq h m–3)–1 for miners with cumulative exposures of below 30 MBq h m–3. There are various sources oferror in the exposure assessment of miners, especially in theearliest years of mining, when exposures were at their highest. Other factors that complicate the analysis of data on
miners include: the high percentage of miners who smoke;
workplace exposure to dust contaminants, such as arsenic,
diesel exhaust in the dust and other pollutants; and periods
spent working in non-uranium mines. The power to detect
any excess risks due to the exposures that miners nowadays
receive is likely to be small, in part because the exposures
are much smaller than those in the early years of mining.
Because of the high exposures in the early days of mining,
it is possible to detect trends in lung cancer incidence andto investigate factors that affect the exposure–response relationship, such as the age at exposure, the effect of exposure
rate and the reduction of risk with increasing time since
exposure, as well as the effect of confounding factors suchas smoking.

526. The annual per caput dose from inhalation of radongas and its decay products represents typically about onehalf of the effective dose received by members of the publicfrom all natural sources of ionizing radiation. For certain
occupations, radon gas is the predominant source of occupational radiation exposure.
527. Radon decay products are well established as lungcarcinogens. However, the doses to other organs and tissues
arising from the inhalation of radon and its decay products
are quite small, usually at least an order of magnitude smaller
than those to the lung. Moreover, epidemiological data provide little support for increased risks of mortality other thanfrom lung cancer.
528. A factor for calculating the dose from a given exposure to radon and its decay products is needed for risk
management, including regulatory purposes, and to allow
comparison with other sources of radiation exposure. There
are two approaches for deriving a dose conversion factor.
A “dosimetric approach” derives the dose from a given
exposure based on the deposition characteristics of radon
decay products in the respiratory tract. An “epidemiological approach” was used by the International Commission
on Radiological Protection (ICRP) to derive the dose conversion factor from epidemiological studies using the ratioof the risk of lung cancer in miners to the overall risk ofcancer in the survivors of the atomic bombings in Japan.
In the UNSCEAR 2000 Report, there was a difference ofabout a factor of 2 between the two approaches. However,
the most recent data that have been published on the risks

530. The BEIR VI model developed from the pooled
analysis of 11 cohorts of underground miners provides awell-established basis for estimating risks from occupational exposures to radon, and accounts for factors such as
the reduced risk with increasing time since exposure. Since
the BEIR VI report was published, studies of various minercohorts have been updated, and confirm the general patternsof risk with dose and with time since exposure that werereported by BEIR VI. They also provide updated coefficients
to take account of the effects of time since exposure on ageing populations. Miner studies therefore provide a strong
basis for evaluating risks from exposure to radon and forinvestigating the effects of modifiers to the dose–response
relationship.
309

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531. Biological and cellular models of the multistage­process of carcinogenesis are used to analyse the data fromstudies on miners, and offer the possibility to assess uncertainties in our understanding of the mechanisms for thedevelopment of cancer and their modelling for the purposesof risk estimation.
532. The extrapolation of radon concentrations in the air inmines to those in homes provides an indirect basis for assessing the risks from residential exposure to radon. However,there are now over 20 analytical studies of residential radon
exposure and lung cancer. These studies typically assess therelative risk from exposure to radon on the basis of estimatesof residential exposure over a period of 25–30 years prior
to the diagnosis of lung cancer. More recent pooled analyses of residential case–control studies support a small but
detectable lung cancer risk from residential exposure, andthis risk increases with increasing radon concentrations. The
excess relative risk of lung cancer from long-term residential
exposure to a radon concentration of 100 Bq/m3 is established with reasonably good precision and is considered to
be about 16% for both smokers and non-smokers (after correction for uncertainties in the exposure assessment), with anuncertainty of about a factor of 3 higher or lower than thisvalue. Because the baseline lung cancer rate for smokers ismuch higher than that for non-smokers, smokers account fornearly 90% of the population risk of lung cancer.
533. Although there are major uncertainties in extrapolating the risks of exposure to radon from the miner studies
in order to assess the risks in the home, there is nevertheless remarkably good agreement between the risk factors
derived from the miner studies and from the pooled residential case–control studies. The ERR per unit radon concentration in air estimated in this annex from miner studies is0.12 (95% CI: 0.04, 0.2) per 100 Bq/m3 (see para. 424); thatfrom the pooled residential case–control studies (based on
the restricted analysis) for Europe is 0.094 (95% CI: 0.034,
0.175) per 100 Bq/m3 [D17, D21] and for North America
is 0.18 (95% CI: 0.02, 0.43) [K1, K26] per 100 Bq/m3 (seetable 24). The studies of uranium miners also provide important information on the effects of modifiers to the exposure–
response relationship, and further follow-up is encouraged.
The pooling of residential case–control studies in Europe,
North America and China now provides an appropriate
basis for estimating the risks from long-term residential
exposure to radon. On the basis of current information, theCommittee considers the use of measurement-adjusted risk
coefficients from pooled studies as an appropriate basis for
estimating the risks to people at home.

ANNEX E: SOURCES-TO-EFFECTS ASSESSMENT FOR RADON IN HOMES AND WORKPLACES

Appendix
Quantities, units and conversion factors relevant to radon and its decay products
Term

Definition

Unit and Selected Conversions

Absorbed dose

The energy absorbed through exposure to radiation divided by
the mass of the body or by the mass of the part of the body thatabsorbs the radiation.

Gy
1 Gy = 1 J/kg = 100 rad

Effective dose

The sum over all tissues and organs of the equivalent doses
weighted by the tissue weighting factor, which represents thecontribution of that organ or tissue to the total detriment resulting
from uniform irradiation of the whole body.

Sv
1 Sv = 100 rem

Equilibrium equivalentconcentration

The concentration of radon in air, in equilibrium with its short-lived
decay products, which would have the same potential alpha energy
concentration as the existing non-equilibrium mixture.

Bq/m3
1 Bq/m3 = 5.56 × 10–9 J/m3

Equilibrium equivalent exposure

Time integral of the corresponding equilibrium equivalent concentration of radon to which the individual is exposed over a given time
period.

Bq h m–3
1 Bq h m–3 = 5.56 × 10–9 J h m–3
1 Bq h m–3 = 1.57 × 10–6 WLM

Mache unit

A measure of radon concentration (historically radium emanation),where 1000 Mache units equals the amount in equilibrium with
1/2000 mg of radium.

Mache unit
1 Mache unit = 275 pCi/L

Potential alpha energy
concentration

The concentration of short-lived radon decay products in air interms of the alpha energy released during complete decay throughpolonium-214.

J/m3
1 J/m3 = 1.80 × 108 Bq/m3

Potential alpha energy exposure

Time integral of the potential alpha energy concentration in air towhich the individual is exposed over a given time period.

J h m–3
1 J h m–3 = 1.80 × 108 Bq h m–3
1 J h m-3 = 282 WLM

Radon levels

Radon concentration in air.

Relative risk coefficient

The ratio of the risk in an exposed population to that in a similar
unexposed population per unit exposure.

—

Working level

Any combination of the short-lived decay products of radon in
one litre of air that will result in the emission of 1.3 × 105 MeV ofpotential alpha energy.

WL
1 WL = 100 pCi/L (assuming
100% equilibrium, i.e. F = 1)
1 WL = 250 pCi/L (assuming
40% equilibrium, i.e. F = 0.4)

Working level month

The cumulative exposure from breathing an atmosphere at aconcentration of 1 WL for a working month of 170 hours.

WLM
1 WLM = 3.54 × 10–3 J h m–3
1 WLM = 6.38 × 105 Bq h m–3
1 WLM = WL × exposure time(h/a)/(170 h/WLM)

pCi/L
1 pCi/L = 0.037 Bq/L
1 pCi/L = 37 Bq/m3

311

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313

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