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QUEENSLAND UNIVERSITY OF TECHNOLOGY
SCIENCE AND ENGINEERING FACULTY
Measurement and Monitoring of
Naturally Occurring Radioactive
Materials for Regulation
Submitted by Sami Alharbi (B.App.Sc, M.App.Sc) to the Science and Engineering Faculty,
Queensland University of Technology, in fulfilment of the degree of Doctor of
Philosophy
2016
ii
Abstract
Naturally occurring radioactive material (NORM) is a commonly used term to describe
material containing primordial radionuclides. Of these, the radioisotopes in the uranium,
actinium and thorium decay series and potassium-40 are important for radiation protection
considerations. Radiation exposure due to NORM has the potential to cause radiological harm to
humans and the environment. Scientific studies have confirmed the association between
inhalation of short-lived radioactive progeny of radon gas and lung cancer. Radon (Rn) is
released from surfaces of materials containing radium isotopes and accumulated in enclosed
spaces. The measurement of the releases of radon as well as its level in indoor air and the
activity concentration of primordial radionuclides in materials are of particular interest for
controlling the resultant exposure. This thesis investigates such measurements, with an
emphasis on the regulation of NORM.
NORM had been largely unregulated in non-nuclear industries until the mid-1980s. The
current regulation of NORM derives from that developed for artificial radionuclides. However,
exposure pathways and scenarios for NORM differ from that for artificial radionuclides, and
consequently, the implementation of such regulation can create confusion among regulatory
authorities. A comprehensive review of recent developments in the radiation protection system
and the current regulation of NORM has been conducted as part of the present thesis. This
review identifies and discusses the challenges that regulatory bodies and industries may face
when implementing NORM regulation. The review leads to the conclusion that the NORM
industry has to be managed on a case-by-case basis because each industry may have its own
exposure characteristics. The implications of the implementation of newly proposed ICRP and
IAEA approaches for radon regulation are also discussed. The regulation of NORM requires
adequate measurements with sufficient accuracy and reliability. As NORM is a vast subject,
three measurement areas were selected because of their importance in the context of radiation
protection: (i) the measurement of radon activity flux density using the activated charcoal
canister technique; (ii) the concentration of radon and thoron in indoor air; and (iii) the analysis
of NORM in solid materials using a broad energy germanium (BEGe) detector.
The measurement methodology of radon activity flux density using an activated
charcoal canister was revisited. The results showed that the vertical distribution of the
adsorbed 222Rn in the charcoal bed of exposed canisters was not the same as in standard
canisters prepared in the laboratory for calibrating measurement instruments. The distribution
also varied according to exposure times. Another observation was the reduction of the
measured 222Rn activity flux with the exposure time. This reduction was attributed to the back
diffusion effect, where 222Rn escaped from the canister through a canister-soil interface and
iii
possibly through the adsorption of water molecules by the activated charcoal. These factors
place limitations on the accuracy of the activated charcoal canister technique.
The activity concentrations of radon (222Rn) and thoron (220Rn) in various workplaces in
Brisbane, Australia, were measured. Basements and confined areas were found to host higher
222Rn activity concentration, most likely because of poor ventilation systems in such locations.
For 220Rn, the maximum activity concentration was detected in storage rooms. Diurnal variation
in 222Rn concentration was observed in workplaces—low during working hours because of the
use of air conditioners and natural ventilation during this time. This variation was observed less
frequently in the case of 220Rn as the change in its activity concentration during and after
working hours was insignificant. The influence of a number of parameters on indoor 222Rn and
220Rn concentrations was also investigated in this study. Floor levels and flooring types were
found to have the greatest impact on indoor 222Rn levels, while distance from walls and surfaces
have been found to be the parameters that affected indoor 220Rn level the most.
The use of the BEGe detector for the analysis of NORM was also explored in this thesis.
The calibration procedure and sample preparation, as well as the required correction factors,
were described. The performance of the detector was tested using two statistical parameters:
relative bias and u-score. The results of these tests revealed no significant difference between
the measured and the reference activity concentration values. In comparison with a
conventional HPGe coaxial detector, the BEGe detector was superior in detecting peaks at low
gamma energy and intensity, such as 46.5, 50.1, 63.3 and 67.7 keV, of 210Pb, 227Th, 234Th and
230Th, respectively. In addition, a rapid technique of determining the activity concentration of
226Ra through the 186.1 keV peak was successfully performed and compared with the
traditional method, which is through 214Pb and 214Bi peaks. Good agreement was reported for
most samples. This method was limited by the large variations in activity concentration and
density between samples and the calibration standard. Overall, the detector was found to be
suitable for the application of NORM.
The findings presented in this thesis advance the understanding of current measurement
techniques and how they should be carried out to adequately regulate NORM. The review on the
current regulation of NORM identified key challenges that are worth addressing in order to
promote its management.
Keywords: NORM, regulation, recommendation, reference level, measurement, radionuclide,
radon, flux density, activity concentration, equilibrium
iv
Acknowledgements
I would like to thank God for giving me the strength to complete this work. I would like to thank
King Saud bin Abdulaziz University for Health Sciences for granting me a full scholarship to
pursue my higher education.
I wish to thank various people for their contribution to this thesis. First of all I would like to
express my very great appreciation to my principal supervisor, Dr Riaz Akber for his enduring
support and guidance. I would like to extend my thank you to my associate supervisor Dr Jamie
Trapp and to the Radiological laboratory technician Kazuyuki Hosokawa for their assistance
throughout my PhD journey.
Finally, I wish to thank my parents, siblings, friends and my wife, who has never once
complained when I have turned up late, for all the generous support and encouragement.
v
Contents
Abstract ................................................................................................................................................................................. ii Acknowledgements ......................................................................................................................................................... iv Contents .................................................................................................................................................................................v List of Figures .................................................................................................................................................................. viii List of Tables ...................................................................................................................................................................... ix List of Publications ............................................................................................................................................................x
Chapter 1 Introduction .................................................................................................................................................... 1
1.1 Naturally occurring radioactive material (NORM) ....................................................................................... 1
1.2 Research objectives ...................................................................................................................................... 2
1.3 Structure of the research .............................................................................................................................. 3 Chapter 2 Literature Review ......................................................................................................................................... 5
2.1 Introduction .................................................................................................................................................. 5
2.2 Origin and distribution ................................................................................................................................ 5
2.3 Terminology ................................................................................................................................................. 6
2.4 Behaviour of NORM ...................................................................................................................................... 8 2.4.1 Uranium ................................................................................................................................................ 8 2.4.2 Thorium ................................................................................................................................................. 8 2.4.3 Radium .................................................................................................................................................. 9 2.4.4 Radon .................................................................................................................................................... 9 2.4.5 Polonium ............................................................................................................................................. 10 2.4.6 Lead ..................................................................................................................................................... 10
2.5 Exposure to NORM ..................................................................................................................................... 11 2.5.1 Internal exposure .............................................................................................................................. 11 2.5.2 External exposure ............................................................................................................................. 12 2.5.3 Natural background radiation .......................................................................................................... 13
2.6 Regulation of NORM .................................................................................................................................. 13
2.7 Measurement of NORM .............................................................................................................................. 16 2.7.1 Measurement of radon releases ....................................................................................................... 17 2.7.2 Measurement of atmospheric radon concentration ....................................................................... 19 2.7.3 Measurement of NORM in solid media ............................................................................................ 20
2.8 Measurement challenges ........................................................................................................................... 22 2.8.1 Measurement cost ............................................................................................................................. 22 2.8.2 Uncertainties ..................................................................................................................................... 24 2.8.3 The background................................................................................................................................. 25 2.8.4 The equilibrium state ........................................................................................................................ 26
2.9 Conclusion .................................................................................................................................................. 28 References ........................................................................................................................................................ 30
Chapter 3 NORM Regulation: Better understanding for better management .......................................... 51
Abstract ............................................................................................................................................................ 53
vi
3.1 Introduction ................................................................................................................................................ 53
3.2 Overview of the radiological protection system ......................................................................................... 54 3.2.1 A brief history of radiological protection ............................................................................................ 54 3.2.2 Quantitative recommendations .......................................................................................................... 55 3.2.3 The conceptual framework ................................................................................................................. 56 3.2.4 Protection beyond humans ................................................................................................................. 57
3.3 The current system of radiological protection ............................................................................................ 57 3.3.1 Exposure situations ............................................................................................................................. 59 3.3.2 Principles ............................................................................................................................................. 60 3.3.3 Exposure categories ............................................................................................................................ 60
3.4 NORM in the radiation protection context ................................................................................................. 60 3.4.1 Radon in dwellings and workplaces .................................................................................................... 62 3.4.2 NORM industries ................................................................................................................................. 62
3.5 Regulatory issues with NORM .................................................................................................................... 63 3.5.1 Absence of a complete framework ..................................................................................................... 64 3.5.2 The interference between planned and existing exposure situations ................................................ 65 3.5.3 The lack of a homogeneous approach ................................................................................................ 65 3.5.4 Ubiquity of NORM ............................................................................................................................... 66 3.5.5 Source of low dose .............................................................................................................................. 67 3.5.6 Societal attitudes ................................................................................................................................ 68 3.5.7 Measurement challenges .................................................................................................................... 68 3.5.8 Waste and residues ............................................................................................................................. 69 3.5.9 Ethical approach ................................................................................................................................. 70
3.6 Case study .................................................................................................................................................. 71 3.6.1 Indoor 222Rn regulation ....................................................................................................................... 72 3.6.2 New policies and recommendations on indoor 222Rn ......................................................................... 73 3.6.3 Implications of the implementation of the new policies .................................................................... 74 3.6.4 The case of indoor 220Rn ..................................................................................................................... 76
3.7 Conclusion .................................................................................................................................................. 76 References ........................................................................................................................................................................................... 78 Chapter 4 Radon-222 Activity Flux Measurement Using Activated Charcoal Canisters: Revisiting
the Methodology .............................................................................................................................................................. 85
Abstract ............................................................................................................................................................ 87
4.1 Introduction ................................................................................................................................................ 87
4.2 Methods ..................................................................................................................................................... 89 4.2.1 The distribution of 222Rn in the charcoal bed ...................................................................................... 89 4.2.2 The influence of exposure time on the measured activity flux........................................................... 89 4.2.3 Analytical technique ........................................................................................................................... 89
4.3 Results and Discussion ................................................................................................................................ 91 4.3.1 The distribution of 222Rn in the charcoal bed ...................................................................................... 91
4.3.1.1 The uniformity of 222Rn in the charcoal bed ............................................................................ 91 4.3.1.2 Detector efficiency at different distance .................................................................................. 92 4.3.1.3 Canisters with different depth of activated charcoal .................................................................. 93
4.3.2 The influence of exposure time on the measured 222Rn activity flux ................................................. 94 4.4 Conclusion .................................................................................................................................................. 99 References ...................................................................................................................................................... 101
vii
Chapter 5 Radon and thoron concentrations in public workplaces in Brisbane, Australia ............. 103
Abstract .......................................................................................................................................................... 105
5.1 Introduction .............................................................................................................................................. 105
5.2 Materials and Methods ............................................................................................................................ 108 5.2.1 Sampling site ..................................................................................................................................... 108 5.2.2 Radon and thoron measurements .................................................................................................... 109
5.3 Results and Discussion .............................................................................................................................. 111 5.3.1 Radon and thoron concentration measurements............................................................................. 111 5.3.2 Diurnal variation ............................................................................................................................... 115 5.3.3 The influence of the measurement heights ...................................................................................... 117 5.3.4 The influence of floor levels .............................................................................................................. 118 5.3.5 The role of flooring on indoor radon and thoron concentrations .................................................... 119
5.4 Conclusion ................................................................................................................................................ 119 References ...................................................................................................................................................... 121
Chapter 6 A broad energy germanium detector for routine and rapid analysis of naturally
occurring radioactive materials ............................................................................................................................. 124
Abstract .......................................................................................................................................................... 126
6.1 Introduction .............................................................................................................................................. 126 6.1.1 Routine analysis of NORM ................................................................................................................ 127 6.1.2 Rapid measurements of 226Ra ........................................................................................................... 127 6.1.3 Determination of 227Ac using gamma-ray spectrometry .................................................................. 128
6.2 Material and methods .............................................................................................................................. 129 6.2.1 Routine analysis of NORM ................................................................................................................ 129 6.2.2 Rapid measurements of 226Ra ........................................................................................................... 133 6.2.3 Determination of 227Ac using gamma ray spectrometry ................................................................... 133 6.2.4 Sample preparation .......................................................................................................................... 134 6.2.5 Analytical technique ......................................................................................................................... 135 6.2.6 Gamma self-attenuation ................................................................................................................... 137 6.2.7 Coincidence summing ....................................................................................................................... 138 6.2.8 Uncertainties ..................................................................................................................................... 139
6.3 Results and discussion .............................................................................................................................. 140 6.3.1 Routine analysis of NORM ................................................................................................................ 140 6.3.2 Rapid measurements of 226Ra ........................................................................................................... 146 6.3.3 Determination of 227Ac using gamma ray spectrometry ................................................................... 149
6.4 Conclusion ................................................................................................................................................ 151 References ...................................................................................................................................................... 153
Chapter 7 Summary and Conclusion ..................................................................................................................... 157 7.1 Conclusion ................................................................................................................................................ 157 7.2 Future Direction ........................................................................................................................................ 159
viii
List of Figures
Figure 2.1 Uranium and thorium decay series ..................................................................................................... 7
Figure 3.1 The basis of radiological protection regulation. ......................................................................... 58
Figure 3.2 The radiological protection system. ................................................................................................. 58
Figure 3.3. Diagram of the current regulation of NORM. .............................................................................. 61
Figure 4.1 The distribution of 222Rn in the charcoal bed for different exposure times. The
correction factors corresponding to the exposure time are shown in parentheses. ......................... 92
Figure 4.2 Normalised 222Rn activity flux in open-trays plotted against exposure time. ................ 95
Figure 4.3 The measured activity flux versus exposure time for charcoal canister fitted over cans
containing samples. ...................................................................................................................................................... 96
Figure 4.4 The measured activity flux of the samples in the closed-can relative to RAD7 values.
............................................................................................................................................................................................... 97
Figure 4.5 Relative count rate in the charcoal canisters versus the exposure time of samples
taken in a closed-can or open-tray. The dashed line represents a situation referred by Equation
4.3. ........................................................................................................................................................................................ 99
Figure 5.1. A map of the sampling site. ............................................................................................................... 109
Figure 5.2 A box plot showing radon and thoron concentrations across workplaces. The boxes
represent 25–75% of the values with median concentrations represented by the lines inside the
boxes. Whiskers represent the range of the distribution and the dots outside the range of 95%
indicate values as outliers. ....................................................................................................................................... 113
Figure 5.3 Correlation between radon and thoron concentrations ........................................................ 113
Figure 5.4 Histogram showing the distribution of the natural logarithms of radon and thoron
concentrations in workplaces with a normal distribution fitted to the data. Geometric means of
radon and thoron concentrations were 7.4 and 5.8 Bq m-3, with a geometric standard deviation
of 2.2 and 2.0 Bq m-3, respectively. ....................................................................................................................... 114
Figure 5.5 The variation in radon and thoron concentrations during the day. ................................. 115
Figure 5.6 Normalised radon and thoron concentrations during and after work hours. .............. 116
Figure 5.7 Radon and thoron concentrations in the air at 5 cm and 120 cm above the floor for
each sampling site. ...................................................................................................................................................... 117
Figure 5.8 Box-whisker plot showing the overall radon and thoron concentrations at the ground
and breathing zones for all location. ................................................................................................................... 117
Figure 5.9 Box-whisker plot showing radon and thoron concentrations categorised by the floor-
level. .................................................................................................................................................................................. 118
Figure 5.10. Box-whisker plot showing A) radon and B) thoron concentrations characterised by
the flooring type. .......................................................................................................................................................... 119
Figure 6.1 (A) Count per emission versus the gamma energy for the reference materials using
the BEGe detector. (B) A comparison between the detection efficiency of a BEGe detector and an
HPGe coaxial detector. The results are for a calibration standard in petri dish geometry (50 g) in
BEGe and Marinelli Beaker geometry (1000 g) in HPGe-coaxial. ............................................................ 137
Figure 6.2 (A) The measured activity concentration (dots) compared to the certified values
(continuous line) for radioisotopes in the reference materials used in the present study, RGU-1,
RGTh-1, and IAEA-434. The comparison is for uranium, thorium, and actinium radioisotopes.
The uncertainty (2) in the certified values is given as dashed lines and in the measured values
as error bars. (B) Activity ratios of several radionuclides in the uranium and thorium series for
the spiked samples (SS and SG) prepared from the NBL standards. The continuous line
ix
represents a ratio of 1:00 and the error bars are the uncertainty to 2. (C) A comparison of 40K
measured activity concentrations and referenced values for the reference samples SS, SG, and
RGK-1 ................................................................................................................................................................................ 142
Figure 6.3 (A) The 226Ra activity concentration measured in the IAEA standards, RGU-1 and
IAEA-434, using (a) 222Rn progeny gamma lines and the 186 keV peak after correcting it for the 235U contribution, calculated from 235U individual gamma lines at (b) 143, (c) 163, (d) 205 keV,
(e) the combined average from these lines, and (f) from the 238U/235U ratio compared to the
certified values. (B) The 226Ra activity measured using different approaches compared to the
expected values in the spiked samples . ............................................................................................................. 147
Figure 6.4 The activity concentrations of 227Ac versus 235U in various samples examined during
the present study. For most samples, the values lie along the line of equality. Samples such as TI,
TII, and BC show the 227Ac:235U ratio 1. This behaviour is most likely due to extraction of
uranium from these samples. ................................................................................................................................. 151
List of Tables
Table 2.1 Examples of radionuclides in equilibrium ...................................................................................... 26
Table 3.1 The development of dose conversion factor over time ............................................................. 73
Table 4.1 Rn-222 emitting samples used in the present study. ................................................................. 90
Table 4.2 Correction factors of 222Rn activity flux for different charcoal depths in a canister with
respect to that of 2.5 cm depth for one day exposure time. ......................................................................... 94
Table 5.1 Summary of the instrument calibration check using certified sources. ........................... 111
Table 5.2 Radon and thoron concentrations in different locations. ....................................................... 112
Table 6.1 Gamma energy lines of key natural radionuclides (Ekström and Firestone, 2004) .... 132
Table 6.2 Sample descriptions ............................................................................................................................... 135
Table 6.3 Correction factors for self-absorption (fs) and random summing (fR); material
densities (g cm-3) are given between the brackets. A brief description of samples is given in
Table. 2. ............................................................................................................................................................................ 139
Table 6.4 U-scores and relative bias for reference materials .................................................................... 142
Table 6.5 The measured activity concentration of several radionuclides of NORM (Bq kg-1) in
various samples used in the present study. ...................................................................................................... 145
Table 6.6 Ratios of 226Ra activity as measured using different approaches to the ratio obtained
via its progeny............................................................................................................................................................... 149
Table 6.7 Activity ratios of 227Ac to 235U, 238U and 226Ra .............................................................................. 150
x
List of Publications
Alharbi, Sami H and Riaz A Akber. 2014. "Radon-222 activity flux measurement using
activated charcoal canisters: Revisiting the methodology." Journal of
Environmental Radioactivity 129: 94-99.
Alharbi, Sami H and Riaz A Akber. 2015. "Radon and thoron concentrations in public
workplaces in Brisbane, Australia." Journal of Environmental Radioactivity 144:
69-76.
Alharbi, Sami H and Riaz A Akber. 2016. " A broad energy germanium detector for
routine and rapid analysis of naturally occurring radioactive materials." Journal
of Radioanalytical and Nuclear Chemistry. (In Press)
Alharbi, Sami H and Riaz A Akber. 2016. " NORM regulation: better understanding for
better management." (In Preparation)
xi
Statement of Original Authorship
The work contained in this thesis has not been previously submitted to meet
requirements for an award at this or any other higher education institution. To the best
of my knowledge and belief, the thesis contains no material previously published or
written by another person except where due reference is made.
Signature
Date:………………………………
QUT Verified Signature
1
Chapter 1 Introduction
1.1 Naturally occurring radioactive material (NORM)
Naturally occurring radioactive material (NORM) is a widely used term to describe
materials that contain radionuclides of natural origin. The term is usually applied to refer to the
primordial radionuclides 40K, 238U, 235U, 232Th and their radioactive decay products. These
radionuclides are ubiquitous in the environment and they present, with a wide variation of
concentration levels, in raw materials, minerals, products, by-products, equipment, waste and
residues, as well as in many industrial processes. Individual radionuclides of NORM have many
applications in various disciplines, such as environmental science, radiometric dating,
atmospheric studies and erosion and sedimentation studies. Exposure to ionising radiation
emitted by individual radionuclides of NORM can be hazardous to humans and the environment.
The radiation dose rate due to exposure to NORM is generally low in most environments.
However, certain human activities can increase the concentration of NORM and/or alter
exposure conditions. Subsequently, this can give rise to above background radiation dose to
receivers. Such exposure needs to be controlled through regulation to ensure that adequate
protection is given to individuals as well as to the environment.
Since the establishment of radiological protection bodies, most of the regulatory
attention to natural sources has been largely focused on the uranium mining and milling
industry, as it is part of the nuclear fuel cycle. However, over the past three decades, there has
been a growing awareness arising from the recognition of increased levels of NORM in non-
nuclear industries, such as phosphate and petroleum industries. The control of exposure
encompasses both indoor and outdoor environments, whether in dwellings or in workplaces,
including industries involving NORM.
Quantifying the activity levels of individual radionuclides of NORM plays an essential
role in managing radiation exposure. The establishment, implementation and compliance with
any radiological protection regime require accurate measurement. In fact, the measurement of
activity concentration is a prerequisite at all stages of a regulation system, starting from the
screening of radionuclides to periodic monitoring checks. It is also essential for the assessment
of radiation dose received by a certain group of human or non-human biota. However, the
measurement of NORM is associated with difficulties and limitations because of the wider range
of radionuclides involved in great variations in their physical and chemical properties.
Over the past century, radiation protection principles and values have evolved. Any new
radiation protection regulations, standards and monitoring regimes for the protection of people
and the environment from NORM should be developed with the latest knowledge in mind. Old
2
monitoring regimes may also need to be reviewed in light of these new developments. National
regulatory bodies may face challenges related to the implementation of recent radiological
protection concepts in their current system of regulation. A review of the regulation of NORM
would contribute to understand the appropriate ways to control it. On the other hand,
measurements for regulation purposes require standardised methods to ensure that specific
criteria have been met. More research studies are required to develop an understanding of
better ways to manage exposure to NORM and how the measurement of NORM should be
conducted to fulfil regulatory requirements.
1.2 Research objectives
The overall aim of this research project is to develop an understanding of the
measurement, monitoring and regulation of NORM. This aim has a number of subsidiary aims:
- To develop an understanding of the mechanism involved in 222Rn absorption in
activated charcoal and, on that basis, to develop an adequate methodology for its use in
222Rn exhalation measured from materials elevated in uranium series activity
concentrations
- To study the distribution of 222Rn and 220Rn activity concentrations in indoor
environments
- To investigate the suitability of a broad energy germanium (BEGe) gamma spectroscopy
system for quantifying the activity concentration of individual radionuclides of NORM.
The stated aims are accomplished through a set of objectives and sub-objectives:
The first objective is to comprehensively review the latest developments in radiation
protection regarding NORM and then identify the main regulatory issues in establishing,
implementing and complying with radiation protection regimes. The publications to be
considered by this review include those from the International Commission on
Radiological Protection (ICRP), the International Atomic Energy Agency (IAEA) and the
United Nations Scientific Committee on the Effects of Atomic Radiation (UNSCEAR), as
well as those from several nationally-based organisations, such as the Australian
Radiation Protection and Nuclear Safety Agency (ARPANSA) and the Queensland
Radiation Health Unit (as the state-based regulator for radiation protection)
The second objective is to further our knowledge about the measurement of NORM to
demonstrate compliance with national and international requirements for radiation
protection. As NORM is a vast subject, three areas have been selected for investigation:
(1) 222Rn exhalation rate using activated charcoal canisters
(2) Atmospheric 222Rn and 220Rn in workplaces in Brisbane, Australia
(3) The use of the BEGe system for the routine measurement of NORM.
3
These areas were chosen because of their significance in the field of radiation protection
and the availability of instrumentation and devices at QUT.
1.3 Structure of the research
This thesis is presented as a combination of scientific papers, which have been
published, accepted for publication or planned for submission in international peer-reviewed
journals.
A literature review on the measurement and regulation of NORM is presented in Chapter
2. The main purposes of this chapter are to provide a fundamental understanding, examine the
current state of knowledge and identify gaps that require further study in the field of NORM.
This chapter starts with the definition of NORM from the point of view of radiological protection
and a description of its origin, distribution, activity measurement, potential health risks and
regulation. It then highlights the measurement challenges associated with quantifying the
activity concentration of NORM and their potential impact on a radiological protection regime.
The radiological protection system of NORM is critically reviewed in Chapter 3. It
discusses various aspects surrounding the current regulation and highlights key regulatory
issues in establishing and implementing a protection system against exposure to NORM. The
chapter contains a case study on the new ICRP and IAEA recommendations in the regulation of
indoor radon (222Rn) and thoron (220Rn) and the implications of the implementation of these
new recommendations. This chapter is being prepared for submission to a peer-reviewed
journal.
Chapters 4, 5 and 6 investigate the measurement and monitoring methods that are
currently in use to determine the activity concentrations of individual radionuclides of NORM in
different environments. These chapters are presented as published papers. The first paper,
Chapter 4, examines the activated charcoal canister technique for measuring 222Rn activity flux
from surfaces. This method is widely used to establish 222Rn exhalation source terms to model
radon transport and subsequent radiation dose, especially in mine sites and land contaminated
with uranium or 226R. This method basically depends on the mechanism of 222Rn adsorption into
activated charcoal, placed in a canister. The charcoal is then analysed using gamma
spectrometry. This method assumes that 222Rn flux remains constant over time, and that all of
222Rn is adsorbed and retained in the charcoal. Such assumptions may not always be warranted,
which may limit the use of this technique and affect the results. This and some other radiation
detection issues require revisiting the methodology relating to activated charcoal canisters and
reporting its limitations.
The second paper is presented in Chapter 5, and it investigates the level of indoor 222Rn
and 220Rn by measuring and monitoring their activity concentrations in workplaces in Brisbane,
4
Australia. Such measurements become important for regulatory purposes, especially after the
latest findings on the radiation hazards of indoor 222Rn exposure. The measurements cover a
wide spectrum of workplace categories with various building characteristics. The paper also
studies the variations of indoor 222Rn and 220Rn levels and identifies the main parameters that
influence their levels.
The final paper has been accepted for publication and presented in Chapter 6. It
explores the feasibility of using the broad energy germanium (BEGe) detector for determining
the activity concentrations of individual radionuclides of NORM. The use of such a detector for
routine measurements of NORM is of significant advantage for the purposes of complying with
regulations and monitoring the environment. This is because of the large crystal size of the
detector that would enhance the sensitivity for low energy gamma-rays in addition to the ability
to simultaneously measure a large number of radionuclides, which can save time and costs.
Although the detector has been widely used in nuclear physics studies, its application in
environmental radioactivity is limited. The paper establishes a regime for the routine use of the
BEGe detector for quantifying activity levels in various types of samples, including standards,
minerals, soil, tailings and organic materials. It provides details of sample preparations,
geometry selection, efficiency calibration, required corrections and counting time. The study
goes further to examine the rapid analysis of 226Ra in several materials and the determination of
227Ac by gamma spectrometry.
Chapter 7 concludes and summarises the main findings of the project and its
significance and then suggests avenues for further research areas.
5
Chapter 2 Literature Review
2.1 Introduction
In the last few decades, there has been a growing concern about exposure to NORM. This
is after the latest findings on health risks associated with the inhalation of radon and its progeny
and the detection of enhanced levels of NORM in non-nuclear industries (IAEA, 2012; ICRP,
2010; WHO, 2009). Numerous studies have been published describing various aspects related
to NORM, including measurement techniques, exposure characteristics and regulation policies.
This chapter aims to summarise the current state of knowledge relevant to the measurement
and regulation of NORM and highlights some gaps in the current knowledge, to which this
research project makes a unique contribution.
2.2 Origin and distribution
Naturally occurring radionuclides are ubiquitous in the environment and form most of
natural background radiation (UNSCEAR, 2008). They are widespread in the environment and
can be found in soil, sand, rock, minerals, air, water and biota (Al-Masri et al., 2000; Kurnaz et
al., 2007; Zare et al., 2012). Naturally occurring radionuclides can be classified into two
categories: (i) primordial radionuclides, which have existed on the Earth’s crust since it was
formed, (ii) and cosmogenic radionuclides, which have been formed by the spallation process of
the constituents of the atmosphere when they interact with the primary cosmic rays in the
upper atmosphere (Beer et al., 2012).
Primordial radionuclides 40K, 238U, 235U, 232Th and their radioactive decay products are of
the greatest importance in the radiation protection context due to their health hazards and long
half-lives, which are comparable with the age of the Earth. These radionuclides undergo
spontaneous radioactive decay, forming a lighter nuclide and emitting alpha or beta particles
accompanied by gamma radiations. Heavy radionuclides 238U, 235U and 232Th decay through a
chain of radionuclides following different decay modes before reaching a stable state (Figure
2.1). Ideally, radionuclides in each decay chain are assumed to be in a state of equilibrium, in
which the activities of all radionuclides within each decay chain are identical. However, this
equilibrium condition is subject to large disturbances in which activities of radionuclides within
the decay series are no longer equal (Ji et al., 2015; Papadopoulos et al., 2013).
The levels of naturally occurring radionuclides vary widely in the environment.
Typically, they are considered to be low, with global average concentrations of 33, 45 and 420
Bq kg-1 for 238U, 232Th and 40K , respectively (UNSCEAR, 2008). The global population weighted
6
annual effective dose rate is 2.03 mSv (UNSCEAR, 2008). However, in specific areas in the world,
such as Guarapari, Brazil; Kerala, India; Ramsar, Iran; and Yangjiang, China, these levels are
elevated naturally due to the geological and geochemical characteristics of the region (Mohanty
et al., 2004), resulting in a significant effective dose level, as much as 131 mSv per year (Sohrabi
and Esmaili, 2002). These areas are termed High Natural Background Radiation (HNBR) areas
(Hendry et al., 2009). There is another situation where the level of naturally occurring
radionuclides has been technologically altered by certain human activities (Baxter, 1996;
O'Brien and Cooper, 1998; Paschoa and Steinhäusler, 2010). This alteration may be deliberate,
as in uranium mines (USEPA, 2008), or unintentional, as in oil and gas industries (Hilal et al.,
2014).
2.3 Terminology
In the context of radiation protection, NORM is defined as material that is designated by
national law or a regulatory body as being subject to regulatory control because of its
radioactivity, containing no significant amounts of radionuclides other than radionuclides of
natural origin (IAEA, 2007b). The acronym NORM is normally used to refer to the primordial
radionuclides 40K, 238U, 235U, 232Th and their radioactive decay products. Elevated levels of NORM
due to certain human activities, which may alter the physical, chemical or radiological
properties of the radioactive material in a way that increases exposure, are also included in this
term. The term is also used by other international organisations, such as the International
Commission on Radiological Protection (ICRP, 2007a) and the European Commission (EC,
2001).
However, the use of the acronym NORM is not globally adopted to describe condition of
natural radioactive materials subject to human activities. For example, Baxter (1996)
mentioned that the use of the term NORM to describe increased levels of natural radionuclides
in the industry was misleading because it implied that the concentration was as it was in nature.
The author suggested using the abbreviation TENORM (technologically enhanced naturally
occurring radioactive material) instead. Historically, the terminology used to describe elevated
exposure due to natural sources in non-nuclear industries was first coined by Gesell and
Prichard (1975) as technologically enhanced natural radiation (TENR). In the United Sates, both
terms, NORM and TENORM, are used to distinguish between natural levels of naturally
occurring radionuclides in materials and concentrated levels, as a result of human activities
(USEPA, 2008). The Canadian authority has also used both terms, but in such a way to
differentiate between non-nuclear and nuclear industries as NORM and TENORM, respectively
(Canada, 2014). Nonetheless, the IAEA intends to continue using the term NORM in its future
publications as shown in the revision draft of its Safety Glossary 2007 Edition (IAEA, 2015b).
7
Element Uranium series Thorium series Actinium series
Uranium
238U 4.47×109
a
234U 2.45×105
a
235U 7.04×10
8 a
Protactinium
234mPa 1.17
min
231Pa 32760
a
Thorium
234Th 24.1 d
230Th 7.5 ×104
a
232Th 1.41×1010
a
228Th 1.912
a
231Th 25.52 h
227Th 18.72 d
Actinium
228Ac 6.15 h
227Ac 21.77
a
Radium
226Ra 1600 a
228Ra 5.75 a
224Ra 3.66 d
223Ra 11.44 d
Radon
222Rn 3.824 d
220Rn 55.6 s
219Rn 3.96 s
Polonium
218Po 3.05 min
214Po 164 μs
210Po 138.4d
216Po 0.145 s
64%
212Po 0.299
μs
215Po 1.781 ms
211Po 0.516 s
Bismuth
214Bi 19.9
min
210Bi 5.01 d
212Bi 60.55
min
211Bi 2.14
min
Lead
214Pb 26.8 min
210Pb 22.3 a
206Pb Stable
212Pb 10.64
h
36%
208Pb stable
211Pb 36.1 min
207Pb stable
Thallium
208Tl 3.053
min
207Tl 4.77
min
Figure 2.1 Uranium, thorium and actinium decay series
8
2.4 Behaviour of NORM
The behaviour of NORM in the environment is governed not only by the physical and
chemical properties of its individual radionuclides, but also by the type of human activities
taking place (Michalik et al., 2013; O'Brien and Cooper, 1998). NORM contains a wide range of
radionuclides with great variations in their physicochemical properties. These properties allow
certain radionuclides to be released into the surrounding environment and migrate from their
source of origin, causing disruption in the radioactive equilibrium state of the natural series.
The radionuclides released into the environment may establish their own subseries away from
the original series. In the 238U decay series, several subseries, headed by 230Th, 226Ra, 222Rn and
210Pb, have the potential to be separated. For example, 222Rn is gaseous and can easily be
released from the ground and enter buildings and establish its own subseries in indoor air
(Wang and Ward, 2002). The short-lived progeny of 222Rn may result in the presence of another
subseries headed by 210Pb in indoor environments (Chen et al., 2009; Samuelsson and
Johansson, 1994) and outdoor surfaces (Doering et al., 2006).
In industrial practices, the behaviour of NORM can be influenced by the type of industry
and operational process (Baxter, 1996; Cooper, 2005; Michalik et al., 2013). Each industry has
its own characteristics, and thus, the same radionuclide can exhibit quite different behaviour
between one industry and another. For example, 226Ra in the scale of the petroleum industry
(Abo-Elmagd et al., 2010; Jibiri and Amakom, 2011) behaves differently compared to how 226Ra
behaves in phosphogypsum in the phosphate industry (Cañete et al., 2008; Haridasan et al.,
2002). It is therefore important to characterise NORM in each industry in order to control it.
2.4.1 Uranium
Uranium (U) has an atomic number of 92 and is found naturally as three different
isotopes: 238U (99.2740%), 235U (0.7200%) and 234U (0.0055%) (Eisenbud and Gesell, 1997). The
behaviour and occurrence of uranium in the environment depends on oxidation and reduction
processes (Landa, 2007; Swarzenski et al., 2003). The most dominant U species in the
environment are tetravalent U(IV) and hexavalent U(VI) (Sheppard, 1980). The latter forms
soluble complexes with carbonate, oxalate and hydroxide ions and hence can be readily
transported through environmental media (Grenthe, Wanner and Forest 1992; Sheppard et al.
2005; Krishnaswami and Cochran 2011; Anderson, Fleisher and LeHuray 1989). On the other
hand, under reducing conditions, U becomes insoluble and occurs in the tetravalent state with
the tendency to bind with organic materials (Fredrickson et al., 2000; Gascoyne, 1992).
2.4.2 Thorium
Thorium (Th) exists naturally as six isotopes—234Th, 232Th, 231Th, 230Th, 228Th and 227Th.
Thorium isotopes are present in all the natural decay series. Th mainly occurs in the tetravalent
Th(IV) state. It is generally an insoluble element and less mobile (Chabaux et al., 2003;
9
Swarzenski et al., 2003). In aqueous media, the level of pH can play a role in Th solubility (Clark
et al., 2011; Degueldre and Kline, 2007; Langmuir and Herman, 1980; Neck et al., 2002). In near
neutral pH, Th is removed from the dissolved phase onto colloids and particulates (Landa, 2007;
Porcelli and Swarzenski, 2003). At low pH, Th undergoes hydrolysis (Katz et al., 1987) and
becomes more soluble and mobile (Altmaier et al., 2005). In soil, the mobility of Th may be less
affected by pH changes (Hansen and Huntington, 1969).
2.4.3 Radium
Radium (Ra) has four natural radioactive isotopes: 228Ra, 226Ra, 224Ra and 223Ra. Like Th,
Ra exhibits only one oxidation state (Ra2+) and is sensitive to pH changes (Almeida et al., 2004;
El Arabi et al., 2006; Martin et al., 2003; Thiry and Van Hees, 2008). Extensive studies on Ra
behaviour have shown that Ra in water is soluble as chloride, sulfate, carbonate, nitrate and
bromate salts, while it is insoluble as carbonate, iodate, oxalate and sulfate salts (Gao et al.,
2010; Grundl and Cape, 2006; Langmuir and Riese, 1985; Lin et al., 2003; Martin et al., 2003;
Martin and Akber, 1999; Sebesta et al., 1981; Zhang et al., 2014). In soil, Ra is adsorbed by soil
particles, such as clay minerals and organic matter (Landa and Reid, 1983; Shoesmith, 1984;
Totten Sr et al., 2007). However, some Ra can leach from soil or rock into the groundwater,
where it can easily migrate (Cañete et al., 2008; Chalupnik, 2005; Desideri et al., 2008).
2.4.4 Radon
Radon (Rn) is a radioactive gas that is derived from the radioactive decay of Ra in three
natural series as 222Rn, 220Rn and 219Rn. As a noble gas, Rn is chemically inactive, and it does not
precipitate in solid phase, although it can compound with fluorine under exceptional conditions
(Chernick et al., 1962). Rn is transported from subsurface to the ground mostly through
diffusion, which is proportional to the gradient of gas concentration in interstitial spaces;
additionally, advection mechanisms, induced by pressure differences created by meteorological
conditions, also aid Rn transport (Andersen, 2001; Chauhan and Kumar, 2013; Jacob and
Prather, 1990). As Rn is exhaled from the ground into the atmosphere and decays, its short lived
progeny reacts with water vapour and trace gases, forming clusters with diameters from 0.5 to
5 nm, which are then attached to aerosol particles (Michielsen and Tymen, 2007; Porstendörfer,
1994; Porstendörfer and Reineking, 1999). The concentration of Rn is by far much higher in
indoor air, where it is accumulated, than in open air, where it is diluted (UNSCEAR, 2006). In
water, Rn is soluble, and its solubility is sensitive to water temperature and salinity (Schubert et
al., 2012). It decreases with increasing temperature and salinity. The solubility coefficient of Rn
at 0°C is 0.51 and drops to 0.25 at 25°C (Ball et al., 1991). At the water-air interface, Rn tends to
readily escape to the air (Dulaiova and Burnett, 2006; Ongori et al., 2015; Schubert et al., 2012).
10
2.4.5 Polonium
Polonium (Po) is present in nature in seven isotopes: 210Po, 211P, 212Po, 214Po, 215Po, 216Po
and 218Po. Po has several oxidation states, the most predominant being the Po(+IV), which is
relatively more soluble in natural water than other states (Ansoborlo et al., 2012; Lehto and
Hou, 2011). Overall, Po solubility is only slight and its concentration in natural water is lower
than Rn and Ra. In the atmosphere, Po can be found as a result of natural processes, such as the
radioactive decay of Rn isotopes and volcanic eruptions (Gauthier et al., 2000; Su and Huh,
2002), or industrial processes, such as fossil fuel burning and metal smelting (Al-Masri et al.,
2000; Boryło et al., 2012). Po is a volatile element and has a strong tendency to condense on ash
particles (Sahu et al., 2014). Atmospheric Po is brought to the Earth’s surface through dry and
wet deposition (Persson and Holm, 2011). The residence time of Po in the atmosphere varies
from few days up to 75 days, depending on several factors, including location, altitude and
meteorological conditions (Baskaran and Shaw, 2001; Moore et al., 1973; Papastefanou, 2009;
Persson and Holm, 2011). Because of atmospheric fallout, Po is concentrated on the top layer of
the soil (Karunakara et al., 2000) and is present in plant leaves (Al-Masri et al., 2008) as well as
on water surfaces (Tateda et al., 2003). In soil, Po is likely to be adsorbed into clay and organic
colloids, which play a role in its mobility (Aslani et al., 2005; Özden et al., 2013).
2.4.6 Lead
Lead (Pb) comprises eight naturally occurring isotopes, four of which are stable (204Pb,
206Pb, 207Pb and 208Pb) and four of which are radioactive (210Pb, 211Pb, 212Pb, and 214Pb). They are
formed through the radioactive decay of radionuclides in the uranium, actinium and thorium
decay series, except 204Pb, which is non-radiogenic (Dickin, 1997). Pb has two common
oxidation states: Pb(II), which dominates in inorganic reaction, and Pb(IV), which dominates in
organic matter (Casas and Sordo, 2006). Like Po, Pb exists in the atmosphere with a residency
time of 2 to 25 days (Ahmed et al., 2004; Gäggeler et al., 1995; Mudbidre et al., 2014) and can be
carried by aerosol particles over long distances before it is removed from the atmosphere
through dry and wet deposition (Halstead et al., 2000; Yamamoto et al., 2006). Diurnal and
seasonal variations have been observed for atmospheric Pb: being higher in the night than in
the day with maximum level during fall and winter and minimum level during spring and
summer (Sheets and Lawrence, 1999). Pb is precipitated or bound to soil solids, and its mobility
is influenced by soil redox potential, available anions (e.g. carbonate, phosphate and sulfate),
pH, organic contents and presence of cations (Kabala and Singh, 2001; Reddy et al., 1995;
Roussel et al., 2000; Sauvé et al., 1998; Violante et al., 2010; Wang and Benoit, 1996). As the soil
pH increases, Pb forms complexes with organic and inorganic ligands which may enhance Pb
mobility. In aqueous solutions, the solubility of Pb decreases as the pH increases (Dando and
Glasson, 1989; Sauvé et al., 1998).
11
2.5 Exposure to NORM
NORM has the potential to pose significant radiological hazards to humans and the
environment (Vearrier et al., 2009). Exposure pathways by which NORM can cause health risks
may depend on the human activity that alters the level of NORM. Several pathways are involved
in complex scenarios, but they can be grouped into two broad categories: (i) internal exposure;
and (ii) external exposure. The determination of exposure pathways is crucial in terms of
assessing the radiation dose from NORM.
2.5.1 Internal exposure
Internal exposure occurs when radionuclide members of NORM enter the body through
inhalation and ingestion. As radionuclides decay inside the body, various types of ionising
radiations are emitted. The greatest effects are associated with alpha radiation, due to its high
linear energy transfer (LET) in tissues (Cember and Johnson, 2008). The presence of radioactive
gases and metallic radionuclides has been observed in indoor and outdoor environments as well
as in dwellings and workplaces (Arafa et al., 2002; Gründel and Porstendörfer, 2004; Murty et
al., 2010; Zeevaert et al., 2006). Airborne radionuclides can be easily transported for longer
distances since their diffusion coefficient in air is higher than in other material media
(Raghunath and Kotrappa, 1979). Radionuclides in the air readily attach to airborne particles,
such as moisture and dust, and their transport mechanism in air is mainly governed by
advection (Leelőssy et al. 2014). The most frequently detected airborne radionuclides are 210Pb,
210Po, 222Rn and 220Rn and their short-lived progeny (Baskaran, 2011; Jia and Torri, 2007; Murty
et al., 2010).
Most of the attention has been given to exposure to 222Rn in indoor environments (IAEA,
2015a; ICRP, 2014). This is because of available data relating health risks to exposure levels and
the presence of 222Rn in all buildings and closed spaces (Darby et al., 2005; UNSCEAR, 2006).
Inhalation of 222Rn and its short-lived progeny is the second main cause of lung cancer (WHO,
2009). The health risk that is commonly attributed to 222Rn is actually associated with its short-
lived progeny radionuclides 218Po, 214Pb and 214Bi that are produced in air; most of inhaled 222Rn
gas is exhaled from the lungs before it decays, due to its relatively long half-life of 3.8 days
(UNSCEAR, 2006). Based on the review of recent epidemiological studies on the association
between exposure to 222Rn and its progeny and lung cancer, the ICRP has announced that the
risk from exposure due to the inhalation of 222Rn progeny is double that of the current
estimation (ICRP, 2010). These epidemiological studies included residential and occupational
exposures and observed a significant association with the risk of lung cancer at average annual
222Rn concentration of approximately 200 Bq m-3 (ICRP, 2010).
Furthermore, some industrial processes, such as dry milling and smelting processes
generate a large amount of dust contaminated with NORM. This radioactive dust was observed
12
in a number of industries, such as the petroleum industry (Bakr, 2010; Potiriadis et al., 2011),
uranium mining and milling (Shawky et al., 2002), minerals sand extraction (Ballesteros et al.,
2008) and the phosphate industry (Samad et al., 2012). The high volatility of some
radionuclides (e.g. 210Po and 210Pb) allows their release and accumulation in industrial
processes involving heating and smelting (Jia and Torri, 2007). In a zircon refining plant, for
example, the main exposure pathway was inhalation of vaporised 210Pb and 210Po in airborne
dust in the processing plant (Cooper, 2005).
NORM can find its way into the human body through ingestion of water and foodstuff
(Abbady, 2006; Awudu et al., 2012). Doses from these pathways depend on the consumption
and the concentration of NORM in the terrestrial and aquatic environment. Some radionuclides,
such as 228Ra and 226Ra, are more soluble in an aqueous medium (Shuktomova and Rachkova,
2011; Veguería et al., 2002) and, hence, can leach from industrial waste and residue streams
and dissolve into the surrounding water. It has been detected in drinking water (Borio et al.,
2007; da Costa Lauria et al., 2012), groundwater (Almeida et al., 2004; Godoy and Godoy, 2006;
Martin and Akber, 1999), produced water during oil and gas extraction (Al-Masri, 2006;
Barescut et al., 2009; Zhang et al., 2014) and all natural water with different concentrations
(UNSCEAR, 2008). The possible exposure pathways may be direct, by drinking water and
inhaling 222Rn released from water, or indirect, by consuming vegetation irrigated with
contaminated water (Martin and McBride, 2012; Medley et al., 2013).
Natural radionuclides in soil can be transferred to plant and vegetation. They have been
detected in different food categories, including fruit, vegetable, dairy, meat and fish (Froidevaux
et al., 2006). The concentrations of natural radionuclides in foodstuff vary according to the food
category, their concentrations in the soil, their physicochemical properties and the prevailing
conditions (Alharbi and El-Taher, 2013; IAEA, 2010). Many studies have shown that
contamination of edible terrestrial organisms can be caused by the discharge of NORM from
industries (Baker and Toque, 2005; Rajaretnam and Spitz, 2000; UNSCEAR, 2008). For example,
Baker and Toque (2005) showed that radium was absorbed by plants in a contaminated site.
210Po in fly ash discharged from a coal-fired power plant was found in food crops nearby
(Zeevaert et al., 2006). Moreover, an elevated level of 226Ra has been detected in vegetation in
the vicinity of oil and gas industry sites (Rajaretnam and Spitz, 2000). However, it should be
noted that for public exposure due to natural radiation, the ingestion of radionuclides is
generally the least significant exposure pathway (UNSCEAR, 2008).
2.5.2 External exposure
External exposure refers to any exposure resulting from a source outside the body. The
dose from an external pathway is derived mainly from gamma radiation, due to its long
travelling distance and high-penetrating power (Martin et al., 2012). A large number of
13
radionuclides of natural origin are gamma emitters (Gilmore, 2008). The population-weighted
annual effective dose from external exposure due to these radionuclides in indoor and outdoor
environments is estimated to be 0.41 and 0.07 mSv respectively (UNSCEAR, 2008).
In indoor environments, external exposure to gamma radiation originates from
primordial radionuclides in the soil and building materials (IAEA, 2015a). Construction
materials such as marble and granite are known to host a high level of 226Ra, 232Th and 40K
(Beretka and Mathew, 1985; Bou-Rabee and Bem, 1996; Pavlidou et al., 2006; UNSCEAR, 2008;
Yu et al., 2000). Their concentrations depend on the industry type and the raw materials used to
manufacture the building materials. Industrial by-products, such as phosphogypsum and
residues, such as red mud generated in aluminium production, which is known to have NORM
contents, are also used to produce building materials (Cooper, 2005; Somlai et al., 2008). These
materials sometimes contain elevated levels of NORM (UNSCEAR, 2008).
In industries, external exposures to NORM are mainly experienced by workers. For a
given activity concentration of radioisotopes, the geometry of the bulk material, which is shaped
by the industry type, plays a critical role in gamma exposure. In the petroleum industry, gamma
exposure comes from gamma-emitting radionuclides 228Ra and 226Ra and their progeny
accumulated in the sludge and scale inside pipes and equipment (Abo-Elmagd et al., 2010; Hilal
et al., 2014; Jibiri and Amakom, 2011; Landsberger et al., 2013). Nevertheless, external exposure
can be controlled by removing the source or by controlling three parameters: duration of
exposure, distance from the source and use of shielding (Martin et al., 2012).
2.5.3 Natural background radiation
According to UNSCEAR (2000) Report Annex E, people working in uranium mines
receive an annual average dose of 5.0 mSv. This value is above the worldwide average annual
exposure to the natural background radiation, which is 2.4 mSv (UNSCEAR, 2008). Examining
detail about the global distribution of NORM and hence NORM-related industries, the annual
dose is quite uneven. The UNSCEAR (2008) Report discusses regional variations in NORM
radiation dose and also highlights certain areas where the dose rises above the global averages.
Annual doses of a few millisieverts is expected depending on the exposure scenario. The Report
concluded that the full picture of worldwide exposure to NORM industries is far from being
completed.
2.6 Regulation of NORM
Although the presence of natural radionuclides in industries was known from the early
20th century, when radium and radon (or radioactive gases at that time) were detected in crude
petroleum (Burton, 1904; Himstedt, 1904), the radiological protection from natural
radionuclides was a more recent concern. The radiological protection community started to
14
recognise NORM as a serious threat in non-nuclear industries in the mid-1980s, when elevated
levels of NORM were reported in the oil and gas industry (Kolb and Wojcik, 1985; Smith, 1987).
Since then, the acronym NORM began to appear in the literature as well as in radiological
protection recommendations and guidelines.
The current radiological protection system is actually developed for artificial
radionuclides. The concept of exemption, for example, is mainly designed for artificial sources of
radiation in planned exposure situations, where the source of exposure is introduced
deliberately (ICRP, 2007b). Recommendations and policies for NORM are included in general
guidance with other radionuclides with no complete guidance for controlling different
situations of NORM (IAEA, 2014; ICRP, 2007a). This has created complexities in regulating
NORM, because scenarios of exposure to NORM are often quite different from those to artificial
radionuclides (Michalik, 2009). A special case of NORM concerns radon in dwellings and
workplaces, and this is treated separately.
In general, exposure to natural sources, including NORM, is considered to be an existing
exposure situation, as the source already existed when the decision to regulate was made; with
the potential to be subject to the requirements of a planned exposure situation because human
intervention and industrial processes could alter its distribution (IAEA, 2014; ICRP, 2007a).
This can generate confusion among users of the standards (Hedemann-Jensen, 2010). Haridasan
(2015) believed that identifying a situation as either an existing or a planned exposure, and
accordingly using appropriate reference levels or dose constraints, is one of the main radiation
protection challenges. The IAEA (2014) also recognised the interference between the two
situations and stated, “Some exposures due to natural sources may have some characteristics of
both planned exposure situations and existing exposure situations”. To address this problem,
the IAEA (2014) has specified a criterion where exposure to natural sources should be treated
as a planned exposure situation:
(1) If the activity concentration of any radionuclides in the 238U and 232Th decay series
exceeds 1 Bq g-1, or 10 Bq g-1 in the case of 40K
(2) Public exposure due to discharge or waste from the situation outlined above
(3) Exposure due to 222Rn and 220Rn and their decay products in which NORM is
controlled as a planned exposure situation
(4) Exposure due to 222Rn and its decay products where its annual concentration in the
workplace remains above the reference level that is proposed by the relevant
regulatory authority, after applying reduction measures.
The values of activity concentrations, that is, 1 and 10 Bq g-1, were selected as the
optimum boundaries between the ubiquitous unmodified concentrations of the nominated
radionuclides in soil and the activity concentrations in ores, mineral sand, industrial residues
15
and waste (IAEA, 2005). No scenarios were involved in deriving the values of activity
concentrations. According to the IAEA (2004a), these values would be unlikely to deliver to
individuals doses exceeding about 1 mSv in a year. In reality, exposure to NORM involves
complex pathways (O'Brien and Cooper, 1998). An example was given by the ICRP (2007b),
which showed that activity concentrations of approximately 0.2 Bq g-1 of 238U and 232Th in
equilibrium with their progeny in waste rock piles could deliver to the public an exposure dose
of more than 1 mSv per year. This indicates that the activity concentration alone does not
deliver the same exposure dose. Moreover, it has been noted that the 1 Bq g-1 criterion has been
misinterpreted as a limit (as concluded by the Seventh International Symposium on Naturally
Occurring Radioactive Material, China, 2013), although the 1 Bq g-1 criterion is widely accepted
(Haridasan et al., 2015).
There is a wide variation across the globe in defining the scope of the NORM regulation
and implementing the regulation. National and international standards on regulating NORM are
sometimes ambiguous (ICRP, 2007b). There is also a notable discrepancy in managing NORM
industries, such as in the case of radium that is well-regulated when it is encapsulated and left
uncontrolled, as when it is found in residues (Michalik et al., 2013). Such situations can lead to
regulatory challenges in controlling exposures to NORM (Chen, 2015; Haridasan, 2015).
Michalik et al. (2011) argues that one of the most generic problems with NORM is the
identification of circumstances where enhanced levels can be found. The IAEA acknowledged
this fact by publishing a safety report to identify industrial sectors and processes, other than
uranium mines, involving NORM that were likely to require regulatory considerations (IAEA,
2006). Eleven industries have already been identified to have enhanced levels of NORM due to
their processes, including extraction of rare Earth elements; production and use of thorium and
its compounds; production of niobium and ferro-niobium; mining of ores other than uranium
ore; production of oil and gas; manufacturing of titanium dioxide pigments; operation in the
phosphate industry; operation in the zircon and zirconia industries; production of tin, copper,
aluminium, zinc, lead, iron and steel; combustion of coal; and water treatment. There are other
industries that have been reported by researchers where NORM could be present and yet to be
identified by regulatory bodies, including fish hatchery (Kitto et al., 1998), geothermal power
installation (Eggeling et al., 2013) and recycling of metal scraps (Ortiz and Carboneras, 2011).
To date, NORM is only managed in five non-industries, namely, the oil and gas industry,
phosphate industry, the zircon and zirconia industries, the production of rare earths from
thorium-containing minerals, and the titanium dioxide and related industries. Guidelines for
managing NORM in these industries are available (IAEA, 2003c, 2007a, 2011, 2012, 2013d).
General guidance, such as Management of NORM Residues (IAEA, 2013a), are applied to the rest
of the industries. However, the use of a general approach may be inappropriate because the
16
exposure scenario may vary on a case-by-case basis, depending upon the type of industry and
the industrial process (ICRP, 2007b).
The regulation of radon in dwellings and workplaces appears to have more attention
than other radionuclide members of NORM. This is because of its presence at many places of
human occupancy and confirmed radiological hazards (Darby et al., 2005; WHO, 2009).
Guidelines and policies on the management and control of radon exposure have already been
published and applied (IAEA, 2014, 2015a; ICRP, 2014). Radon is normally managed as an
existing exposure situation where a reference level has to be set (IAEA, 2014). Regulatory
authorities are encouraged to establish an action plan for controlling exposure to indoor radon,
which requires conducting a measurement to identify radon-prone areas and set appropriate
reference levels (IAEA, 2015a; WHO, 2009). Reference levels of less than 300 Bq m-3 was
mentioned internationally for indoor radon as in the recent published recommendations (EU,
2013; IAEA, 2014; ICRP, 2014). In two cases, radon exposure is managed as planned exposure
situations: (i) in practices where NORM is managed as a planned exposure situation; (ii) in
workplaces where the concentration of radon exceeds the specified reference level (IAEA,
2014).
There is emphasis on regulatory bodies to consider methodologies for the sampling,
averaging, monitoring and detection of radionuclides when deriving activity concentrations for
regulatory purposes (IAEA, 2005). Errors in measurement can result in errors when including
or excluding materials from a regulation. For example, Lavi et.al. (2004) found that, in gamma
spectrometry, ignoring the interference that occurred between gamma lines of 228Ac and 40K at
1460 keV could lead to errors up to 150% in 40K result. In oil and gas industries, the
measurement of radioactivity in scale inside pipes, pumps, valves and other equipment was
affected by pipe geometry, which represents one of the major issues in managing NORM in
waste stream (ARPANSA, 2005). Lack of an adequate infrastructure for analysing and
interpreting NORM was found to be a problem in managing NORM (Haridasan, 2015).
Therefore, a number of reports have been provided by regulatory bodies to standardise
analytical procedures for the determination of natural radionuclides and to ensure that specific
criteria of accuracy are met (USEPA, 1993; IAEA, 2013c; ISO, 2012).
2.7 Measurement of NORM
A wide range of methods and techniques are available for measuring the activity
concentration of various radioisotopes in materials (Knoll, 2010; Turner, 2007). The selection of
an analytical technique and measurement method is primarily dictated by the purpose for
which the measurement is being conducted (Macášek, 2000). Radionuclides of natural origin
are ubiquitous and present in almost all environments, including soil, sediments, waste,
17
residues, by-products, building materials and rocks (UNSCEAR, 2008). These materials are also
the prime source of radon gas, which is released from them and accumulated in enclosed spaces
(Pasculli et al., 2014; UNSCEAR, 2006). Therefore, measurements of natural radionuclide
members of NORM in solid media as well as measurements of radon released from the ground,
and radon trapped in indoor air are of utmost importance in the context of radiological
protection. Further, the implementation of radiological protection standards and the
demonstration of regulatory compliance rely on such measurements.
Beside radiological assessment, the measurement of individual radionuclides in the
uranium and thorium series has many applications in various disciplines. For example, 210Pb has
been extensively used as an environmental radiotracer in sediment dating (Abril and Brunskill,
2014; Putyrskaya et al., 2015; Sanchez-Cabeza and Ruiz-Fernández, 2012), atmospheric studies
(Brattich et al., 2015; Heijnis et al., 1993; Lamborg et al., 2000), ocean science (Stewart et al.,
2007; Verdeny et al., 2009) and pollution investigation (Cheng and Hu, 2010; Sakan et al., 2015).
2.7.1 Measurement of radon releases
The release of radon from materials to the atmosphere takes place when radon is
emanated from grains into the interstitial space and then transported to the surface, where it
can be released (Nazaroff, 1992). The amount of radon release over a specific area from a
material surface is referred to as exhalation flux density and is given as the activity per unit of
area and time (Bq m-2 s-1). This exhalation is commonly measured by placing a chamber
(accumulator) directly into the sampling area in an upturned position (Ferry et al., 2002; Ielsch
et al., 2002; Schery et al., 1989), or by enclosing the sample materials in an airtight container
(Jonassen, 1983; Samuelsson, 1990; Tufail et al., 2000) using either passive detectors, which do
not require electrical power to operate but require laboratory analysis, or active detectors that
require electrical power to operate (ISO, 2012). The measurement can be distinguished on the
basis of the sampling duration into instantaneous, continuous and integrated (over a short or
long period of time) measurement methods.
Over the past few decades, a number of passive and active techniques have been
developed to measure radon release from surfaces (Cohen and Cohen, 1983; Iimoto et al., 2008;
Kotrappa et al., 1990; Spehr and Johnston, 1983; Wilkening and Hand, 1960). Passive detectors
are usually cheaper and more suitable for integrating measurement, and a large number of
detectors can be deployed over a wide area for a reasonable period of time. On the other hand,
active detectors are expensive and require periodic maintenance, which renders them less
desirable for field measurement.
One of the best known passive techniques is the nuclear track detector (Fleischer et al.,
1965). It has been used for measuring radon flux density in building materials (Al-Jarallah and
Abu-Jarad, 2001; El-Bahi, 2004; Najam et al., 2013) as well as in soil (Abd-Elzaher, 2012; Oufni,
18
2003; Saad et al., 2013) and fly ash samples (Kumar et al., 2005; Mahur et al., 2008). However,
the sensitivity of nuclear track detectors is relatively low, which makes them more suited for
long-term radon concentration averages than for short-term exhalation flux density (IAEA,
2013b). Electret detectors have also been used to measure radon flux density (Kotrappa and
Stieff, 2009; Kovler et al., 2004; Righi and Bruzzi, 2006). Nonetheless, the performance of
electret detectors is affected by both background gamma radiation (Kotrappa et al., 1990;
Usman et al., 1999) and extreme humidity conditions (Mahat et al., 2001).
Activated charcoal canister is another passive technique that has been used extensively
to measure radon exhalation flux density (Bollhöfer and Doering, 2015; Cosma et al., 1996; De
Jong et al., 2005; Doering et al., 2006; Lawrence et al., 2009; Nisti et al., 2014; Petropoulos et al.,
2001; Spehr and Johnston, 1983; Wang et al., 2009). In industrial applications, activated
charcoal canisters were used to estimate the activity from residues and waste sites (Bollhöfer
and Doering, 2015; Dueñas et al., 2007; Nisti et al., 2014). In an unpublished work by Al-Farsi
(2008), radon flux density from petroleum industry waste samples, including sludge and scale,
were measured using activated charcoal detectors as well as an emanometer detector (an active
radon technique). A good correlation was found between the two methods. The activated
charcoal detector was also used for determining radon exhalation from building materials (De
Jong et al., 2005; Petropoulos et al., 2001).
The mechanism of the activated charcoal technique is based on the adsorption of radon
gas on charcoal. When measuring the exhalation flux density, canisters are deployed on the
target area where they are exposed for a few days, typically between one to five days, and then,
the adsorbed radon in the activated charcoal is determined through gamma radiations emitted
from its daughters using gamma spectrometry (Bollhöfer et al., 2006; Countess, 1976; Dueñas et
al., 2007; Lawrence et al., 2009). The technique has many advantages, including low cost,
reusability and simplicity. It is also recommended by some regulatory bodies for regulatory
purposes (ARPANSA, 2008). However, changes in environmental conditions, such as humidity,
temperature and pressure, have the potential to affect the sensitivity of activated charcoal (El
Samman and Arafa, 2002; Iimoto et al., 2008; Ronca-Battista and Gray, 1988; Vargas and Ortega,
2007). Moreover, this technique is based on the assumptions of constant radon flux density over
exposure time and retaining all the radon that entered the canister (Spehr and Johnston, 1983).
Although these assumptions may not always be valid in the real world, it has been widely
accepted for this method.
Radon flux density can also be measured with an emanometer (Whittlestone et al., 1998;
Zahorowski and Whittlestone, 1996). This is an active detector that is based on a flow-through
or accumulator method. It consists of two scintillation cells coupled with two separated
photomultiplier tubes (PMT) connected to a data acquisition and control unit. The two
19
scintillation cells are separated with a delay line enough for the decay of thoron to allow
thoron/radon discrimination (Zahorowski and Whittlestone, 1996). Lawrence et al. (2009) used
an online emanometer and charcoal canisters to independently measure radon exhalation rates
at different sites in tropical northern regions of Australia. Their measurements covered both the
dry and wet seasons, with their large variations in soil and air moisture. The authors reported
obtaining statistically overlapping results with both methods. The emanometer is limited by the
diffusion losses including the build-up of radon/thoron concentrations in the chamber, back
diffusion and leakages at edges of the chamber, which can be significant (Zahorowski and
Whittlestone, 1996).
2.7.2 Measurement of atmospheric radon concentration
Radon presents in both outdoor and indoor environments. In the outdoors, however, it
is readily diluted to low concentration levels, typically around 10 Bq m-3 with a wide range from
1 to 100 Bq m-3 (UNSCEAR, 2006). In a nationwide survey conducted in Japan, the mean outdoor
radon concentration was found to be 6.1 Bq m-3, which corresponded to 40% of the indoor value
(Oikawa et al., 2003). Popic et al. (2011) observed seasonal variations in outdoor radon
concentrations in decommissioned Norwegian iron and niobium mining sites that were rich in
naturally-occurring radionuclides, being higher in the summer, with an arithmetic mean in the
range of 29–82 Bq m-3, than in fall. This result is in contrast to that obtained by Chan et al.
(2010), who found that outdoor radon concentration was lower in the summer season. This
variation in outdoor radon levels was influenced by metrological conditions, such as rainfall,
wind speed, temperature and humidity, which play an essential role in controlling radon
releases from the ground (Akber et al., 1994; Lawrence et al., 2009; Martin et al., 2004).
Nonetheless, time spent outdoors is generally much less than that spent indoors, which makes
outdoor radon less significant in terms of radiological protection.
On the contrary, radon tends to build up in enclosed spaces and when there is less air
volume exchange, such as buildings, tunnels, caves and underground mines. Epidemiological
studies of residential and occupational exposure have shown evidence of a relationship between
indoor radon exposure and lung cancer (Darby et al., 2005; Krewski et al., 2006; Tomasek et al.,
2008; WHO, 2009). These findings have increased awareness of indoor radon exposure,
whether in dwellings or in workplaces (IAEA, 2015a; ICRP, 2014). Controlling such exposure
can significantly reduce its hazards, and the first step to do so is to determine the activity
concentration of indoor radon levels and the parameters that may influence that concentration
(WHO, 2009).
Most indoor radon is driven from the ground through cracks in the floor (UNSCEAR,
2006). The level of radon in the ground is largely influenced by the underlying geology (Moreno
et al., 2008; Sundal et al., 2004). The authors found a significant correlation between the level of
20
indoor radon and the geological composition of the ground (Moreno et al., 2008; Sundal et al.,
2004). Building materials and water supply can also contribute to indoor radon (Cosma et al.,
2013; Najam et al., 2013; Song et al., 2005). Bavarnegin et al. (2013) found that radium content
in building materials used in Ramsar, Iran, a region with high natural background radiation, was
the main source of indoor radon. The contribution of radon in the water supply to indoor air
was much less compared to other sources, and most of the risks arise from the ingestion of
water containing radon (WHO, 2011).
Seasonal and diurnal variations in indoor radon levels have been observed in dwellings
(Bochicchio et al., 2005; Faheem and Mati, 2007; Prasad et al., 2012; Singh et al., 2005). The
highest level was usually obtained during winter, while the lowest was in summer. These
variations were mostly attributed to ventilation systems, such as air-conditioning and heating
systems (Lee and Yu, 2000; Marley and Phillips, 2001). In regions with colder climates, the
heated air was intentionally retained indoors, where radon builds up. If measurements were
made for a period shorter than a seasonal cycle, then some correction factors would need to be
applied to get an annual average (Baysson et al., 2003; Faheem and Mati, 2007; Pinel et al.,
1995). Alternatively, measurement may be conducted during the coldest months (IAEA, 2003b;
NORDIC, 2000). Seasonal variations of indoor radon were observed less in workplaces, and
some have exhibited no seasonal patterns (Inoue et al., 2013; Papachristodoulou et al., 2010;
Tokonami et al., 1996). Instead, diurnal variations were reported more often in workplaces
(Vaupotič et al., 2012; Zahorowski et al., 1998).
Less attention has been given to indoor thoron (220Rn), and the available data are largely
limited (Burghele and Cosma, 2012; Inoue et al., 2013; Vaupotič et al., 2012). This is perhaps
due to measurement difficulties associated with the short half-life of thoron (56 s). It should be
noted that health risks associated with the inhalation of thoron and its decay products are
comparable with those of radon and its progeny (ICRP, 1993; Shang et al., 2005; UNSCEAR,
2000). Moreover, the concentration of thoron in some indoor spaces has been found to be four
times greater than that of radon (Ramola et al., 2012; Yamada et al., 2006). Additionally, the
exhalation flux density of thoron is generally greater than that of radon (Sharma and Virk,
2001). In the indoor environment, building materials represent the main source of thoron gas
(McLaughlin, 2010; Meisenberg and Tschiersch, 2011), and its level is more influenced by the
distance from the source, which includes the walls and floors in buildings (Fujimoto et al., 1994;
Urosevic et al., 2008).
2.7.3 Measurement of NORM in solid media
The radioactivity in a material such as soil, sediment, rock, products, by-products, waste
and residues is often quantified through ionising radiation emitted during the radioactive decay
of the radionuclide within that material. Alpha, beta and gamma spectrometry are most
21
frequently used for such purposes (Bonotto et al., 2009; El-Daoushy and Hernández, 2002;
Martin and Hancock, 1992; Murray et al., 1987; Vesterbacka et al., 2009). Mass spectrometry
has also been employed to measure mass concentration, hence the activity concentration of
long-lived radioisotopes (Roos, 2014). These measurement methods, with the exception of
gamma spectrometry, involve chemical separation, which is destructive and inconvenient
(Bickel et al., 2000). Therefore, gamma-ray spectrometry can be the best choice for quantifying
the activity concentration of natural radionuclides being able to simultaneously detect a large
number of natural radionuclides in a non-destructive manner.
Germanium (Ge) detectors are by far the most common gamma detectors for activity
concentration measurement. They have higher gamma energy resolution, which allows
distinguishing gamma emission for different radioisotopes. The higher atomic number of
germanium compared to the previously-used silicon (Si) detectors also makes it applicable for a
wide range of energy (Gilmore, 2008). Over the last few decades, several types of high-purity
germanium (HPGe) detectors have been developed. These detectors have different crystal
configurations to fit particular applications. The coaxial HPGe detector has been widely utilised
in numerous environmental studies (Alghamdi and Aleissa, 2015; De Corte et al., 2005; El-
Daoushy and Hernández, 2002; Kaste et al., 2006; Lenka et al., 2009; Mauring et al., 2014;
Putyrskaya et al., 2015; Scholten et al., 2013; Siegel, 2013; Van Beek et al., 2010; Yücel et al.,
2010). Lenka et al. (2009) used a coaxial HPGe detector to measure 238U in soil through the 63
keV gamma peak of its daughter radionuclide 234Th. Al-Masri and Suman (2003) measured
radium isotopes from the NORM waste site of an oil and gas industry using a coaxial detector.
Nonetheless, the coaxial detector has relatively lower efficiency at low gamma energy compared
to the well-type and planar detectors (Han and Choi, 2010).
The well-type detector has also been used to determine the activity concentration of
natural radionuclides (Gourdin et al., 2014; Legeleux and Reyss, 1996; Reyss et al., 1995; Van
Beek et al., 2008). Despite the high efficiency of this detector, it suffers from a high coincidence
summing effect due to the 4π solid angle (Sima and Arnold, 1996). This effect can limit the use
of the well-type detector for measuring radionuclides of natural origin because of their complex
decay schemes and complicated spectral interference (Gilmore, 2008). Another type of HPGe
detector is the planar detector with a small and thin crystal shape. Planar detectors showed a
higher sensitivity at low gamma energy (Loaiza et al., 2011) and have the advantage of
successfully detecting natural radionuclides with gamma emissions in that region, such as 210Pb
and 234Th (Al-Masri et al., 2010; Di Gregorio et al., 2007). However, at a higher energy, the
efficiency of conventional planar detectors is lower than that of well and coaxial detectors (Han
and Choi, 2010). One of the HPGe detectors (manufactured by Canberra Industries Inc.) has
developed a planar detector whose sensitivity at a higher gamma energy has been promoted to
22
be comparable with that of the coaxial detector (Canberra, 2013). The new detector is known as
broad energy germanium (BEGe) detector. According to the manufacturer, the BEGe detector
can effectively cover an energy range from 3keV to 3MeV with suitable resolution and efficiency
for NORM radioactivity measurements. It is, therefore, well suited for the application of NORM.
The BEGe detector has been used extensively in nuclear physics (Aalseth et al., 2011;
Agostini et al., 2011; Budjáš et al., 2013; Di Vacri et al., 2009) and medical science (Fantínová
and Fojtík, 2014; Kramer et al., 2009; Liye et al., 2007; Webb and Kramer, 2001). However, the
use of BEGe detectors in environmental studies involving radionuclides of natural origin is still
limited (Condomines et al., 2010; Putyrskaya et al., 2015). Zhang et al. (2009) used the BEGe
detector to accurately determine the activity concentration of 226Ra in soil. The authors
deconvoluted the interference between 226Ra and 235U at 186 keV peak using Aatami software.
The employment of the BEGe detector for routine analysis of natural radionuclides can bring
benefits to the field of NORM, including the detection of low gamma energy emitted from
radionuclides such as 210Pb and 234Th and the determination of radionuclides with low gamma
intensity, such as 230Th. However, like other Ge detectors, there are a number of issues that
need be addressed regarding the use of BEGe detectors in such a field, including efficiency
calibration, geometric effects, coincidence summing and self-absorption effects.
2.8 Measurement challenges
Since the discovery of radioactivity until the present day, there has been significant
development in techniques and instrumentations to quantify the radioactivity of materials. In
the case of NORM, the measurements involve sampling and analysis of a broad spectrum of
radionuclides, which vary in their physicochemical properties in various types of material. In
such cases, difficulties are expected to arise and pose challenges to the field of NORM. These
challenges include, but are not limited to, the cost of analysis, uncertainty of measurement, the
background and the state of equilibrium.
2.8.1 Measurement cost
The cost of the measurement is one of the most important factors in choosing the
appropriate analytical method, in addition to the desired application and detection limit (Tyler,
2014). The appropriate analytical procedure for measuring natural radionuclides of interest
needs to be selected carefully to maximise the benefits and minimise the cost. This cost includes
financial cost, which is required for instrumentation, resource, operation and maintenance and
time required to complete the analysis. NORM radioisotopes' analytical methods generally
involve time and resources and, therefore, are more expensive (Aycik, 2009; L'Annunziata,
2012). NORM has many radionuclides that vary in their chemical and physical properties and
have a ubiquitous distribution within industrial sites, operations, processes, products, by-
23
products, residues and waste (Al-Saleh and Al-Harshan, 2008; Hilal et al., 2014; Kumar et al.,
2005; Palomo et al., 2010; Pontedeiro et al., 2007; Tufail et al., 2006). In each case, specific
radionuclides dominate. The measurement of individual radionuclides is likely to involve
different sampling and analysis techniques due to the differences in the chemical and physical
properties of these materials, which would raise the financial costs of the analysis.
A number of studies have attempted to reduce the cost by reducing the time required for
the analysis (Johnston and Martin, 1997; Palomo et al., 2011; Sengupta et al., 2015; Zapata-
Garcia et al., 2009). Dowdall et al. (2004) measured 226Ra directly from the peak 186 keV by
subtracting the contribution of 235U at the same peak using the activity ratio 235U/238U instead of
waiting for three weeks to achieve secular equilibrium between 226Ra and its daughters 214Pb
and 214Bi, which are usually used to determine the activity concentration of 226Ra. Lin et al.
(2015) developed a method for the rapid and simultaneous detection of alpha and beta
radioactivity in food using liquid scintillation counting. Their method resulted in a reduction of
sample preparation time for a batch of eight samples from three days in traditional methods to
around five hours.
Another way to reduce the cost of the analysis is by using an in situ measurement
technique system (Al-Masri and Doubal, 2013; Aldenkamp et al., 1992; Clouvas et al., 2005;
Tyler, 2014; Wilson et al., 1999). This technique eliminates the need to collect and send samples
to a laboratory for analysis. Povinec et al. (2014) found in situ underwater gamma-ray
spectrometry to be a cost-effective technique for many applications in marine radioactivity. A
comparison between in situ and laboratory analysis of 40K, 208Tl, and 214Bi in seabed samples was
conducted by Androulakaki et al. (2015). They observed a satisfactory agreement between the
two methods with the advantage of in situ measurement in saving time required for laboratory
analysis.
Monte Carlo (MC) simulation is also employed to minimise the cost of NORM
measurement. It has been used to calibrate gamma detectors instead of purchasing expensive
calibration standard sources (Abbas et al., 2002; Britton et al., 2013a; Bronson, 2003; Dejeant et
al., 2014; Guerra et al., 2015; Habib et al., 2014). A sourceless efficiency calibration software
using MC simulation was tested by Bronson (2003) for 13 detectors. The author concluded that
using such a method was more convenient and quicker than the source-based calibration
method. Moreover, the MC simulation was used to determine correction factors required for
some measurements, such as coincidence summing and self-absorption effects (Agarwal et al.,
2011; Britton et al., 2014; Garcıa-Talavera et al., 2001; Huy et al., 2013; Quintana and Montes,
2013). However, the summing effect may be more complicated in samples containing
radionuclides from the uranium and thorium series due to the presence of a large number of
radionuclides that emit several gamma radiations. Additionally, simulation is limited by the lack
24
of precise data from sample geometry and composition, as well as from detector characteristics,
which can be an additional source of error (Britton et al., 2013a). Nonetheless, developing a
cost-effective technique for measuring NORM will remain one of the major challenges of NORM
measurement.
2.8.2 Uncertainties
Radioactive decay is a random process and, therefore, uncertainty is intrinsic in the
analysis of radioactivity. Additionally, systematic errors can also creep into the measurement.
Uncertainties in radioactivity measurement give some indication of the quality of the results
(EURACHEM., 2012). At the same time, they can affect the precision and accuracy of the
measurement of the activity concentration (IAEA, 2003a). Due to its importance in analytical
measurement, a number of guidelines have been published to provide tools for evaluating and
quantifying uncertainty across a wide range of measurement (e.g. EURACHEM., 2012; IAEA,
2004b; ISO, 1995).
The uncertainty associated with the measurement of NORM, in general, can arise from
many components that could be related to the sample itself, including matrix, weighting, water
content, radioactivity and homogeneity; instrumentations, including spectral acquisition,
efficiency calibration and background; and nuclear properties, including radioactive decay,
probability of emission and half-life (Heydorn, 2004; IAEA, 2004b; Xhixha et al., 2013). The
determination of the main source of uncertainty components associated with a particular
measurement is specified by the analytical technique (IAEA, 2004b). For example, in the in situ
gamma measurement of NORM, soil humidity was found to be the main contributor to the total
uncertainty, while in laboratory gamma spectrometry, counting statistics dominated (Al-Masri
and Doubal, 2013). Shakhashiro and Mabit (2009) reported heterogeneity and moisture content
in samples as the largest contributors to the overall uncertainty in their gamma analysis. In
another measurement method, that is, a delayed coincidence counter for the determination of
radium isotopes in water, the associated uncertainties were influenced by the sample volume,
counting time, and elapsed time between sampling and measurement (Garcia-Solsona et al.,
2008).
In order to achieve better statistics when using a detector, such as gamma spectrometry,
long counting time, large sample size and close sample-detector distance are recommended
(Garcia-Solsona et al., 2008; Gehrke and Davidson, 2005; Herranz et al., 2008). However, the
larger sample volume will result in a self-absorption effect (Huy et al., 2013), and the closer
distance to the detector may increase the probability of cascade summing (Garcıa-Talavera et
al., 2001). In addition, the longer counting time will cause the short-lived radionuclides to decay
during the measurement (Abo-Elmagd et al., 2010). The corrections required for these factors
25
are also associated with uncertainties that must be added to the propagated uncertainty of the
measurement.
It should be noted that these uncertainties are associated with the measurement of
radioactivity only. They are essential for any further use of the resulting values. When it comes
to dose assessment, for example, the estimated uncertainties of the radioactivity measurement
must be considered with other sources of uncertainties, such as uncertainties of radionuclide
discharge and transfer to the environment, as well as dose coefficient of intake (Betti et al.,
2004).
2.8.3 The background
Primordial radionuclides from 238U, 235U and 232Th decay series as well as 40K are
ubiquitous and can be found in constructional materials, ambient environments, shielding
materials and even the equipment itself. Hence, a major aspect of the measurement of NORM is
that the radionuclides being measured in the sample are present in the background. They form,
along with cosmic rays, a significant component of the background of the measurement
detectors (Knoll, 2010). Unfortunately, detectors cannot distinguish between radiation
originating from the sample and that from the background. Therefore, the background should
be determined and subtracted from the spectrum (Gilmore, 2008).
Over the last few decades, considerable efforts have been devoted to the reduction of the
background counts of detectors (Alghamdi and Aleissa, 2015; Neder et al., 2000; Niese, 2014;
Povinec, 2008; Semkow et al., 2002). In gamma spectroscopy, lead shielding is usually used to
protect the detector from background radiations. Smith et al. (2008) examined three types of
lead shielding assemblies and found differences in the background radiation in the shielding
material itself. Obviously, low background shielding should be preferred. On the other hand, the
interactions of background radiations with shielding materials can produce a low-energy
bremsstrahlung continuum and induce characteristic X-rays (Alessandrello et al., 1998; Britton
et al., 2013b; Mrđa et al., 2007). Graded shielding, where the internal surface is coated with
elements such as tin, aluminium and copper, was developed for gamma detectors in order to
absorb the lead X-ray and reduce general background radiation (Kramer and Hauck, 2005;
Singh et al., 2011). Moreover, a Compton suppression system was also introduced to reduce
Compton continuum denominated in below 1 MeV of gamma spectrum (Nolan et al., 1994;
Schumaker and Svensson, 2007). In the Compton suppression system, the HPGe detector is
surrounded by an assembly of guard detectors, which will collect the escaping photons from the
HPGe detector.
A further development in reducing background radiation was the introduction of
underground laboratories (Busto et al. 1992; Helbig and Niese 1986; Kaye et al. 1972; Hult et al.
2006; Laubenstein et al. 2004; Povinec, Commanducci and Levy-Palomo 2005; Zeng et al. 2014).
26
Underground laboratories are more efficient in reducing cosmic rays. Nevertheless, such
laboratories are not of great benefit to some detectors, such as gamma detectors because
primordial radionuclides still present in construction materials, and liquid scintillation
detectors because 50% of the background is caused by detectors themselves (Niese, 2014).
Although these solutions are effective in minimising the background in the measurement of
NORM, they may lead to the first challenge, which is cost.
2.8.4 The equilibrium state
In the measurement of radionuclides in the natural decay series, the activity
concentration of a parent radionuclide is often estimated from its daughters and vice versa (El-
Daoushy and Hernández, 2002; Huy and Luyen, 2004; Scholten et al., 2013; Siegel, 2013). Such
measurement is based on the assumption of equilibrium conditions, where the activity
concentrations of the parent and daughter radionuclides are equal in a secular equilibrium or
can be calculated in a transient equilibrium (Debertin and Helmer, 1988). Understanding the
state of equilibrium is vital in terms of dose assessment due to NORM (Lombardo and Mucha,
2008).
Equilibrium conditions are utilised when the radionuclide cannot be directly
determined by the detection method or to overcome difficulties in the measurement. In gamma
spectrometry, for example, the activity concentration of 226Ra is commonly determined through
gamma-rays emitted from its decay products 214Pb and 214Bi after achieving secular equilibrium
(Al-Masri and Aba, 2005; Landsberger et al., 2013; Sartandel et al., 2014). This is more accurate
than using the deconvolution (Saïdou et al., 2008; Zhang et al., 2009) or 238U/235U activity ratio
(Dowdall et al., 2004) when measuring 226Ra directly from its gamma peak at 186.1 keV that
interferes with the 185.7 keV peak of 235U. More examples of radionuclide measurement using
equilibrium are given in Table 2.1.
Table 2.1 Examples of radionuclides in equilibrium
Radionuclide of interest
Measured radionuclide
Typical delay time
Reference
238U 234Th and 234mPa
4 months (Huy and Luyen, 2004; Lenka et al., 2009; Yücel et al., 2009)
226Ra 214Pb and 214Bi 3 weeks (Dowdall et al., 2004; Landsberger et al., 2013; Murray et al., 1987)
228Ra 228Ac 36 hours (Lourtau et al., 2014; Xhixha et al., 2013) 228Th 224Ra
212Pb and 208Tl 3 weeks 2 days
(Awudu et al., 2012; Condomines et al., 2010)
227Ac 227Th and 223Ra 3 months (Köhler et al., 2000; Van Beek et al., 2010)
223Ra 219Rn A minute (Desideri et al., 2008; El Afifi et al., 2006)
27
One of the drawbacks of using equilibrium conditions in the measurement of
radioactivity is the waiting time required to achieve equilibrium, especially for relatively long-
lived daughters. In a well-known example, the measurement of 226Ra through its daughters
required at least three weeks of delay period to attain secular equilibrium (Siegel, 2013). In the
case of the measurement of 238U through 234Th and 234mPa, the delay time is up to four months
(Kaste et al., 2006). Several studies attempted to overcome this issue by developing a shortcut
method (De Corte et al., 2005; Johnston and Martin, 1997; Li et al., 2015; Martin et al., 1995).
For example, Li et al. (2015) studied the possibility of measuring 226Ra through gamma
spectrometry using the 214Pb-ingrowth method without waiting for equilibrium. They reported
good agreement with the results obtained through the traditional method, that is, after
equilibrium.
However, the equilibrium state can be disrupted under certain conditions. This occurs
when members of the radioactive decay series are removed or added to the system, creating a
condition of disequilibrium. This disequilibrium can occur naturally or technologically due to
some human activities, such as in uranium mines, where uranium isotopes are separated from
the rest of the primordial radionuclides in the uranium decay chain (Bollhöfer et al., 2006;
Dejeant et al., 2014). In the environment, disequilibrium is controlled by the behaviour of
individual radionuclides; physicochemical properties play a role in causing leachability and
mobility of radionuclides (Cañete et al., 2008; Rajaretnam and Spitz, 2000; Wang et al., 2012).
This disequilibrium has been largely observed in water (Labidi et al., 2010), sediment (Stewart
et al., 2007) and NORM industrial sites (Abo-Elmagd et al., 2010). Once the equilibrium is
disturbed, it requires days, weeks, months, thousands or even millions of years to be restored,
depending on the half-lives of the radionuclides.
Disequilibrium can be a serious source of error in the measurement of radionuclides in
the decay series (IAEA, 2003a; Mattinson, 2010). Assuming equilibrium, which is not the case in
reality, can lead to an erroneous estimate of the activity concentration of radionuclides of
interest. Charette et al. (2012) attributed the poor agreement in 226Ra gamma measurement
among the participating laboratories of an inter-laboratory comparison experiment on radium
isotopes to insufficient equilibration caused by the escape of radon gas from samples. They
recommended special care to be exercised to ensure the retention of radon gas and to allow
adequate time to establish equilibrium. Charette et al. (2013) studied the influence of a sealing
technique on radon loss from sample containers when measuring 226Ra with gamma
spectrometry. Their results indicated < 6% loss of radon from the sealed container containing
reference materials and found the use of epoxy resin as a sealing material to be the best way to
prevent radon loss.
28
In the field, disequilibrium, especially in the uranium series, has been reported in many
sites, including industrial wastes and residues (Condomines and Rihs, 2006; De Corte et al.,
2005; Gascoyne, 1992; Patra et al., 2013). 226Ra is not expected to be in equilibrium with its
decay products, due to the possibility of radon releases from the surface layer of the soil (Miller
et al., 1994), or due to the leachability of 226Ra from the site (Rajaretnam and Spitz, 2000).
Therefore, the use of in situ measurement at such sites with the assumption of equilibrium can
be misleading. Miller et al. (1994) used 235U peaks at 144 and 163 keV to determine 226Ra in soil
using in situ gamma measurement to avoid the disequilibrium effect due to radon exhalation.
Nuccetelli and Bolzan (2001) attributed the discrepancy between 226Ra activity concentrations
in building materials obtained directly via 186 keV and indirectly from 214Pb using in situ
technique to radon release from building materials.
2.9 Conclusion
This chapter has presented a review of the current literature available on the
measurement, monitoring and regulation of NORM. The regulation of NORM has been given
more attention due to the latest epidemiological findings on the association of lung cancer with
the inhalation of radon and its progeny along with the existence of enhanced levels of NORM in
the oil and gas industry. New research findings in the literature reported the identification of
radon-prone areas and enhanced NORM levels in many non-nuclear industries, including
extraction of rare Earth elements; production of thorium; production of niobium; mining of ores
other than uranium ore; manufacture of titanium dioxide pigments; the phosphate industry; the
zircon and zirconia industries; production of tin, copper, aluminium, zinc, lead, iron and steel;
combustion of coal and water treatment. Each industry has its own exposure characteristics and
pathways. The standards and policies currently employed to regulate NORM are originally
designed for artificial radionuclide. This adoption has caused confusion and has created new
challenges in the radiation protection system across the globe. Some regulatory bodies have
published recommendations and standards, which cover limited aspects of NORM. Current
monitoring and regulation regimes need to be reviewed in light of the most recent knowledge.
Potential regulatory issues that may impact on the development, implementation and
compliance with NORM legislation should be identified.
The core of any NORM regulation is measurement. Three types of measurement are of
utmost importance for the purpose of regulation: radon exhalation flux density from material
surfaces, concentration of radon in indoor air and activity concentration of natural
radionuclides in materials. Radon exhalation flux density is commonly measured using the
activated charcoal canister technique and is recommended to be used for NORM sites by
regulatory bodies (e.g. ARPANSA, 2008). This technique is based on the assumptions that radon
29
exhalation is constant over time and all of the exhaled radon is adsorbed and retained in the
charcoal. While a number of studies have shown the limitations of activated charcoal canisters
due to meteorological conditions, the validity of the stated assumptions is not fully understood.
Therefore, further investigation on the measurement of radon exhalation flux density using the
activated charcoal technique would provide a unique opportunity to examine the set of
assumptions and better understand the potential limitations of such a widely-used technique.
Released radon from materials is either diluted in the open air or accumulated in
enclosed spaces. As radon is readily diluted, its concentration in the open air is of less interest
for radiological protection. More attention has been given to radon in indoor environments.
Epidemiological studies have demonstrated the association between radon exposure and lung
cancer. The first step to control such exposure is to determine the level of radon in the indoor
environment, whether in dwellings or in workplaces. The previously reported data indicate the
dependency of indoor radon on the underlying geology of the region and ventilation systems.
Seasonal and diurnal variations on indoor radon levels have also been observed. However, there
are limited data on indoor levels of radon, as well as thoron, in Australia and other parts of the
world. National and international regulatory bodies encourage conducting measurements of
indoor radon in order to assess radon levels in a country and to establish a reference level as
well as an action plan for controlling radon exposure. Conducting such measurement in
workplaces would be beneficial to assess the levels of radon and thoron in those places and to
investigate the parameters that may influence those levels.
The third important measurement for regulatory applications is quantifying the activity
concentration of natural radionuclide members of NORM in materials. Although alpha, beta and
mass spectrometry were used to measure the radioactivity in materials, these methods involve
chemical separation, which is destructive. Gamma spectrometry is unique in this manner as it is
not destructive and can measure many radionuclides simultaneously. While several types of
gamma spectrometry have been developed, the BEGe detector represents the latest technology
of these detectors with high resolution and good efficiency over a wide spectrum of gamma
energies. This detector has been extensively used in nuclear physics and medical science.
However, in environmental science and NORM, the use of the BEGe detector is very limited. The
employment of such a detector will provide a valuable tool for the routine analysis of NORM.
This will be accomplished with the investigation of the potential use of the BEGe detector in the
field of NORM.
30
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Chapter 3 NORM Regulation: Better understanding for
better management
Sami H. Alharbi*, Riaz A. Akber
*School of Physical and Chemical Sciences, Queensland University of Technology, 2 George
Street, Brisbane, Qld 4000, Australia
52
Statement of Contribution of Co-Authors for Thesis by Published Paper The authors listed below have certified that:
1. they meet the criteria for authorship in that they have participated in the conception,
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expertise;
2. they take public responsibility for their part of the publication, except for the
responsible author who accepts overall responsibility for the publication;
3. there are no other authors of the publication according to these criteria;
4. potential conflicts of interest have been disclosed to (a) granting bodies, (b) the editor
or publisher of journals or other publications, and (c) the head of the responsible
academic unit, and
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53
Abstract
Naturally occurring radioactive material (NORM) is present in various aspects of human
lives. Radiation exposure resulting from NORM can be hazardous to humans and the
environment. Such exposure may be increased as a result of certain human activities involving
exploration, processing and production of raw materials. The regulation of NORM has been
given more attention since the 1980s, when the enhanced level of NORM was identified in the
petroleum industry. The presence of NORM was also observed in many other industries
involving minerals and raw materials production, including phosphate, zircon and rare earth.
National and international regulatory bodies attempted to control NORM exposure in some of
these industries by applying the radiation protection system that has been designed and
developed primarily for artificial radionuclides. This has created regulatory issues and
challenges that were identified and discussed in this review on the light of best available
knowledge. In addition, the latest findings on radon have confirmed the health risks not only for
workers, but for dwellings residents as well. The implication of the implementation of the
recent ICRP and IAEA recommendations and policies on indoor radon and thoron protection
were also examined in this review.
3.1 Introduction
Radiation protection is a field that has evolved over the last century. It is a dynamic
concept, with its underlying principles and guidelines being regularly revised, reworked and
amended. The main goal of the radiological protection system is to protect humans and the
environment from the harm of ionising radiation, regardless of its size and origin (ICRP, 2007a).
However, NORM has received less attention than other sources of ionising radiation, although
the resultant exposure from NORM can be more significant (Steinhäusler, 2010). Interestingly,
certain human activities involving exploration and production of raw materials have been found
to cause alterations to levels of NORM (IAEA, 2006a), resulting in a potential increase in
radiation exposure to workers, the public and environment. This source of exposure, in most
cases, has not been introduced deliberately and, more importantly in other cases, it is increasing
as new situations are identified. Human and non-human species need to be protected against
such a source of exposure.
Recently, there has been a growing tendency to protect mankind and the environment
against exposure to NORM caused by human activities. This requires a greater focus on the
management of NORM in different practices and its related aspects. The aim of this chapter is to
review the radiological protection system related to naturally occurring radioactive materials
and highlight critical issues associated with NORM that need to be considered when establishing
54
a framework for regulating NORM. The information presented in this chapter has been compiled
from publications provided by the following organisations: ICRP, IAEA, UNSCEAR, WHO, EU,
ARPANSA, and USEPA. Relevant scientific studies and reviews available in the literature have
also been used.
3.2 Overview of the radiological protection system
3.2.1 A brief history of radiological protection
Some literature reviews have described the historical development of radiation
protection organisations worldwide (Clarke and Valentin, 2005; ICRP, 2009b; Walker, 2000). A
summary of these references is outlined in this Section.
The story of ionising radiation began after the discovery of X-rays and radioactivity in
the 1890s. The use of X-ray devices then became widespread throughout medical and scientific
disciplines. It soon became clear, however, that exposure to this form of radiation caused
numerous adverse effects among patients, physicians and scientists. The same phenomenon
was observed after the discovery of radium. At that time, a lack of knowledge regarding the
hazards presented by radiation led to serious illness and death for those handling radium, such
as watch dial painters (ICRP, 2009b; Walker, 2000). In the early years of the 20th century, the
international community recognised the harm caused by exposure to ionising radiation, as well
as the need to protect people and workers from excessive amounts of radiation. Hence, a
number of national and international organisations have been created for the purpose of
radiation protection.
The First International Commission on Radiology was formed in London in 1925 by a
number of radiologists, physicians and scientists. Their hope was to formulate consistent
radiological metrics and measurements to mitigate and control exposure to ionising radiation.
Three years later, the Second International Commission on Radiology established the
International X-Ray and Radium Protection Committee (IXRPC). At that meeting, the first report
on radiological protection was published. The report did not include any quantitative
recommendations; it focused mainly on shielding needed to protect workers from harmful
exposure (IXRPC, 1928). The IXRPC was later renamed, in 1950, the International Commission
on Radiological Protection (ICRP).
With the growing use of radioactive substances and nuclear energy, there was a need to
establish a committee to collect and evaluate scientific information on the effects and levels of
ionising radiation. This was the main reason for creating the United Nations Scientific
Committee on the Effects of Atomic Radiation (UNSCEAR) in 1955. Two years later, the
International Atomic Energy Agency (IAEA) was established in response to the deep fears and
expectations resulting from the discovery of nuclear energy. These international organisations
55
have worked to protect humans and the environment from the harm of ionising radiation, and
to provide a framework to achieve their goal (ICRP, 2009b).
3.2.2 Quantitative recommendations
In 1934, the radiological protection system entered a new era with the introduction of
numerical recommendations for limiting occupational exposure. The IXRPC recommended a
dose limit of one roentgen per week for occupational exposure to X-rays (IXRPC, 1934). This
corresponded to an annual occupational effective dose of 500 mSv. It was a remarkable
improvement for the radiological protection system. A few years later, the IXRPC revised this
figure, believing the dose was too close to the theoretical threshold for adverse effects to occur.
Thus, the limit was reduced to 0.3 roentgen per week, which corresponded to an annual
occupational effective dose of 150 mSv for both X-rays and gamma radiation (ICRP, 1951). The
report included a table of relative biological effectiveness (RBE) and maximum permissible
body burdens for eleven nuclides, including 226Ra.
The 1954 ICRP report contained further improvements to the quantitative
recommendations. Before that report, there was no such limit set for general members of the
public; the ICRP then recommended a maximum permissible dose for the public as a tenth of the
occupational limit (ICRP, 1955). The concept of the critical organ was introduced in the same
report to relate the dose limit to whole body exposure. The numerical values were reduced in
1958 for both occupational workers and members of the public to 5 rem (50 mSv) and 0.5 rem
(5 mSv), respectively. In this recommendation, the dose limits were replaced by the limits for
the accumulative dose equivalent (in rem), which was derived from the RBE.
With advanced developments in the fields of physics and biology, it was realised that the
degree to which ionising radiation affected biological tissues varied with different types of
radiation and the nature of the tissues and organs being exposed. As a result, the effective dose
limit was introduced in 1991 to take into account the type of radiation (radiation weighting
factor) and sensitivity of different organs (tissue weighting factor) exposed. The annual effective
dose limit was adjusted to 20 mSv for occupational exposure and 1 mSv for the general public.
The occupational and public dose limits have remained unchanged since the ratification of
the ICRP recommendations in 2007. However, based on new information, the ICRP changed the
emphasis on the radiation sensitivity of different organs in the human body (ICRP, 2007a).
These changes included increased tissue weighting factors for the breast (due to the higher
breast cancer risk), decreased tissue weighting factors for the gonads (due to the reduced
hereditary risk) and the introduction of new tissue weighting factors for the brain and salivary
glands (due to the increased cancer risk of these two organs).
The impact of changing the tissue weighting factors was investigated in a study conducted
by Lee et al. (2007), which concluded that the effective dose determined in 1991 was influenced
56
more by the gonads than the breasts, and vice versa in the effective dose of 2007. Lee et al.
attributed their findings to the significant changes in the tissue weighting factor for both the
gonads (from 0.20 to 0.08) and the breasts (from 0.05 to 0.12), as per the ICRP 2007
recommendations.
3.2.3 The conceptual framework
In addition to quantitative recommendations, the principles and concepts of radiological
protection have developed. They have become integral to the radiological protection system
because the numerical advice was insufficient for a comprehensive framework. A number of the
current fundamental principles of radiation protection were implicitly stated in the earlier
recommendations. For example, the terms “maximum permissible exposure” and “probable
threshold for adverse effect” were used in the recommendations of 1951 to imply the principle
of the dose limit. Hence, the main principle of radiation protection was to keep all exposure
below the relevant thresholds to prevent deterministic effects and to minimise stochastic
effects. However, the growing need to balance the costs, benefits and detriments in regards to
radiation exposure led the ICRP to introduce the current principles of radiation protection in
1977: justification, optimisation and limitation (ICRP, 1977).
In the 1990 recommendations, the ICRP introduced a new description for human
activities in which it categorised them into practices and intervention (ICRP, 1991). Practices
increase the overall exposure to radiation due to the introduction of new sources or the
modification of exposure, while intervention decreases the overall exposure by influencing
existing sources of exposure. These concepts were replaced by exposure situations in 2007,
when the ICRP distinguished between three types of exposure situations: planned, existing and
emergency (ICRP, 2007a). The three fundamental principles of radiological protection remained
unchanged.
A special concept was proposed for protecting general members of the public and
known as the critical group. This was a group of individuals who received the highest radiation
dose from a particular source for the purpose of compliance with the regulation (ICRP, 1965).
However, the expression “critical group” was found to be confusing because the adjective
“critical” may be understood as meaning crisis, which was not the intended purpose.
Additionally, the word “group” usually refers to a few tens of people and does not reflect the
individual’s dose being assessed (ICRP, 2007a). Therefore, the “critical group” was replaced by
the concept of the “representative person” (ICRP, 2006). The representative person is defined as
an individual who receives a dose that is representative of more highly exposed individuals in
the population. It is expected that regulatory authorities throughout the world will eventually
phase out the use of the concept of the “critical group” in favour of the “representative person”.
57
3.2.4 Protection beyond humans
The radiological protection system was initially designed to protect humans against
ionising radiation. It was thought that the level of safety required for the protection of human
beings was likely to be sufficient to protect other biota (ICRP, 1977). However, it is difficult to
determine whether the environment is adequately protected from the potential impact of
radiation exposure under different circumstances because there are no criteria that address
environmental protection (ICRP, 2003). For that reason, the ICRP has proposed a mechanism to
assess the radiological impact of practice on living organisms, other than humans, which is yet
to be adopted by a number of regulatory organisations worldwide (ICRP, 2009a). This new
approach is based on the concept of Reference Animals and Plants (RAP). This concept has
defined by ICRP as: “A hypothetical entity, with the assumed basic biological characteristics of a
particular type of animal or plant, as described to the generality of the taxonomic level of family,
with defined anatomical, physiological, and life history properties, that can be used for the
purposes of relating exposure to dose, and dose to effects, for that type of living organism” (ICRP,
2009a).
A set of reference animals and plants has been reported and these were grouped at the
family level (ICRP, 2009a). However, certain types of animals belong to species that may be
associated with more than one type of habitat. For example, some frog species may be counted
more in terrestrial environments, while others are counted more in aquatic environments
(Hosseini et al., 2010). Consequently, the correct type of habitat needs to be associated with
each species.
3.3 The current system of radiological protection
The system of radiological protection has been developed as collaboration between
international organisations as illustrated in Figure 3.1. Basic scientific studies, which are the
primary sources of information, are evaluated and compiled by UNSCEAR. The ICRP in turn uses
this information along with other reports provided by national organisations, such as the
Biological Effects of Ionizing Radiation (BEIR) reports, to devise recommendations. These
recommendations are used by the IAEA to produce its Basic Safety Standards (BSS) that are
adopted by national authorities worldwide.
58
Figure 3.1 The basis of radiological protection regulation.
The current system of radiation protection consists of three exposure situations and
three protection principles to protect four individual categories (IAEA, 2014; ICRP, 2007a). This
system was developed based on a combination of three components: scientific knowledge,
practical experience and social and ethical values (Figure 3.2). Recommendations are
incorporated into regulations and standards, and adopted by regulatory authorities in many
countries worldwide.
Figure 3.2 The radiological protection system.
Basic scientific studies
UNSCEAR
(Reports)
ICRP
(Recommendations)
IAEA
(Standards)
National authorities
(Regulation)
Radiation Protection System
Principles
Justification
Optimisation
Dose limits
Situations
Planned
Emergency
Exsisting
Categories
Occupational
Medical
Public
Environmental
Values
Experience
Science
59
3.3.1 Exposure situations
The ICRP recognised three situations in which individuals could be exposed to radiation:
planned exposure, existing exposure and emergency exposure. These situations replaced the
previous approach of “practices” and “interventions” used by the ICRP. According to the ICRP,
the new situations are capable of covering the entire range of exposures (ICRP, 2007a).
The planned exposure situations involve the deliberate introduction and operation of
sources; they may result in the rise of both normal exposure (i.e. those expected to occur) and
potential exposure (i.e. those that are not expected to occur during the operation) (ICRP,
2007a). The previous operation, categorised as a “practice”, is included in this situation. In
planned exposure situations, the protection against ionising radiation can be planned in
advance of the installation of the source. The magnitude of exposure from a source in these
situations can be predicted and, hence, actions can be taken to control exposure.
Existing exposure situations are those that already exist when a decision regarding
control has to be taken, including prolonged exposure situations after emergencies. For
example, exposure caused by natural sources (e.g. radon in a dwelling or workplace and NORM).
Cosmic radiation is also considered by the ICRP as an existing exposure situation (ICRP, 2016).
Existing exposure can result from other sources that are related to some human activities (ICRP,
1999). The Chernobyl nuclear plant accident in the Ukraine, for instance, may be initially an
emergency exposure situation, but it resulted in the contamination of the surrounding
environment and the establishment of prolonged exposure (Krolak et al., 2012; Weiss, 2011).
Emergency exposure situations are any unexpected events that require urgent action to
avoid or reduce any undesirable consequences. A recent example is the nuclear reactor accident
in Fukushima, Japan, that occurred in 2011 during the operation of a planned situation. This
situation caused a release of various radionuclides into the environment and, consequently, an
existing exposure situation will be established after the accident has passed (Hirose, 2011; Taira
et al., 2012). Emergency and existing exposure situations require appropriate action to reduce
the exposure if it exceeds a certain reference level that are usually set by the regulatory
authority.
There is a likelihood that more than one exposure situation can exist simultaneously. In
this case, it is important to know that each exposure situation has its own merits and, therefore,
each exposure situation has to be treated separately. The exposure situation could start as a
planned exposure and end up as an existing exposure, for example, through an emergency event
which acts as a transition situation. As an example, Chernobyl started as a nuclear power plant
(planned exposure situation), then an accident occurred leading to an emergency exposure
situation and, eventually, environmental contamination is being controlled as an existing
exposure situation (Krolak et al., 2012; Weiss, 2011).
60
3.3.2 Principles
In 1977, the ICRP introduced the fundamental principles of radiological protection:
justification, optimisation and limitation. Justification is realised when the perceived benefits of
a radiation practice outweigh the harm associated with it. Optimisation aims to keep all
exposure as low as reasonably achievable, with economic and social factors being taken into
account. The third principle is the dose limits that are applied to ensure that individual
exposure does not exceed an acceptable level of risk. These principles ensure that the
deterministic effects are avoided and potential stochastic effects are reduced. The principles of
justification and optimisation are applied in all exposure situations and types. The principle of
dose limit is only applied in planned exposure situations, with an exemption for medical
exposure, while the reference level is used in existing and emergency situations.
3.3.3 Exposure categories
Individuals who actually receive, or will receive, exposure are grouped by the ICRP into
three exposure types: occupational exposure, which refers to all exposure incurred by workers
as a result of their work; medical exposure, which refers to any exposure received by patients as
the result of their treatment or diagnostics; and public exposure, which refers to anyone
receiving exposure that is neither occupational nor medical. Another type of exposure is known
as environmental exposure and it refers to any exposure received by any living organisms, other
than humans.
3.4 NORM in the radiation protection context
The presence of natural radionuclides in industries was recognised in the early 20th
century when radium and radon (or radioactive gas at that time) were detected in crude
petroleum (Burton, 1904; Himstedt, 1904) and in natural gas (Satterly and McLennan, 1918).
However, exposure due to such sources, and NORM in general, did not receive attention from
the radiological protection community until the mid-1980s, when elevated levels of NORM were
reported in oil and gas industry (Kolb and Wojcik, 1985; Smith, 1987). The exception was
uranium mining and milling because this was considered to be part of the nuclear industry. This
omission occurred because exposure to NORM was not considered to be a source of risk to
humans. It was framed in the context of natural background radiation, which was assumed to be
uncontrollable and constant everywhere.
However, this perspective was altered when the concentration of NORM was found to
vary considerably from one place to another depending on the geological and geographical
formation. In addition, some human activity resulted in an elevation or modification of NORM
concentration. Fortunately, exposure in a number of these situations can be controlled, such as
in the case of radon in dwellings and workplaces.
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The current radiation protection system is applied based on the exposure situation. All
exposure to ionising radiation from any source is subject to the radiological protection system,
regardless of its size or origin (ICRP, 2007b). Exposure due to natural sources of ionising
radiation including NORM is generally considered as an existing exposure situation. However,
exposure due to NORM is subjected to the requirements of planned exposure situations as
follows (IAEA, 2014):
i. Exposure due to material in practices where the activity concentration of any
radionuclide in the uranium or thorium decay chains exceeds 1 Bq g-1, or 10 Bq g-1 for
40K in that material
ii. Public exposure delivered by discharges or arising from radioactive waste of the above
practices
iii. Exposure due to 222Rn and 220Rn and their progeny in workplaces in which occupational
exposure, due to other radionuclides in the uranium or thorium decay chains, is
controlled as a planned exposure situation
iv. Exposure due to 222Rn and its progeny in the workplace where the annual average of
222Rn activity concentration remains above the reference level.
Figure 3.3. Diagram of the current regulation of NORM.
Figure 3.3 shows regulation situations applicable to NORM. The regulation of NORM can
be divided into two headings: radon in dwellings and workplaces, and NORM industries. In
these headings the principle of a graded approach to regulation is adopted if the activity
NORM
Regulated
Existing Exposure Situation
222Rn and220Rn in dwellings and workplaces
Practices 238U,232Th ˂ 1Bq/g and 40K˂ 10Bq/g
Planned Exposure Situation
Practices238U,232Th >1Bq/g and 40K>10Bq/g
Radioactive waste arising from the above practices
222Rn and 220Rn associated with regulated practices
222Rn in workplaces > Reference level
Not regulated
Exposures deemed to be not amenable to control (e.g. 40K
in human body)
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concentration of any radionuclides in the uranium and thorium decay series exceeds 1 Bq g-1
and 10 Bq g-1 of 40K or the relevant reference level for 222Rn (IAEA, 2014). In that case, more
details have to be considered about the practice, including particular types of operation,
process, material involved and exposure dose to workers and the public (IAEA, 2006a). The
graded approach has four levels of regulatory control, which are exemption, notification,
registration and licensing (IAEA, 2014). Exemption is the basic level of the graded approach in
which the regulatory body may decide not to apply regulatory requirements to a material or
practice. The exemption is granted if a specific criterion, proposed by the regulatory body, is
met. If the regulatory body found the exemption was not the optimum option, the next level in
the graded approach would be required, in which the legal person must submit a formal
notification of the intention of operating a facility to the regulatory body. If the regulatory
authority found that exemption or notification were insufficient for the level of exposure,
authorisation would be required in a form of either a registration or a licence. The difference
between the latter two levels is essentially in the level of regulatory stringency. In registration,
only limited obligations on the legal person may be enough to ensure adequate protection, while
in licensing, enforcement of more stringent exposure control measures are more appropriate
for acceptable levels of protection. However, the use of licensing for practices involving
exposure to NORM is likely to be limited to material with very high activity concentrations
(IAEA, 2013a).
3.4.1 Radon in dwellings and workplaces
Radon exposure in dwellings and workplaces has been given more attention than other
exposure from natural sources. This is because exposure due to radon and its decay products is
the major contributor to public exposure from natural sources (UNSCEAR, 2006) and the risks
of exposure to radon and its decay products have been confirmed by several recent
epidemiological studies (Baysson et al., 2004; Darby et al., 2005; ICRP, 2010; Vacquier et al.,
2009). Moreover, exposure to radon in the indoor environment is controllable as its parameters,
such as entry points and ventilation, can be controlled.
A series of recommendations and policies regarding exposure to radon in dwellings and
workplaces have been published recently (IAEA, 2015; ICRP, 2014; UNSCEAR, 2006; WHO,
2009). More discussion on the recent development of radon exposure regulation is provided in
Section 6 of this chapter.
3.4.2 NORM industries
NORM is present at enhanced levels in a wide range of industrial operations, including
exploration, extraction of raw materials, processing, products, by-products and waste. The main
industries in which NORM has been identified are: the oil and gas industry (Abo-Elmagd et al.,
63
2010; Al-Masri and Suman, 2003; Bakr, 2010; White and Rood, 2001), minerals extraction and
processing industries, including uranium and mineral sand (Ballesteros et al., 2008; Haridasan
et al., 2006; Righi et al., 2005), the bauxite and aluminium industry (Abbady and El-Arabi, 2006;
Jobbágy et al., 2009), the phosphate industry (Beddow et al., 2006; Tufail et al., 2006), iron and
steel production (Brown et al., 2003; Iwaoka et al., 2010), coal mining and coal-fired power
generation (Cevik et al., 2007; Parami et al., 2010), the building materials industry (Aslam et al.,
2012; Turhan et al., 2011) and water treatment (Kleinschmidt and Akber, 2008; Palomo et al.,
2010). Enhancements can be either deliberate, as in uranium mining where a higher ore grade
is sought, or inadvertent, as in petroleum industries where the activity is often associated with
the barite phase in produce water, scale and sludge.
Enhanced levels of NORM may have the potential to pose risks to human and the
environment (Vearrier et al., 2009). However, most exposure from NORM industries relates to
workers. The ICRP has devised a broad framework for radiological protection against NORM
industries, without including details of each type of operation (ICRP, 2007a). NORM industries
are simply regulated based on the intention of modifying or introducing the source. If the source
of exposure is modified intentionally in such a way as to increase the radiation dose, the source
falls under the scope of regulation. The specific details are the responsibility of national and
international regulatory bodies. However, in some situations, the national authorities may lack
experience in dealing with NORM.
The IAEA has published a number of standards for managing NORM in a small number
of industries, including the oil and gas industry (IAEA, 2003), minerals and raw materials
production (IAEA, 2006a) and the phosphate industry (IAEA, 2013b). The concept of
exemptions, clearance and exclusion are preferable for most exposure situations in industries
involving NORM (IAEA, 2005).
3.5 Regulatory issues with NORM
In recent years, there has been a growing need to control exposure caused by NORM.
National and international organisations are aware of the importance of managing exposure to
NORM. Some steps have been taken in this direction so far. However, until now, there has been
no complete vision of the regulation of NORM. A broad approach has been used through the
principle of optimisation utilising the concept of exemptions, clearance and exclusion for most
NORM exposure situations (IAEA, 2005).
The controllability of exposure plays an essential role in determining the scope of
regulation (ICRP, 2007a). Controllable exposure includes those situations in which radiological
protection actions are warranted to reduce exposure, such as dwellings and workplaces where
the radon concentration in air exceeds IAEA, ICRP or a national recommended reference level.
64
In other situations, exposure to NORM is uncontrollable, or is not amenable to control, which
means that exposure cannot be controlled by regulations and protective actions are not feasible.
For example, public exposure due to 40K in the human body is considered to be an uncontrolled
situation and therefore, is excluded from radiological protection legislation.
Regulatory authorities may face a number of issues in attempting to establish a regulation
standard for managing exposure to NORM. Nine such issues have been repeatedly found to be
central components of the regulation of NORM. These issues have a direct influence on the
decision regarding managing exposure to NORM. The issues are listed and discussed as follows:
3.5.1 Absence of a complete framework
There has been no comprehensive framework for regulating NORM in non-nuclear
activities. This is reasonable, as NORM regulation was ignored for a long period of time and
most importantly, it has to be treated case-by-case. The characteristic of NORM exposure varies
depending on the type of industrial activity. Studies have shown that the behaviour of individual
radionuclides of NORM can differ from one practice to another and even from field to field for
the same practice (Haridasan et al., 2015; O'Brien and Cooper, 1998; Xhixha et al., 2013). Dose
assessments and exposure pathways change accordingly across practice. For example, in the
petroleum industry, exposure to NORM is caused by radium isotopes, which dissolve in the
produced water and are brought to the surface, precipitated in sludge and scale (Abo-Elmagd et
al., 2010; IAEA, 2003), while in the phosphate industry, they present more in residue,
phosphogypsum, and tend to leach into groundwater (Beddow et al., 2006; Cañete et al., 2008;
Haridasan et al., 2002; IAEA, 2013b). Thus, the criterion for controlling NORM in the oil and gas
industry is not appropriate for the application to the phosphate industry. For this reason, the
regulation applied to NORM in one industry cannot be generalised to others.
Regulatory bodies have recognised this issue and even in the upcoming NORM
framework currently being prepared by the ICRP, the approach is going to be broad enough to
cover a wide range of industrial sectors involved without entering into detail (Lochard, 2015).
Currently, guidance for managing NORM has been issued for a limited number of industries,
including oil and gas, phosphate, rare earths, titanium and zircon and zirconia industries (IAEA,
2003, 2007a, 2011, 2012, 2013b). For other industries where NORM has been identified, (e.g.
geothermal power installation and recycling of metal scraps), guidance is still unavailable.
Meanwhile, general guidelines (e.g. IAEA 2004, 2006a, 2013b) may be applied. The management
of NORM, therefore, should start with identifying industries in which elevated concentrations of
natural radionuclides can be found. It should also characterise exposure, practice and
radionuclide behaviour, as well as resolve waste and residue issues.
65
3.5.2 The interference between planned and existing exposure situations
In the recent ICRP situation-based approach, two situations have been applicable to
exposure from NORM: planned exposure situations and existing exposure situations (ICRP,
2007a). They correspond to the practice and interventions of the previous approach (ICRP,
1991), which are still being used by many national organisations. By definition, exposure to
natural sources is an existing exposure situation. However, if exposure is directly related to the
work, the protection requirements for occupational exposure in planned exposure situations
may be applied. When applying a radiation protection system, each exposure situation is treated
separately, and the radiological protection system is applied according to the exposure
situation.
Sometimes it is difficult to identify whether exposure due to NORM is an existing or
planned exposure situation (Haridasan, 2015; Hedemann-Jensen, 2010). This issue has been
mentioned by the IAEA, which has admitted that in some circumstances the distinction between
exposure situations may be vague (IAEA, 2014). In its recent International Basic Safety Standard
(IBSS), the IAEA (2014) declared that “…some exposures due to natural sources may have some
characteristics of both planned exposure situations and existing exposure situations”. Even in
the previous approach, i.e., practice and intervention, this issue was observed. In waste disposal,
the ICRP found that distinguishing between practices and intervention is not always easy (ICRP
1997).
In situations where the boundary between planned and existing exposure situations is
unclear, it is the national regulatory authority's responsibility to decide which exposure to
NORM should be considered as planned, and which should be considered as an existing
situation (ICRP, 2007b). Most national regulatory authorities around the world, however, may
lack experience and may have not yet adopted the new situations-based approach, with the
practice and intervention-based approach still being used. The regulatory authorities in these
countries may face challenges when transferring their regulation systems to the new approach
and when deciding which exposure situation is applicable to NORM. This may result in
confusion for these national regulatory authorities.
3.5.3 The lack of a homogeneous approach
One of the major challenges facing regulatory authorities is the inconsistency in the
protection system against exposure to NORM. Indeed, the ICRP has recognised the variability of
the regulation of exposure to NORM in its recommendation. The ICRP (2007b) declared that
there is a need for an international consensus on including or excluding exposure to NORM from
the scope of regulations. The main cause of this inconsistency may be the absence of guidance
on the protection against exposure to NORM. Furthermore, the definitions of the scope of the
66
NORM regulations vary across countries, which means that what is regulated in one country
may be exempted in other countries.
Another possible cause of the heterogeneous regulation of NORM could be the rapid
progress in the radiological protection system, which may prevent national authorities keeping
pace with new policies. Consequently, the latest standards are likely to be adopted in some
countries sooner than others. The language and desire to be consistent can also be effective
factors in this regard (Rochedo and Lauria, 2008). The terms and phrases used in the
recommendations may be interpreted or translated incorrectly. For example, the reference level
may be interpreted as a limit, which is not what is meant by the ICRP.
Inconsistency appears mostly in the variations in the clearance and exemption levels of
radionuclides in the natural decay chains. Within Australia, for example, the exemption level of
226Ra is 10 kBq in Queensland, while in the bordering state, New South Wales, it is 40 kBq
(ARPANSA, 2005). There is also inconsistency in the terms used to describe NORM in the
policies. For example, in the US, the acronym TENORM is used for enhanced level of NORM
industries (USEPA, 2008), while in IAEA publications, TENORM is included in the acronym
NORM (IAEA, 2007b).
The absence of a homogeneous approach to regulate NORM may have economic, social
and political impact. The international trading of raw materials has been affected by this lack of
a harmonised approach (Wymer, 2007). For instance, some industries may migrate to countries
or territories with less strict regulatory systems. Consequently, workers may move to a new
place or lose their jobs, which in turn would lead to social and financial difficulties.
Nevertheless, much effort has been recently devoted to achieving a homogeneous
system to regulate NORM across international and national jurisdictions. At the international
level, the IAEA has published a general safety guide that is directly applicable to NORM: on the
Classification of Radioactive Waste (IAEA, 2009), in addition to a safety guide on the application
of the concepts of exclusion, exemption and clearance (IAEA, 2005). The ICRP is preparing a
guideline for NORM, which will be available in the near future. On the national level, in Australia
for example, the Australian Radiation Protection and Nuclear Safety Agency (ARPANSA) has
recently established the National Directory for Radiation Protection to provide a uniform
framework for radiation safety, including NORMs to be adopted by all jurisdictions across
Australia (ARPANSA, 2013). More efforts are still required to overcome this problem.
3.5.4 Ubiquity of NORM
Naturally occurring radionuclides are ubiquitous in all materials and, to some extent, all
materials are NORM. It is present naturally in sand, minerals, ores, products, by-products,
commodities, foodstuff, air, water and species' bodies. This nature of NORM can create unaware
67
complex exposure situation, which can deliver a radiation dose to human and non-human
species. According to Steinhäusler (2010), the lack of awareness of the presence of NORM by
many industries has resulted in the release of millions of tons of NORM into the environment
without control. It may last for a long time post-operation and the resultant contamination can
extend to the surrounding environment and cause a prolonged exposure situation.
Cumulative evidence has shown a concentrated level of natural radionuclides in vegetables in
the vicinity of coal-fired stations (Baeza et al. 2012; Papp et al. 2002) as well as around
petroleum industry facilities (Rajaretnam and Spitz, 2000). The challenge that regulatory bodies
may face is the determination of which industrial process can enhance the concentration of
NORM, and which of them should fall within the scope of regulation.
3.5.5 Source of low dose
The current radiation protection system is based on the linear no-threshold (LNT)
model. This model is used to extrapolate the cancer risks associated with the exposure dose.
According to this model, the carcinogenic risk is linearly proportional to the radiation dose. This
means that even a very low dose ˂ 1 mSv can carry a potential risk (BEIR, 2006). In contrast,
evidence has shown that a low level of radiation can be beneficial, as the radiation hormesis
model predicts (Feinendegen, 2014).
The LNT model adequately describes the dose response at a high radiation dose;
however, at low doses, it is controversial. Statistical errors in epidemiological and experimental
carcinogenesis studies are dominant at low (˂ 10 mSv) and very low (˂ 1 mSv) doses, as well as
the dose rate (ICRP, 2005). These errors have raised doubts about the LNT model for estimating
cancer risk at low doses. The BEIR (VII) report has recognised the validity of the LNT model, and
has concluded that even the smallest dose has the potential to cause an increase in cancer risk
(BEIR, 2006).
On the other hand, the French Academy of Sciences reported that the biological
mechanisms at low doses are different from those at high doses. The academy concluded that
the use of the LNT model to assess the risk of doses below 20 mSv is unjustified and,
consequently, should be discouraged (Tubiana et al., 2005). Apparently, the LNT fails to
estimate the cancer risk at the low dose and the low dose rate region (Cohen, 2012).
Comprehensive reviews of recent findings on the effect of low dose levels have been published
(Averbeck, 2009; Calabrese and O'Connor, 2014; Dauer et al., 2010; ICRP, 2005; Tubiana et al.,
2006).
What is apparent is that the effects of low-level radiation doses are not fully understood.
The result may cause doubts about the current and future regulation of exposure due to NORM,
68
which may be considered as a source of low dose exposure. Solving the dilemma of dose
response in the low dose zone will reflect positively on the regulation of NORM.
3.5.6 Societal attitudes
The goal of the radiation protection system is to protect people and the environment
from the harm of ionising radiation, regardless of its size and origin (ICRP, 2007a). However,
natural radionuclides have received less attention than artificial radionuclides. This is perhaps
because of the lack of public awareness of the risks that exposure to natural sources may cause.
People’s perceptions about the risks from radiation are usually associated with artificial
radionuclides, which are well-regulated and contribute little to overall public exposure (Lee,
1992), although exposure to workers and members of the public from natural sources
considerably exceeds that from nuclear technologies (Steinhäusler, 2010). These perceptions,
usually voiced loudly, have led the radiological protection community to focus more on artificial
radionuclides. This attitude may have historical roots; developed with time from occupational
exposure due to X-rays and 226Ra to the arrival of nuclear power (ICRP, 2007b). From a public
point of view, nuclear power and weapons are related to artificial radionuclides, while natural
radionuclides are related to natural background radiation. In addition, the potential of a
catastrophic accident, such as a nuclear accident, does not exist in the case of NORM.
Societal attitudes have a large influence on the final decision of the level of radiation
safety (ICRP, 2007a). They are related to the principles of justification and optimisation (ICRP,
1977). The influence of societal attitudes can be shown clearly in the variability of decisions by
regulatory bodies regarding their definitions of the scope of NORM regulation from one region
to another (ICRP, 2007b). Thus, in many countries, NORM is not regulated by nuclear
commission bodies but by local jurisdictions or environmental-related bodies. For instance, in
the US, NORM is regulated by the Environmental Protection Agency (EPA), while man-made
radionuclides are regulated by the Nuclear Regulatory Commission (NRC). Nevertheless, the
goal of radiation protection needs to be achieved regardless of the source of radiation, whether
natural or artificial (ICRP, 2007b).
3.5.7 Measurement challenges
In the regulation context, measurements are necessary to verify compliance with the
requirements of protection and safety. Measurements are performed during, pre- and post-
operation phases for several reasons and applications, including risk assessment and periodic
monitoring checks of sources.
As mentioned earlier, NORM contains a range of radionuclides that vary in their
chemical and physical properties. They also interact with the ambient environment and can
migrate to other areas (Michalik et al., 2013; Swarzenski et al., 2003). Therefore, quantifying the
69
activity concentrations of individual radionuclides of NORM and estimating the risk assessment
are not straightforward tasks. They involve several complex parameters and exposure
pathways. Modelling and simulations are usually employed for estimating the risk assessment
for NORM (O'Brien and Cooper, 1998; Pontedeiro et al., 2007).
Measurement is an important factor that is considered when establishing a regulation
standard (IAEA, 2014). The current available techniques and associated difficulties are reflected
in the regulatory standards. Using radon regulation as an example, the actual health risks are
caused by its progeny; however, reference levels are usually given in terms of radon gas
concentrations as a surrogate for radon progeny. The reason is related to the measurement
simplicity and cost-effectiveness of radon gas over the progeny measurement (WHO, 2009). On
the other hand, the data on thoron gas and its progeny is scarce because of their measurement
difficulties. Regarding its impact on the regulation, thoron receives less attention than radon,
even though the health risk of thoron progeny is comparable with that of radon progeny (Shang
et al., 2005).
The measurement and management of NORM are interrelated. The improvement of the
measurement techniques of NORM will have a positive impact on its regulation. National
authorities may need to provide procedures and protocols for the measurement techniques to
assist operators and ensure the accuracy of results. For example, the USEPA has published
protocols that provide a technological guidance for measuring radon and its decay products in
homes (USEPA, 1993).
3.5.8 Waste and residues
NORM, in most industries, presents accidently and therefore, it is mainly concentrated in
waste and residues, which are the main pathways of the release of NORM from practice to the
environment (O'Brien and Cooper, 1998). NORM in waste and residues can have a negative
impact on humans and the environment (Michalik et al., 2013; Steinhäusler, 2005). It can
interact with the surrounding environments and individual radionuclides may leach and
transport causing contamination to the environment (Rajaretnam and Spitz, 2000). Once
radionuclides are released or leached from waste sites, it becomes difficult for them to be traced
in the environment.
Numerous studies have reported contamination of the surrounding environment due to
the discharge of NORM from industries (Boryło et al., 2012; Cañete et al., 2008; Michalik, 2008;
Paschoa, 1998; Pontedeiro et al., 2007; Thiry and Van Hees, 2008). Zeevaert et al. (2006)
observed 210Po in food crops near a coal-fired power plant. They attributed the presence of 210Po
to the atmospheric discharge of fly-ash from that facility. Rajaretnam and Spitz (2000) also
detected a concentrated level of 226Ra in vegetation in the surrounding area of a petroleum
70
industry site. The IAEA has acknowledged this issue by publishing a number of guidelines on
NORM which focus mainly on the management of waste and residues (IAEA, 2003, 2007a, 2012,
2013b).
Industrial waste and residues are usually generated in large quantities in different forms
and, most importantly, they contain long-lived radionuclides. The activity concentrations of
these radionuclides are found to be elevated in several industries (Abo-Elmagd et al., 2010;
Chanyotha et al., 2012; Hilal et al., 2014). The management of such waste and residues is one of
the major challenges facing the radiological protection community. Regulatory bodies strongly
encourage recycling and reusing of residue rather than disposing it of as waste (IAEA, 2006b).
This would depend on the type of residue, the rate of generating it, location of the facility and
market conditions (Haridasan, 2015). An example of recycling residues is the use of
phosphogypsum, which is a residue of a phosphate fertiliser processing plant, as a by-product to
manufacture building materials (Cooper, 2005; Somlai et al., 2008). However, utilising residues
requires large spaces for storage and may also need special treatment, which may raise an
economic issue.
The recycling strategy is important to be considered for managing NORM waste and
residues. However, if recycling is not feasible, then the residues have to be treated as waste
(IAEA, 2013a). A graded approach is recommended to regulate and manage radioactive waste in
practices involving NORM (IAEA, 2006a, 2013b). This will require characterising NORM in that
industry and ensuring that the waste management plan is safe, technically optimal and cost-
effective (IAEA, 2013a). There are situations where the regulation of NORM waste is unclear,
such as waste containing both NORM and artificial radionuclides in a mix. In such situations, the
rules designed and developed for one category of radioactive waste cannot be easily applied to
another (Michalik et al., 2013).
3.5.9 Ethical approach
In 1977, the ICRP introduced the three fundamental principles of the radiological
protection system: justification, optimisation and limitation. These principles represent the
ethical foundation of the radiation protection system. The justification principle ensures that
the benefit outweighs the harm and the optimisation principle aims to keep all exposure as low
as reasonably achievable under the prevailing circumstances. These two principles correspond
to utilitarian ethics and the third principle, the dose limit, satisfies deontological ethics
(Hansson, 2007). The justification also represents a virtue ethic. In utilitarianism, actions are
judged by their overall consequences. The positive and negative effects of actions are weighted.
The benefits or the good of actions must be maximised and outweigh the harm. On the other
71
hand, deontology focuses on the rightness or wrongness of actions themselves, rather than their
consequences.
The ethical dimensions of radiological protection have been proposed mainly for man-
made radionuclides or planned exposure situations, in which the source is introduced
intentionally. On the other hand, NORM exposure is adventitious in most practices, and the
source is not deliberately introduced. Moreover, some industries can cause dissemination of
NORM radiation in consumer products, which may raise indoor exposure, such as in building
materials with elevated uranium and thorium content (Kovler, 2009).
An important ethical issue for the radiological protection system is the naturalistic
fallacy, in which ethical problems are reduced to purely scientific problems, as in the case of
natural background radiation (Shrader-Frechette and Persson, 2002). The author argues that
background radiation should not be used as a criterion for ethically acceptable exposure levels.
An alternative ethical dimension may be required to be implemented within, to the
protection from exposure to NORM. One suggestion is to introduce a pragmatic and flexible
approach to radiological protection against NORM (Cool, 2015; Loy, 2015). The pragmatic
approach is already applied in the nuclear industry, especially for waste, which may pose long-
term risks for humans and the environment (Blowers, 2010; Shrader-Frechette, 2013). In
addition, flexibility is expected to be recommended in the future ICRP guideline being prepared
specifically for NORM exposure, and it will apply to the application of dose limits in the planned
exposure situation (Lecomte, 2015).
3.6 Case study
Radon is a special case of NORM and is treated separately because of its possibility to
accumulate, capability to harm and feasibility to control. These have raised the awareness of
radiation exposure due to radon among international and, to a lesser extent, national
organisations. Consequently, a number of recommendations, standards and guidelines have
been published and others will be available in the future (IAEA, 2015; ICRP, 2014; WHO, 2009).
The publications provide a framework for managing and controlling exposure due to radon gas
and its progeny with the goal of protecting people against radon exposure.
There are three isotopes of radon: 222Rn, 220Rn and 219Rn derived from the three natural
decay series 238U, 232Th and 235U, respectively. The most significant source of radon exposure is
delivered by 222Rn due to its relatively long half-life (3.82 days) compared with 220Rn (55.6
seconds) and 219Rn (3.96 seconds). The exposure dose from the latter isotope is negligible due
to its short half-life and low abundance of 235U in nature.
72
3.6.1 Indoor 222Rn regulation
The current regulation of indoor 222Rn distinguishes between two exposure situations:
existing and planned (IAEA, 2014). Generally, indoor 222Rn in dwellings and workplaces is
managed as an existing exposure situation because it is a natural source of radiation and existed
when the decision of the regulation was made. In such a situation, workers are treated in the
same way as general members of the public and the protection from exposure is achieved by the
use of reference levels and optimisation. National authorities are responsible for setting an
appropriate reference level for 222Rn concentrations in indoor environments (IAEA, 2015). If the
concentration of indoor 222Rn exceeds the proposed reference level, intervention actions may be
needed. In workplaces, if for any reason 222Rn concentration after the intervention remains
above the reference level, then the requirements of occupational exposure in a planned
exposure situation have to be applied to workers (IAEA, 2014). Another situation in which
exposure to 222Rn is considered as a planned exposure is in a workplace where other
radionuclides in the uranium decay chains are managed as a planned exposure situation.
National authorities are encouraged to establish a radon action plan for controlling
222Rn exposure to the public in indoor environments. Recommendations and guidance have
been provided by international organisations in this regard (IAEA, 2014, 2015; ICRP, 2010,
2014; WHO, 2009). The 222Rn action plan is suggested to include the establishment of an
appropriate reference level, the identification of radon prone areas and the implementation of
actions to reduce 222Rn exposure in existing and new buildings (IAEA, 2015). The national
authority may add other components to the action plan according to the situation.
The reference level is the key parameter in controlling indoor 222Rn. Several reference
levels have been recommended for the purpose of protection against 222Rn exposure in
dwellings and workplaces. It is suggested to set reference levels based on an annual effective
dose of around 10 mSv per year (IAEA, 2015). However, for practical reasons, the reference
level is provided in terms of radon concentration. The conversion between the two quantities
requires a dose conversion factor to calculate the effective dose. It is usually associated with the
assumption of a 0.40 equilibrium factor for 222Rn indoors and annual occupancy factors of 7000
hours indoors or 2000 hours at work. Table 3.1 shows the development of the dose conversion
factors over the past few decades. These factors were derived using either the dosimetric
approach, which is based on the Human Respiratory Tract Model (HRTM) developed by the
ICRP (ICRP, 1994), or epidemiological approach, which is based on a direct comparison to the
risk to uranium miners (ICRP, 1993). Most recently, the ICRP reported that it intended to use
the dosimetric model again (ICRP, 2010).
Reference levels of less than 300 Bq m-3 are recommended internationally for 222Rn in
indoor environments, as reported in the latest recommendations of ICRP, IAEA, WHO and EC.
73
On a national level, some countries have not yet updated their 222Rn regulations, while other
countries’ indoor 222Rn levels are yet to be regulated. In many countries, 222Rn regulation is
limited to workers in a planned exposure situation (e.g. mines). In Australia, for example, a
reference level of 200 and 1000 Bq m-3 is recommended by ARPANSA for dwellings and
workplaces, respectively (ARPANSA, 2002). However, public exposure to indoor 222Rn is not
under the scope of regulation in any jurisdiction.
The tools for controlling indoor 222Rn exposure are preventive measures for new
buildings and corrective actions for existing buildings (WHO, 2009). The aim of these actions is
to prevent 222Rn ingress into facilities and reduce the current elevated 222Rn level in indoor
environments. There are some circumstances in which the regulatory authority decides when
the intervention actions are mandatory or voluntary (IAEA, 2015). It is always mandatory to
take corrective actions when the annual dose due to exposure to 222Rn in dwellings and other
buildings is greater than 100 mSv (IAEA, 2015). Co-operation with other governmental
organisations may require ensuring the effective implementation of an indoor 222Rn control
plan. The emanation of 222Rn from building materials should be controlled by limiting 226Ra
content in these materials (WHO, 2009). The release of 222Rn from the water supply may also
need to be controlled by following the approach provided by WHO (2011).
Table 3.1 The development of dose conversion factor over time
Factor Up-until Publication 65 Publication 103
Publication 115
Publication 126
1993 (ICRP, 1993) (ICRP, 2007a) (ICRP, 2010) (ICRP, 2014) Risk approach Dosimetric Epidemiological Epidemiological Dosimetric Dosimetric Protection level (mSv y-1) Action
level (10 ) Action level (3–10 )
Reference level (10)
Reference level (10)
Reference level (10)
Dose Conversion factor (nSv (Bq h m-3)-1)*
15 9 9 Under revision Under revision
222Rn level (Bq m-3)
Dwelling 200–600 600 300 300 Workplace 1500 500–1500 1000–1500 1000 300
* Based on UNSCEAR calculations in 1982, 1988 and 2000 taking into account radon progeny in terms of equilibrium equivalent concentration (EEC) which represents the concentration of 222Rn gas, in equilibrium with its progeny that would have the same potential alpha energy as the existing non-equilibrium mixture.
3.6.2 New policies and recommendations on indoor 222Rn
In view of the latest findings, relevant international organisations have revised and
updated their guidelines and policies for protection against 222Rn exposure (EU, 2013; IAEA,
2015; ICRP, 2014). The revised guidelines are focused more on public exposure due to 222Rn
indoors, because occupational exposure due to 222Rn in planned exposure situations is well-
regulated. The reference levels of 222Rn in dwellings and workplaces have been reduced, with
slight variations across international organisations. The ICRP recommends a reference level
ranging between 100 and 300 Bq m-3 for all buildings, without distinguishing between dwellings
and workplaces (ICRP, 2014). These replaced the previously recommended levels of 300 and
74
1000 Bq m-3 for dwellings and workplaces, respectively (ICRP, 2010). The IAEA has adopted the
same approach as an alternative to its main approach for indoor 222Rn policy. The IAEA
recommends a reference level that does not exceed an annual average 222Rn concentration of
300 Bq m-3 for dwellings as well as workplaces with a high occupancy factor for members of the
public. In workplaces with a low occupancy factor, the reference level of 222Rn concentration
should not exceed 1000 Bq m-3 (IAEA, 2015). The European Commission has reduced its
reference levels from 200 and 500 Bq m-3 to 100 and 300 Bq m-3 for both dwellings and
workplaces, respectively (EU, 2013).
The new policies emphasise the regulator’s role in providing information on indoor
222Rn levels in buildings. The information includes the average activity concentrations and
associated risks of 222Rn in dwellings and other buildings with a high occupancy factor (IAEA,
2015). This may be accomplished by conducting a national survey. It may also include
information about 222Rn testing techniques and protocols (ICRP, 2014). This information may
then be used to establish and implement an action plan for controlling public exposure due to
indoor 222Rn. Another new policy for controlling indoor 222Rn is the involvement of individual
responsibility, the so-called “legal responsible person” (ICRP, 2014). It is a responsibility of a
person of a building towards the users of that building. It could be stakeholders towards
occupants, an employer towards employees or house owners towards tenants. This legal
responsibility should be addressed and reflected in the indoor 222Rn protection strategy plan.
Although the health risk of indoor 222Rn exposure in smokers is higher than non-
smokers (Darby et al., 2005; Leenhouts, 1999), the new policies of the 222Rn protection system
do not distinguish between the two groups (ICRP 2014). Instead, national authorities are
encouraged to consider the conjunction of 222Rn policy with other health policies, such as non-
smoking and indoor air quality (IAEA, 2015), considering that ceasing smoking is likely to
reduce the health risks of 222Rn (Bochicchio, 2008; Méndez et al., 2011).
3.6.3 Implications of the implementation of the new policies
National authorities are encouraged to take steps to adopt the new recommendations on
indoor 222Rn. Despite the benefits of protecting humans against exposure to indoor 222Rn, the
implementation of the new 222Rn policies and recommendations may challenge national
authorities. Financial and practical implications are the main concerns of this section.
The cost implications of the new policies may affect the corresponding governmental
body as well as stakeholders and residents. The relevant authorities are required to undertake a
national survey in order to determine 222Rn exposure to the population and set an appropriate
reference level. The cost of such a survey depends on several factors, including the size of the
population, the measurement technique, staff training and sampling duration. Individuals,
75
whether stakeholders, employers or residents, may also be required to conduct measurements
for indoor 222Rn levels in premises under their responsibility.
The reduction of the reference level in the new recommendations means that more
buildings would be above the proposed reference level. Intervention actions in certain buildings
may be required to reduce the 222Rn level below the reference level. Some cost-effective
measures for existing and new buildings were highlighted in the WHO (2009) Report. These
extra costs can affect the decision of the legal responsible person or organisation regarding
complying with the requirement of 222Rn regulation and, consequently, influence the efficient
implementation of a protection strategy against indoor 222Rn.
In 2010, the ICRP announced that the nominal risk coefficient for 222Rn exposure should
be double the value previously reported and thus, new dose conversion factors will be
published in the near future. Subsequently, the cases in which corrective action is mandatory,
when the annual effective dose due to 222Rn concentration is greater than 100 mSv, are likely to
increase. In addition, some workplaces that may previously have been below the reference level
may now exceed it, if it is reduced in response to the new dose conversion factors.
Practically, the implementation of the new policies on 222Rn protection may cause some
confusion to the regulatory authorities. This is because the new policies are structured on a
situations-based approach of a radiological protection system, which replaced the
practice/intervention classification. Some countries are still regulating ionising radiation based
on the previous approach. The main confusion could arise from the fact that two situations are
applicable to indoor 222Rn, namely existing and planned exposure situations. This problem has
had a negative impact on the implementation of the standard and this has been recognised by
some authors (Chen, 2015; Lochard, 2015).
Another confusing issue is the level of occupancy factor mentioned by IAEA- BSS (2014).
Exposure to indoor 222Rn in buildings with a high occupancy factor for the public are treated as
dwellings, while buildings with a low occupancy factor are treated as workplaces, with a
correspondingly higher reference level and potential implementation of occupational exposure
to workers. There is no clear definition of when occupancy is deemed to be high or low, except
for some examples, such as schools attended by students and offices attended by visitors.
Instead, the IAEA recommends that the national authority decides which buildings should be
considered as having a high occupancy factor for general members of the public.
In recent years, there have been a number of changes in indoor 222Rn policies. These
changes may not be adopted by some authorities, which could create a heterogeneous pattern of
222Rn management across nations and states. National regulatory authorities need to keep pace
with the latest changes to avoid such situations.
76
3.6.4 The case of indoor 220Rn
Indoor 220Rn has received less attention than 222Rn due to the very short half-life of 220Rn
compared with 222Rn. However, the risk presented by the inhalation of 220Rn and its decay
products is considered to be equal if not greater than that of 222Rn because 50 % or more of the
total potential alpha energy concentration (PAEC) is caused by 220Rn progeny (Shang et al.,
2005). Therefore, determining the activity concentration of 220Rn alone may not reflect the
actual risk level.
The level of indoor 220Rn gas drops significantly as it travels away from walls or its
origin (Meisenberg and Tschiersch, 2011). It can lose more than 99 % of its value in the indoor
environment at a distance of 1 m from the wall surface (McLaughlin, 2010). However, 220Rn
progeny may remain in indoor air at a considerable level causing potential exposure to the
occupants.
On the other hand, the data available for 220Rn is limited, perhaps because of the
measurement difficulties associated with its short half-life (Tokonami, 2010; Zhang et al., 2010).
There is also lack of epidemiological studies on the exposure of 220Rn. These factors have a
negative impact on the current regulation of 220Rn. Perhaps it is for these reasons, 220Rn is not
regulated in most parts of the world, although the IAEA has included it in its recent guidelines
(IAEA, 2015). The IAEA recommends indoor 220Rn to be regulated only if necessary, when the
activity concentration of 220Rn in indoor air is high, but no value is provided. The control of
indoor 220Rn can be accomplished by controlling 232Th in building materials. The management of
indoor 220Rn is identical to that of 222Rn in indoor environments (IAEA, 2014). The same
procedure of developing an action plan and establishing a reference level for 222Rn is followed
for 220Rn with differences in dose conversion factors. The current dose conversion factor (DCF)
derived for 220Rn is 40 nSv (Bq h m-3)-1 for equilibrium equivalent concentrations (EEC) as
proposed by UNSCEAR (2000).
3.7 Conclusion
Protection against NORM is an ongoing concern for regulatory authorities. Exposure to
NORM is currently regulated with some ambiguity and confusion with the absence of a
comprehensive framework. It is considered as an existing exposure situation but also can be a
planned exposure situation. There are nine regulatory issues and challenges related to NORM
that have been highlighted in this review, including, heterogeneous approach, ubiquities nature,
wastes management, social attitudes, measurement methods and ethical values. These issues
created gaps in the current radiological protection system. Addressing these issues will not only
be crucial for characterising NORM exposure in different human activities, it will be essential for
improving the radiological protection system. The diversity of human activities involving NORM
77
and the complexity of exposure pathways dictate continuing efforts to formulate a framework
for protecting the public, workers and the environment. The framework should be broad
enough to accommodate a variety of industrial activities with enhanced levels of NORM,
including exploring, extraction, processing, production and disposal. New recommendations
and policies are currently underway with the aim of providing a clear and implementable
approach to regulating NORM.
78
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Chapter 4 Radon-222 Activity Flux Measurement
Using Activated Charcoal Canisters: Revisiting the
Methodology
Sami H. Alharbi*, Riaz A. Akber
*School of Physical and Chemical Sciences, Queensland University of Technology, 2 George
Street, Brisbane, Qld 4000, Australia
Journal of Environmental Radioactivity 129 (2014), 94-99.
http://dx.doi.org/10.1016/j.jenvrad.2013.12.021
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Statement of Contribution of Co-Authors for Thesis by Published Paper The authors listed below have certified that:
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In the case of this chapter:
Radon-222 Activity Flux Measurement Using Activated Charcoal Canisters: Revisiting the
Methodology
Contributor Statement of contribution
Sami
Alharbi
Signature:
Date: 31/03/2016
Conducted experimental work, prepared samples of analysis,
analysed and interpreted the data, wrote the manuscript
Riaz Akber Original idea, Assisted with editorial process,
helped in experimental design, reviewed the manuscript.
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87
Abstract
The measurement of radon (222Rn) activity flux using activated charcoal canisters was
examined to investigate the distribution of the adsorbed 222Rn in the charcoal bed and the
relationship between 222Rn activity flux and exposure time. The activity flux of 222Rn from five
sources of varying strengths was measured for exposure times of one, two, three, five, seven, 10,
and 14 days. The distribution of the adsorbed 222Rn in the charcoal bed was obtained by
dividing the bed into six layers and counting each layer separately after the exposure. 222Rn
activity decreased in the layers that were away from the exposed surface. Nevertheless, the
results demonstrated that only a small correction might be required in the actual application of
charcoal canisters for activity flux measurement, where calibration standards were often
prepared by the uniform mixing of radium (226Ra) in the matrix. This was because the diffusion
of 222Rn in the charcoal bed and the detection efficiency as a function of the charcoal depth
tended to counterbalance each other.
The influence of exposure time on the measured 222Rn activity flux was observed in two
situations of the canister exposure layout: (a) canister sealed to an open bed of the material and
(b) canister sealed over a jar containing the material. The measured 222Rn activity flux
decreased as the exposure time increased. The change in the former situation was significant
with an exponential decrease as the exposure time increased. In the latter case, lesser reduction
was noticed in the observed activity flux with respect to exposure time. This reduction might
have been related to certain factors, such as absorption site saturation or the back diffusion of
222Rn gas occurring at the canister-soil interface.
Keywords: 222Rn; activity flux; activated charcoal canisters
4.1 Introduction
Radon (222Rn) is a natural radioactive gas produced in rocks and soil by the
disintegration of its parent nuclei radium (226Ra). A small fraction of the produced 222Rn can find
its way to the surface and can be released from the ground. The activity flux of 222Rn released
from the ground is commonly measured by placing a chamber (accumulator) directly into the
sampling area in an upturned position (Ferry et al., 2002; Ielsch et al., 2002; Schery et al., 1989)
or by enclosing the sample materials in an airtight container (Jonassen, 1983; Samuelsson,
1990; Tufail et al., 2000). In both situations, the exhaled 222Rn from the sample surface enter the
chamber and the concentration of 222Rn increases. This concentration can be measured using
passive or active detectors. Passive devices, such as alpha track detectors and activated charcoal
canisters do not require electrical power to function, whereas active devices, such as
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scintillation cell and semiconductor detectors require electricity to operate. In both cases, the
flux density is then deduced using a proper formula.
Over the past few decades, a number of techniques have been developed to measure
222Rn activity flux from surfaces. A commonly and extensively used technique for measuring
222Rn utilises activated charcoal packed in canisters. It can be used to measure 222Rn
concentration in air (Cohen and Cohen, 1983; George, 1984; Scarpitta, 1992) as well as the 222Rn
activity flux as it escapes from the ground (Bollhöfer et al., 2006; Countess, 1976; Lawrence et
al., 2009; Spehr and Johnston, 1983; Wang et al., 2009) and building materials (De Jong et al.,
2005; Petropoulos et al., 2001). This technique is based on 222Rn adsorption in the charcoal,
which affords many advantages including: low cost, reusability and simplicity.
Since Countess (1976) published a paper on measuring 222Rn using an activated charcoal
canister, numerous studies have been conducted about the applications and limitations of this
method (El Samman and Arafa, 2002; Iimoto et al., 2008; Ronca-Battista and Gray, 1988; Vargas
and Ortega, 2007). These studies have shown that meteorological conditions, including
humidity, temperature and atmospheric pressure have the greatest influence on 222Rn
adsorption in activated charcoal. Most of these studies focused on the measurements of 222Rn
concentration in the air whereas only a few studies have been done on activity flux.
When 222Rn activity flux is measured, canisters are deployed on the target area and
exposed for a few days, typically between one to five days (Bollhöfer et al., 2006; Countess,
1976; Dueñas et al., 2007; Lawrence et al., 2009). The adsorbed 222Rn in the activated charcoal
is commonly measured through the gamma radiation detection of its short-lived progeny nuclei.
A sodium iodide (NaI(Tl)) detector is usually used for this purpose. The detection efficiency of
the gamma detector is determined by counting a standard canister, which is generally prepared
by adding a known amount of 226Ra to the charcoal and mixing it uniformly into the charcoal bed
(Bollhöfer et al., 2006; Spehr and Johnston, 1983). This may not always reflect the actual
situation as the distribution of the adsorbed 222Rn in the measured canister may not necessarily
have the same distribution as in the standard canister.
The 222Rn adsorption rate in this method is assumed to be constant over the exposure
period. Although this assumption may not be always valid in the real world, it has been widely
accepted for this method. This paper aims to visit two aspects of the current methodology of
measuring 222Rn activity flux using activated charcoal canisters. The first is the distribution of
222Rn in the charcoal bed. The second is the influence of exposure time on the measured value of
222Rn activity flux.
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4.2 Methods
A total of 293 individual measurements of the 222Rn activity flux were carried out for five
222Rn sources (Table 4.1). Of the canisters, 180 were used to examine the distribution of 222Rn in
the charcoal bed and 77 of them were used to investigate how the time influenced the measured
222Rn activity flux. All measurements were made under a controlled environment in the
laboratory. The relative humidity and temperature were 40%–60% and 22oC–25oC,
respectively. The canisters were cylindrical with a height of 0.060 m, a diameter of 0.070 m and
an open face area of 0.0038 m2. Each canister was filled with 25g of activated charcoal (Chem-
Supply, Australia). Prior to use, the canisters were annealed for 24 hours at 110oC to drive off
any residual 222Rn in the charcoal and then sealed with a polyethylene lid and adhesive tape.
4.2.1 The distribution of 222Rn in the charcoal bed
The charcoal bed in each canister was divided into six layers that could be sectioned and
counted separately for estimating the vertical distribution of the adsorbed 222Rn. Each layer was
filled with 5g of activated charcoal providing about 5 mm of layer thickness separated by a
metal mesh and secured with a wire spring. The canisters were exposed to different222Rn
sources for one, three, seven, 10, and 14 days. At the end of the sampling time, each layer in the
canister was transferred into a separate container and measured. The reason for counting each
layer separately was to maintain the same geometry throughout. The counting system used for
this purpose was a 2" × 2" NaI (Tl) detector. This setup is further described in Section 4.2.3.
4.2.2 The influence of exposure time on the measured activity flux
Activated charcoal canisters are integrating devices. When 222Rn activity flux is
measured, the obtained values actually represent the average activity flux over the exposure
time. In this procedure, the measured activity flux from the surfaces was examined as a function
of the exposure time. Two sets of experiments were performed, in the first, canisters were
placed directly onto the sources that were hosted on a tray. The tray had a diameter of 40 cm
and a depth of 5 cm and could host seven canisters simultaneously. Canisters were pressed to a
depth of 1 cm into the material to prevent gas leakage. In the second set, canisters were firmly
sealed over sources hosted in polystyrene cylindrical containers of 10 cm height and 6.3 cm in
diameter. The thickness of the materials inside the container was 8 cm. The open-tray method
may have more accurately represented field situation, but the material depth was only 3 cm.
Canisters in both sets were exposed for one, two, three, five, seven, 10, and 14 days.
4.2.3 Analytical technique
After the exposure, canisters were resealed and given four hours to allow the ingrowth
of 222Rn progenies 214Pb and 214Bi. The activity of the adsorbed 222Rn on the charcoal was
determined by detecting gamma rays that were emitted from its short-lived progenies 214Pb and
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214Bi using a 2" × 2" NaI (Tl) detector. Detector efficiency was 11% at the selected region of
interest (143 - 740 keV), which covered the salient peaks 214Pb at 242, 295, and 352 keV, as well
as 214Bi at 609 keV. The detector was housed in a lead castle to reduce background counts which
were obtained by counting an unexposed canister. The activity flux of 222Rn was inferred using
the following formula (Spehr and Johnston, 1983):
𝐽 =𝑅. 𝜆2. 𝑡𝑐 . 𝑒𝜆.𝑡𝑑
𝜀. 𝑎. (1 − 𝑒−𝜆.𝑡𝑐). (1 − 𝑒−𝜆.𝑡𝑒) (4.1)
where J is the measured 222Rn activity flux (Bq m−2 s−1), R is the count rate (s-1) after background
subtraction, λ is the decay constant (s-1) for 222Rn, ɛ is the efficiency of the NaI(Tl) detector
(cps/Bq), a is the surface area covered by the charcoal canister (m2), tc is the counting time, te is
the exposure time, and td is the delay time between the end of the exposure until the start of the
counting. For uniformity in units, tc, te and td were all measured in seconds. This formula is
based on two basic assumptions. The first assumption is that all of the 222Rn that enters the
canister is adsorbed by the charcoal and retained in it. The second assumption is that the 222Rn
activity flux from the surface is constant over the exposure period.
The results were given relative to the measured values of the activity flux after one day
of exposure. Equation (1), in that case, led to a relationship between the measured activity flux
after one day (J1) and n days (Jn) as
𝐽𝑛
𝐽1=
𝑅𝑛(1 − 𝑒−𝜆.𝑡𝑒1)
𝑅1(1 − 𝑒−𝜆.𝑡𝑒𝑛) (4.2)
where R1 and Rn are the count rates after the background subtraction (s-1) for one and n days of
exposure, respectively and te1 and ten are the exact exposure times for nominally one and n days
of exposure (s). Equation (2) assumed that the counting times tc1 and tcn were equal—1000
seconds during the present work. The delay times td1and tdn for one and n days of exposure
were the same for each sample (about four hours). Reported uncertainties represent one
standard deviation calculated from the statistics of counting.
In addition to the activated charcoal canisters and gamma count rate, an independent
method was also used to measure 222Rn activity flux from the five sources. This method was
based on the use of a solid-state alpha detector incorporated into two RAD7 systems
(DURRIDGE, 2000). The results of this independent method (Table 4.1) were compared with
those obtained from the activated charcoal canisters.
Table 4.1 Rn-222 emitting samples used in the present study.
Sample ID Description Method of use Expected activity flux(mBq m−2 s−1)a O – 1 Rock Closed-can 4528 ± 94 O - 24 Rock Closed-can 1427 ± 112 O - 31 Rock Closed-can 1551 ± 106 T Uranium tailing Closed-can, Open-tray 43 ± 1, 14.7 ± 0.1 B Uranium tailing Closed-can, Open-tray 313 ± 98, 102 ± 49 a: Measured using a calibrated RAD7 device.
91
4.3 Results and Discussion
4.3.1 The distribution of 222Rn in the charcoal bed
4.3.1.1 The uniformity of 222Rn in the charcoal bed
As mentioned in Section 4.2.1, after the exposure different 5 mm depth sections of the
total 30 mm (30 g) charcoal bed were removed from the canister. Each section was then
separately counted under the same source-detector geometry. The distribution of 222Rn in the
charcoal bed obtained is shown in Figure 4.1. The results were for five different exposure times
covering the range from one to 14 days. All results have been plotted relative to the value in the
section closest to the 222Rn source, which was furthest from the canister base as they have been
exposed in an inverted position. This experiment was performed using four sources of 222Rn
covering activity flux ranging from nearly 45 to 4500 mBq m−2 s−1. The values in Figure 4.1 were
the average for all sources. However, the level of the activity flux within our experimental range
appeared not to affect the relative distribution of 222Rn in the charcoal bed.
The distribution of 222Rn in the charcoal bed appeared less uniform at short exposure
periods. For one day of exposure, the closest layer to the canister base had 27 ± 1 % of the
amount of 222Rn that was in the layer furthest from the canister base. This difference became
smaller with the increase of exposure time. Most likely, as the exposure time increased more
222Rn nuclei were infused gradually into the deeper layers of the charcoal bed. Our findings
were supported by Uroševic et al. (1999) who theoretically simulated the time dependency of
the adsorbed 222Rn activity in the thickness of the charcoal layer. They found that the adsorbed
222Rn activity in the charcoal approached equilibrium as time proceeded.
The detection efficiency of the gamma detector might have been affected by the
distribution of 222Rn in the charcoal bed, especially for short exposure times, where the degree
of uniformity was low. This is because the geometry of 222Rn in the measured canister may not
be similar to that of the standard canister, which has a uniform distribution of 222Rn across the
charcoal bed. For one day of exposure, most of the 222Rn accumulated in the surface layer that
faced the source (Figure 4.1). However, after the exposure and during the counting period this
layer became the furthest from the detector and was therefore counted with less efficiency.
92
Figure 4.1 The distribution of 222Rn in the charcoal bed for different exposure times. The correction factors corresponding to the exposure time are shown in parentheses.
4.3.1.2 Detector efficiency at different distance
The effect of the low degree of uniformity of 222Rn in the charcoal bed on the detection
efficiency required investigation to obtain a realistic value of the activity flux. For this purpose,
detector efficiency at different heights was first checked using two certified point sources of
133Ba and 137Cs in the activity proportion of 0.64:1.00 housed in an empty canister. This
combination gave a spectrum that was comparable to our canisters within our region of interest
(143–740 keV).
Detector efficiency, as a function of distance from the canister base and the data of
Figure 4.1, had been used to work out a correction factor between a uniform distribution of
222Rn in the calibration standard and that which actually occurs due to 222Rn infusion during
exposure. The correction factor was obtained by the differences in the counting geometries
between the uniformly distributed standards and measured canisters. This correction factor is
shown in the legend of Figure 4.1. Its numerical value was 1.14 and 1.05 for exposure periods of
one day and three days, respectively, and then it levelled off to1.02 for 14 days.
93
Interestingly, the phenomenon of the reduction in detection efficiency as a function of
sample height above the detector and the adsorption of 222Rn in the charcoal bed tended to
counterbalance each other, at least for the setup we used. Consequently, the correction factor
that was due to the difference between the 222Rn distribution in the standard and exposed
canister for different days became smaller. The highest value of the correction factor was for the
first day of exposure due to the lower degree of uniformity of the 222Rn distribution in the
charcoal bed in this period as indicated in Figure 4.1. A correction factor of this magnitude is
likely to be well within other sources of uncertainty in the 222Rn measurement.
The numerical values of the activity flux given by Equation (1) should be divided by this
correction factor to obtain the expected value. Care should be exercised to generalise the use of
the correction factor given in this report because its value may change with the detector
characteristics, shape of the canister, and the properties of the activated charcoal material.
4.3.1.3 Canisters with different depth of activated charcoal
A number of charcoal canisters are simultaneously deployed during most applications of
activity flux measurements. Some differences in the material quantity may exist between these
cups. Differing amounts of activated charcoal in a canister corresponds to a different geometry,
which may require a correction factor. On the basis of the results presented in Figure 4.1, this
factor is expected to be negligible when the difference in the amount is small. The following
experiment was carried out to observe the effect of a dramatic difference of charcoal amount in
the cups.
A set of canisters were filled with five, 10, 15, 20, 25, and 30 g of activated charcoal
corresponding to bed depths of 0.5, 1.0, 1.5, 2.0, 2.5, and 3.0 cm, respectively. Five such sets
were prepared and exposed to different 222Rn sources as described in Table 4.1. This
experiment was, however, limited to one day exposure time which, as shown in Figure 4.1, gave
the most dramatic differences in the adsorbed 222Rn across the charcoal layers. The results
given in Table 4.2 were normalised to 2.5 cm charcoal bed depth because this is the depth we
normally use. Variation with respect to charcoal amount was significant, but little variations
were observed with respect to the 222Rn source strength within the investigated range of the
activity flux (45 - 4500 mBq m−2 s−1).
In Table 4.2, the charcoal amount correction factor is a value by which the activity flux
given in Equation (1) should be divided if charcoal depth is different from the calibration
reference (2.5 cm in our case). This correction factor was for one day exposure in the field. In
light of the observations given in Figure 4.1, the correction factors for large exposure times are
expected to be lesser in magnitude. As shown in Table 2, the amount of the adsorbed 222Rn in
the 0.5 cm layer was double that of the 2.5 cm layer. This indicates that the thin layer of charcoal
94
bed is counted with higher detection efficiency. Scarpitta (1996) investigated the dependence of
222Rn adsorption and desorption as a function of the depth of an activated charcoal bed.
Scarpitta (1996) reported that the thin layer of charcoal bed has higher 222Rn adsorption
coefficient. However in our case, the difference in the counting geometry as a function of
charcoal depth might also have played a role.
Table 4.2 Correction factors of 222Rn activity flux for different charcoal depths in a canister with respect to that of 2.5 cm depth for one day exposure time.
Activated charcoal mass Charcoal bed depth (x) Correction factor (g) (cm) 5 0.5 2.1 ± 0.3 10 1.0 1.6 ± 0.2 15 1.5 1.4 ± 0.2 20 2.0 1.2 ± 0.1 25 2.5 1.0 30 3.0 0.9 ± 0.2
4.3.2 The influence of exposure time on the measured 222Rn activity flux
The activity flux of 222Rn over the open-tray of the material was measured for different
exposure times using two independent 222Rn measurement systems: RAD7 detector and
activated charcoal canisters filled to 2.5 cm depth and exposed for one, two, three, five, seven,
10, and 14 days. The results of this experiment were plotted in Figure 4.2. These results were
corrected for the uniformity distribution of 222Rn in the charcoal bed using the correction
factors given in Figure 4.1. For the purpose of comparison, the measured values of the 222Rn
activity flux have been normalised to the value obtained at one day of exposure. The measured
activity flux using activated charcoal canisters appeared to decrease as the exposure time
increased. The value after 14 days of exposure was about half of that obtained after one day. The
relationship between the measured activity flux and the exposure time was
Jn /J1= (1.00 ± 0.03). e− (0.05 ± 0.01) n−1 where n≥1
where J1 and Jn represented the 222Rn activity flux after one day and n days of exposure time,
respectively. The data fit was obtained with R2 = 0.88.
95
Figure 4.2 Normalised 222Rn activity flux in open-trays plotted against exposure time.
Through a broad comparison with the results of RAD7, the measured 222Rn activity flux
appeared to agree with those obtained using activated charcoal canisters for exposure periods
between one to seven days, given an average of around four days as may be seen in Figure 4.2.
There might be multiple factors involved in this situation, causing the difference
between the two methods. The measurements of the 222Rn activity flux were not carried out
simultaneously with the charcoal canisters and the RAD7 detector. Although the measurements
were conducted under laboratory conditions, temperature and humidity may have influenced
the results for the two measured systems. The effect of temperature and humidity on the 222Rn
activity flux measurement methods had been reported by a number of authors (Iimoto et al.,
2008; Jha et al., 2000; Lawrence et al., 2009).
Interestingly, the reduction in the activity flux with respect to canister exposure time
was less when the samples were held in an airtight can (closed-can method as described in
Section 4.2.2) Figure 4.3. The values of the 222Rn activity flux that were measured after 14 days
of exposure remained at 80% or more of those obtained after one day of exposure. In Figure
4.3, the 222Rn activity flux that was measured with activated charcoal canisters was normalised
against the values obtained after one day of exposure. The influence of the exposure time on the
96
measured activity flux was small, as statistically indicated by the nearly flat slopes of the linear
fit of the activity flux versus exposure time in Figure 4.3. The observed values demonstrated
that not all decreases in the measured activity flux with exposure time were statistically
significant. Comparing the results with the values obtained by the RAD7, the values of the
activated charcoal agree within 25 % of those from RAD7 throughout the exposure time Figure
4.4.
Figure 4.3 The measured activity flux versus exposure time for charcoal canister fitted over cans containing samples.
97
Figure 4.4 The measured activity flux of the samples in the closed-can relative to RAD7 values.
To explore the influence of exposure time on the measurement of the 222Rn activity flux,
count rates were plotted versus the exposure time for the closed-can and open-tray samples in
Figure 4.5. Assuming continued and constant 222Rn activity fluxes from the sample and no loss of
222Rn by other processes than radioactive decay results in a trend of the normalised count rate
in the charcoal canisters shown by the dashed line in Figure 4.5 and described by equation (4.3).
𝑅𝑛
𝑅1=
(1 − 𝑒−𝜆.𝑡𝑒𝑛)
(1 − 𝑒−𝜆.𝑡𝑒1) (4.3)
The observed values in Figure 4.5 tended to level off well below this trend, particularly
for the open-tray samples. For the closed-can, it may be so that the 222Rn activity concentration
in the air gap between the sample and the charcoal could have kept on building; therefore the
increase in the count rates are closer to the expected values. On the other hand, the build-up of
222Rn initially continued in the same fashion under the canister in the open-trays. However,
after three days or so of exposure, the build-up did not occur as expected; indicating the loss of
222Rn from the canister due to a process other than radioactive decay.
We suggested that the divergence between the open-tray and closed-can behaviours
was related to the differences in the design of these two methods. Placing the canister directly
98
to the sample surface did not isolate it completely from the ambient environment because the
horizontal extent of the sample material underneath the canister was in direct contact with the
atmosphere. Therefore, a back diffusion of 222Rn from the canisters into the soil gas below is a
possibility. However, the 222Rn concentration in the charcoal canister grew at a lower than
expected rate, hence only a portion of 222Rn might have diffused back into the soil gas
underneath and possibly into the air.
The possibility of such back diffusion was also proposed by Mayya (2004) who analysed
the effect of the presence of an inverted cup on the diffusion of 222Rn from the soil underneath in
a two-dimensional framework. Mayya’s study demonstrated that the use of an inverted cup can
underestimate the measured activity flux due to the back diffusion phenomena that occurs at
the soil-cup interface. According to Mayya (2004), the underestimation of the activity flux
would require a correction factor that depends on the time period of the measurement,
dimensions of the cup and soil properties. The effect of back diffusion on the measured 222Rn
activity flux had also been mentioned in a number of other radon detectors, such as RAD7
(Tuccimei et al., 2006), Lucas cell (Aldenkamp et al., 1992) and alpha track detectors (Faheem,
2008).
It should be noted that the presence of activated charcoal in the canisters may mitigate,
but not eliminate the effect of back diffusion due to the ability of the charcoal to adsorb 222Rn
gas. In this case, another factor, adsorption of moisture by activated charcoal, might have also
contributed to the observed reduction in the measured 222Rn activity flux with exposure time in
the open-tray method. The adsorption of moisture by activated charcoal is known to reduce
222Rn adsorption in the charcoal (Ronca-Battista and Gray, 1988; Vargas and Ortega, 2007).
Water molecules might have been picked up from the ambient air through the sample material
during the exposure. This could explain the slight reduction in the measured activity flux in the
closed-can set. This is because enclosing the sample with charcoal canister in an airtight
container can prevent water molecules from the ambient air getting inside the container during
the exposure period. Therefore, the adsorption of 222Rn in the charcoal bed remained close to
the expected value as shown in Figure 4.5.
99
Figure 4.5 Relative count rate in the charcoal canisters versus the exposure time of samples taken in a closed-can or open-tray. The dashed line represents a situation referred by Equation 4.3.
4.4 Conclusion
The method of measuring 222Rn activity flux using activated charcoal canisters was
revisited through lab measurements using samples with activity flux ranging between 45 and
4500 mBq m-2 s-1. Two aspects were studied: (a) the distribution of 222Rn in the charcoal bed
and (b) the influence of the exposure time on the measured activity flux. The distribution of
222Rn in the charcoal bed was found to become more uniform with longer exposure times. A
correction factor was introduced by the differences in the counting geometries between the
uniformly distributed standards and measured canisters with non-uniform distribution. Its
value for one day of exposure, which recorded the lowest degree of uniformity, was small at
14 % or less. This was because the lower uniformity due to the diffusion of 222Rn in the charcoal
bed and the detection efficiency as a function of the charcoal depth appeared to counterbalance
each other. The values of the correction factors in this study may vary somewhat depending on
the type of detection system used and the shape of the charcoal canister.
100
The influence of the exposure time on the measured 222Rn activity flux using activated
charcoal canisters was observed in two different sets of experiments: (a) canisters introduced
to an open bed of the material and (b) sealed over a jar containing the material. The measured
activity flux decreased as the exposure time increased from one to 14 days. The reduction was
significant in the open-bed samples, while it was slight in canisters sealed over the jars
containing the same material. The reduction in the measured activity flux with exposure time
may be attributed to the back diffusion of 222Rn from the canister through canister-soil interface
and possibly to the adsorption of water molecules by the activated charcoal. The measurement
of the activity flux over the time period between one to seven days appeared to agree with the
values obtained using the RAD7 detector.
The activated charcoal canister is a useful method for routine measurement of 222Rn
activity flux. The statistical uncertainty due to counting is not a limiting factor in obtaining
results. Even for our lowest strength sample (about 45 mBq m-2 s-1) and for exposure time as
short as 1 day, one standard deviation uncertainty due to statistics of counting has been 10%.
However, limitations of the methodology are due to other factors, some of which have been
discussed in this paper and summarised above.
Acknowledgements
Sami Alharbi is grateful for the financial support provided by King Saud bin Abdulaziz
University for Health Sciences during PhD study. Many thanks go to Kazuyuki Hosokawa for
laboratory technical support.
101
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103
Chapter 5 Radon and thoron concentrations in public
workplaces in Brisbane, Australia
Sami H. Alharbi*, Riaz A. Akber
*School of Physical and Chemical Sciences, Queensland University of Technology, 2 George
Street, Brisbane, Qld 4000, Australia
Journal of Environmental Radioactivity 144 (2015), 69-76.
http://dx.doi.org/10.1016/j.jenvrad.2015.03.008
104
Statement of Contribution of Co-Authors for Thesis by Published Paper The authors listed below have certified that:
1. they meet the criteria for authorship in that they have participated in the conception,
execution, or interpretation, of at least that part of the publication in their field of
expertise;
2. they take public responsibility for their part of the publication, except for the
responsible author who accepts overall responsibility for the publication;
3. there are no other authors of the publication according to these criteria;
4. potential conflicts of interest have been disclosed to (a) granting bodies, (b) the editor
or publisher of journals or other publications, and (c) the head of the responsible
academic unit, and
5. they agree to the use of the publication in the student’s thesis and its publication on the
Australasian Research Online database consistent with any limitations set by publisher
requirements.
In the case of this chapter:
Radon and thoron concentrations in public workplaces in Brisbane, Australia
Contributor Statement of contribution
Sami Alharbi
Signature:
Date: 31/03/2016
Identified concept for the research paper,
conducted experimental work, analysed and interpreted the data,
wrote the manuscript
Riaz Akber Assisted with editorial process, aided in concept development, and
reviewed the manuscript.
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I have sighted email or other correspondence from all Co-authors confirming their certifying
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105
Abstract
Radon and thoron are radioactive gases that can emanate from soil and building
materials, and it can accumulate in indoor environments. The concentrations of radon and
thoron in the air from various workplace categories in Brisbane, Australia were measured using
an active method. The average radon and thoron concentrations for all workplace categories
were 10.5 ± 11.3 and 8.2 ± 1.4 Bq m-3, respectively. The highest radon concentration was
detected in a confined area, 86.6 ± 6.0 Bq m-3, while the maximum thoron level was found in a
storage room, 78.1 ± 14.0 Bq m-3. At each site, the concentrations of radon and thoron were
measured at two heights, 5 cm and 120 cm above the floor. The effect of the measurement
heights on the concentration level was significant in the case of thoron. The monitoring of
radon and thoron concentrations showed a lower radon concentration during work hours than
at other times of the day. This can be attributed to the ventilation systems, including the air
conditioner and natural ventilation, which normally operate during work hours. The diurnal
variation was less observed in the case of thoron, as the change in its concentration during and
after the working hours was insignificant. The study also investigated the influence of the floor
level and flooring type on indoor radon and thoron concentrations. The elevated levels of radon
and thoron were largely found in basements and ground floor levels and in rooms with concrete
flooring.
Keywords: Radon; Thoron; Workplaces; Indoor
5.1 Introduction
Radon (222Rn) and thoron (220Rn) are radioactive gases formed by the decay of radium
isotopes within the 238U and 232Th series. Radon and thoron emanate from the soil, rocks and
construction materials, and tend to accumulate in enclosed spaces, such as buildings, tunnels
and underground mines. The inhalation of radon and thoron gases, as well as their progenies, is
recognised as a potential cause of lung cancer (WHO, 2009) and considered to be the major
contributor to public exposure from natural sources (UNSCEAR, 2000).
Indoor radon and thoron are of particular concern due to the possibility of their
concentrations being elevated to a considerable level. In some environments, indoor radon was
found to be greater than outdoors by a factor of up to 50 (Porstendorfer et al., 1978; Song et al.,
2005). In general, in cities such as Brisbane, Australia, people tend to spend more time indoors,
whether in dwellings or workplaces, increasing their exposure to indoor sources. A
considerable amount of time is often spent at the workplace, and exposure due to radon and
thoron in some workplaces can be significant (Gillmore et al., 2000). Fortunately, exposure due
to radon and thoron in indoor environments is feasible to control (IAEA, 2003; ICRP, 1993,
2014; UNSCEAR, 2006).
106
Radon and thoron concentrations can be measured using passive or active methods
(Alharbi and Akber, 2014; IAEA, 2013; Laiolo et al., 2012). Passive detectors integrate the
average radon and thoron concentrations over one season or longer while active detectors
provide continuous radon and thoron monitoring during the day. However, workplaces are only
occupied for a certain time period on specific days of the week; thus, the average concentration
over the working hours may differ from that obtained over the total time. In this case, using
passive detectors may not be appropriate, as most of these detectors are incapable of
monitoring radon and thoron concentrations selectively during work hours.
Generally, the levels of indoor radon and thoron vary widely in time and space
(Bochicchio et al., 2005; Faheem and Mati, 2007; Prasad et al., 2012; Singh et al., 2005), being
higher in the winter and lower in the summer. This seasonal pattern is mainly attributed to
ventilation systems, including air conditioning and heating systems, which play essential roles
in changing radon levels in indoor environments (Lee and Yu, 2000; Marley and Phillips, 2001).
A correction factor for such variations has been proposed (Baysson et al., 2003; Faheem and
Mati, 2007; Pinel et al., 1995); alternatively, the measurement may be taken only during the
coldest months, as recommended by some authorities (IAEA, 2003; NORDIC, 2000). The
seasonal trend of indoor radon, however, has been less observed in workplaces, and some have
exhibited no seasonal patterns (Inoue et al., 2013; Papachristodoulou et al., 2010; Tokonami et
al., 1996). Instead, the diurnal variation has been found more often in workplaces (Vaupotič et
al., 2012; Zahorowski et al., 1998).
Another factor that can affect the level of indoor radon is the underlying geology. A
significant correlation has been found between the level of indoor radon and the geological
composition of the ground (Moreno et al., 2008; Sundal et al., 2004). Radon in the soil gas can
enter buildings through cracks in the floor by way of a pressure-driven mechanism. The
relatively long half-life of radon (3.8 d) gives it enough time to travel, which can be clearly seen
in multi-story buildings where radon levels were found to be higher in the basement and on the
ground level (Papachristodoulou et al., 2010; Shaikh et al., 2003). The influence of the
underlying geology is less pronounced in indoor thoron, possibly due to its short half-life (56
seconds). Instead, thoron is more influenced by the distance from the source, which includes the
walls and floors in buildings (Fujimoto et al., 1994; Urosevic et al., 2008).
There are other parameters that might influence indoor radon and thoron levels, but
with less significance. These include meteorological conditions (Miles, 2001), the buildings’
characteristics (Gunby et al., 1993) and the occupants' habits1. These factors can contribute in
different degrees to the radon and thoron variations in indoor environments.
1 Such as keeping opening windows and doors, introducing floor and wall covering using fans or air conditioning. etc.) —Note: footnotes were added to the thesis, not present in the published paper.
107
During the past few years, there has been an increasing awareness of the exposure of
radon and thoron in workplaces other than mines (IAEA, 2003, 2011; UNSCEAR, 2006; WHO,
2009). These workplaces are often located underground, such as in subways and tunnels, or in
aboveground places, such as offices, schools, shops and factories. Mines that are open to tourists
are also included in these areas. Workers and members of the public may be adventitiously
exposed to radon and thoron in these workplaces. Recently, the ICRP has recommended a
reference level in the range of 100 to 300 Bqm-3 for radon concentration in buildings without
distinguishing between dwellings, workplaces and mixed-use buildings (ICRP, 2014).
A broad spectrum of workplaces have been surveyed worldwide for indoor radon
concentrations (Inoue et al., 2013; Martín Sánchez et al., 2012; Solomon et al., 1996; Vaupotič et
al., 2012). In Spain, a wide range of workplaces was measured, including offices, shops, spas,
caves, museums, etc. (Martín Sánchez et al., 2012), and the average radon concentration
exceeded 400 Bq m-3 in 16% of the measured workplaces. In a study conducted in Greece by
Clouvas et al. (2007), the indoor radon concentrations in schools were found to be higher than
in other workplaces.
On the other hand, the available data on indoor thoron is largely limited, especially for
workplaces (Burghele and Cosma, 2012; Inoue et al., 2013; Vaupotič et al., 2012). This is
perhaps due to the measurement difficulties associated with its short half-life; additionally,
there is no desire to regulate indoor thoron in most countries. It is worth noting that the health
risks associated with the inhalation of thoron and its decay products are considered to be equal
if not greater than those of radon and its progeny (ICRP, 1993; Shang et al., 2005; UNSCEAR,
2000). Moreover, the concentration of thoron in some indoor spaces was found to be four times
greater than that of radon (Ramola et al., 2012; Yamada et al., 2006).
In Australia, a very limited number of workplaces, other than those for which the
exposure is controlled as a planned exposure situation, were surveyed for indoor radon
concentrations (Solomon et al., 1992; Solomon et al., 1996; Zahorowski et al., 1998). A radon
level of 6330 Bq m-3 was detected in Australian show caves, giving off a yearly effective dose of
up to 5.9 mSv for tour guides (Solomon et al., 1996). An action level of 200 and 1,000 Bq m-3 of
radon in dwellings and workplaces, respectively, was proposed by the Australian Radiation
Protection and Nuclear Safety Agency (ARPANSA) based on the Australian National Health and
Medical Research Council (NHMRC) recommendation, jointly with the National Occupational
Health and Safety Commission (NOHSC) (NHMRC, 1995).
Universities are educational premises where a wide range of workplaces can be found with
a high occupancy factor for members of the public. Radon and thoron levels in universities may
pose potential long-term health risks to the full-time staff and students who may extend their
work until late, to conduct research. This paper aims to assess the levels of indoor radon and
108
thoron across a wide spectrum of workplace categories at the Queensland University of
Technology.
5.2 Materials and Methods
5.2.1 Sampling site
The measurements of indoor radon and thoron were conducted at different sites at the
Queensland University of Technology (QUT) Gardens Point campus. The campus is located in
the central business district of Brisbane city, the state capital of Queensland, Australia, at a
latitude of 27°29ʹ South and a longitude of 153°2ʹ East (Figure 5.1). The campus consists of a
wide range of multi-story buildings designed for mixed-use purposes. These buildings are
constructed of concrete, brick, metal and cement, and they are occupied by students, staff
members and visitors during the daytime. Currently, the University has 4,349 staff members
including academic, technician and administrative staff. All buildings are air-conditioned to
adjust the temperature in the summer and winter during work hours.
The study area is underlain by basement rocks that consist mainly of meta-sedimentary
and metamorphic formation covered by pale, shallow and stony soil (Willmott et al., 2012).
Brisbane has a subtropical climate and the rain is distributed between the summer and autumn
months.
The sampling sites have been chosen to cover a broad range of workplace categories.
The examined workplaces in this study include offices, classrooms, research rooms,
laboratories, a library, main halls, confined areas, basements, storage areas and a car park.
These places are normally occupied by professional academic staff members and students.
109
5.2.2 Radon and thoron measurements
The study was conducted at the selected workplaces throughout the cooler months,
between June and November 2013. All measurements were carried out during the weekdays
(Monday through Friday) using a solid-state alpha detector RAD7 designed by Durridge, USA.
The radon and thoron concentrations were measured on several floors of the multi-
story buildings. At each sampling site, the radon and thoron levels were monitored for two days
at the ground (5 cm) and at the breathing levels (120 cm). The reason for measuring the
concentrations at two heights was to compare the variations in radon and thoron levels with the
distance2. The radon and thoron concentrations were monitored simultaneously and
continually over the examination period, and the average radon and thoron concentrations over
the work hours, as well as over the whole day, were obtained.
Aside from the measurement of radon and thoron concentrations, the data has been
handled in such a way as to investigate various parameters that may influence the radon and 2 Distance: height above the floor surface—Note: footnotes were added to the thesis, not present in the published paper.
" Brisbane
Figure 5.1. A map of the sampling site.
110
thoron concentrations in indoor environments. The effect of the floor level on the concentration
of radon and thoron has been studied. The influence of flooring type on indoor radon and
thoron levels was investigated.
The detector used in this study, Rad7, has an internal cell of a 0.7-litre volume
hemisphere with a solid-state, ion-implanted, planar, silicon alpha detector installed in the
centre. The inner side of the internal cell is coated with an electrical conductor. A high voltage is
applied across the internal cell to create an electric field, which directs positive charges toward
the centre, where the detector sits. The sampled air is pumped through a 0.45 µm filter, which
excludes radon and thoron progeny, into the internal cell with a flow rate of 650 ml/min. In the
cell, radon and thoron decay into polonium isotopes by emitting detectable alpha particles,
which in turn are converted into electrical signals. A spectrum of alpha energy in the range of 0–
10 MeV is produced and grouped into eight windows (A – H). Windows A and C are used to
determine the 222Rn level by collecting alpha energies from 218Po (6.0 MeV) and 214Po (7.69
MeV), respectively. Windows B and D are used to determine the 220Rn through 216Po (6.78 MeV)
and 212Po (8.78 MeV) levels. The rest of the windows on the detector are not applicable to our
measurements. A drying tube containing desiccant (CaSo4) is used to maintain the relative
humidity of the RAD7 below 10%, as its sensitivity is degraded at high humidity levels.
The RAD7 was set to a 3-hour cycle time to obtain better statistics. Prior to each
measurement, the instrument was purged for 10 minutes to remove any remaining radon or
thoron gas from the internal cell. During the sampling period, the values of radon and thoron
are monitored and printed out every 3 hours. The data is then transferred from the RAD7 to the
computer using the Capture software supplied by the manufacturer. The software allows the
values to be corrected for the humidity and the spillover of the counts that occur between the
adjacent windows of the RAD7.
The instrument was calibrated by the manufacturer for radon, with an overall accuracy
of about 5%; however, thoron sensitivity is considered by the manufacturer to be half of the
radon sensitivity. For verification purposes, the RAD7 was also calibrated using the radon and
thoron flow through type sources (Pylon Model Rn-1025 and TH-1025, Pylon Electronics Inc.,
Canada). The radon and thoron concentrations from the Pylon source were predicted using a
formula provided by the manufacturer. The measured radon and thoron concentrations from
the Pylon sources with the RAD7 are within 10% of the calculated concentrations (Table 5.1).
As a quality control procedure, the measurements of radon and thoron concentrations in
three locations were repeated twice in different days using the same detector and setup. There
were no significant differences between the repeated measurements of radon and thoron
concentrations (p = 0.570 and p = 0.880; respectively) in the three locations.
111
Table 5.1 Summary of the instrument calibration check using certified sources.
Isotopes Flow rate (m3 min-1) Concentration (Bq m-3) Ratio (%)
Measured Calculated
222Rn 0.001 2343 ± 158 2158 ± 86 1.09 ± 0.11
220Rn 0.001 38800 ± 1100 40457 ± 3237 0.96 ± 0.10
5.3 Results and Discussion
5.3.1 Radon and thoron concentration measurements
The average radon and thoron concentrations from each sampling site are given in Table
5.2. The averages are worked out for 5 and 120 cm heights combined for the entire two days
period of measurements. The radon concentration ranged from 0.7 to 86.6 Bq m-3, with an
arithmetic mean of 10.5 ± 11.3 Bq m-3. The maximum radon level was detected in a confined
room on the ground floor, and the minimum level was found in an office on the third floor. A
comparison of radon concentrations across all sites for both the floor level and 120 cm above
the floor is shown in Figure 5.2. A one-way analysis of variance (ANOVA) revealed a significant
difference (p < 0.05) in radon concentrations across the workplaces. A post hoc multiple
comparison test indicated that the radon concentration in basements and in the confined room
were significantly higher than in other locations.
All the values for the radon concentrations were below the reference level (200 Bq m-3)
recommended by the Australian authority for such workplaces (NHMRC, 1995). The reported
radon concentration is comparable to the average indoor radon concentration in Queensland
homes, 6.3 ± 4.4 Bq m-3 (Langroo et al., 1991), and it is lower than the typical global indoor
radon level, 40 Bq m-3, (UNSCEAR, 2000).
The level of thoron varied between 0.6 and 78.1 Bqm-3, with an arithmetic mean of 8.2 ±
3.3 Bq m-3. This average is comparable to that of radon. However, a poor correlation (R2 =
0.015) has been observed between their concentrations, as shown in Figure 5.3. Thoron
concentration was significantly higher in the basements and storage rooms. Interestingly, the
lowest thoron concentration was observed in the confined room where the radon concentration
was the highest. It is worth noting that no ventilation system was operating in the confined
room at the time of conducting the measurement. This implies that the thoron concentration
build-up is less influenced by the ventilation system. In contrast, the radon gas reached its
highest value in the confined room, and places with poor ventilation systems are generally ideal
for elevating radon levels (Alghamdi and Aleissa, 2014).
112
Table 5.2 Radon and thoron concentrations in different locations.
Workplace Number of Building Floor Radon Concentration (Bq m-3) Thoron concentration (Bq m-3) Category workplaces type* level Range Mean ± SD Range Mean ± SD
Office 7 A, B, C multiple 0.7 – 28.1 6.3 ± 3.2 2.2 – 25.5 6.2 ± 3.5 Classroom 3 B, C 3 1.7 – 23.0 7.3 ± 5.2 2.2 – 21.6 3.8 ± 1.1 Laboratory 6 A, B, C multiple 1.2 – 41.7 9.3 ± 7.1 2.2 – 15.2 4.4 ± 1.9 Library 3 C 2, 3, 5 1.1 – 10.2 3.8 ± 1.3 2.6 – 11.5 4.6 ± 1.3 Basement 2 B, C basement 3.8 – 54.9 26.5 ± 14.6 7.9 – 46.8 23.7 ± 3.9 Research room 2 C 2, 5 0.9 – 22.0 8.9 ± 3.4 2.3 – 14.0 6.1 ± 1.5 Storage 3 B, C 1, 4 1.7 – 35.6 11.1 ± 5.9 2.8 – 78.1 23.9 ± 23.9 Confined area 1 B ground 29.3 – 86.6 64.6 ± 6.0 0.6 – 8.4 2.2 ± 1.5 Car park 1 C ground 0.9 – 13.2 6.3 ± 2.7 3.0 – 19.6 10.4 ± 3.9 Main hall 1 B ground 20.8 – 63.1 35.8 ± 5.1 2.6 – 11.6 4.3 ± 2.2 All 29 0.7 – 86.6 10.5 ± 11.3 0.6 – 78.1 8.2 ± 1.4 * A : Metal cladded B : Brick veneers C : Prefabricated concrete
113
Workplace category
Office
Classroom
Library
Laborato
ry
Basement
Research R
oom
Stora
ges
Ra
do
n c
once
ntr
atio
n (
Bq
m-3
)
0
10
20
30
40
Workplace category
Office
Classroom
Library
Laborato
ry
Basement
Research R
oom
Stora
ges
Tho
ron c
once
ntr
atio
n (
Bq
m-3
)
0
20
40
60
Figure 5.2 A box plot showing radon and thoron concentrations across workplaces. The boxes represent 25–75% of the values with median concentrations represented by the lines inside the boxes. Whiskers represent the range of the distribution and the dots outside the range of 95% indicate values as outliers.
y = 0.12x + 7.02
R² = 0.015
Radon concentration (Bq m-3
)
0 20 40 60 80
Tho
ron c
once
ntr
atio
n (
Bq
m-3
)
0
20
40
60
80
Figure 5.3 Correlation between radon and thoron concentrations
The levels of radon and thoron in the present study were relatively low when compared
to other international reported values (Clouvas et al., 2007; Martín Sánchez et al., 2012).
Although the concentration of indoor radon is generally low in Australia (Langroo et al., 1991),
the ventilation system used in the workplaces is a possible cause of lowering radon levels, as
most examined workplaces are air-conditioned. The operation of the air conditioner can
increase the air exchange between the indoor and outdoor air through the pressure driven and
hence reduce radon levels (Lee and Yu, 2000). The location of Brisbane can be another reason,
as coastal sites have lower radon concentrations than inland sites (Arnold et al., 2009; Crawford
et al., 2011). This is because of the poor radon concentrations in the air coming from the ocean.
Additionally, the radium content, the progenitor of radon, in Australian building materials was
found to be lower than in many other places of the world (Beretka and Mathew, 1985). These
114
factors might have contributed in varying degrees to the overall radon and thoron levels in the
workplaces investigated in this study.
Among all the examined sites, the basement showed a significant increase (p < 0.05) in
both radon and thoron concentrations when compared to other locations (Figure 5.2). The
median radon and thoron concentrations in the basements were 34.5 and 22.1 Bq m-3,
respectively. Due to the close proximity of basements to the ground underneath, which
represents the main source of indoor radon, radon in the soil gas can easily enter through
cracks in the floor of the basement. Furthermore, the floors of basements are made of concrete
and generally not covered by carpets or any other materials that can disturb radon flux from the
cracks. Another possible reason is the lower air exchange in the basements, where access is
limited and the ventilation is relatively poor. Additionally, the walls in all the measured
basements were not coated with paint, which is known to reduce the exhalation rate of radon
(O'Brien et al., 1995) as well as of thoron (Wang et al., 2012). The latter is likely the main cause
of higher thoron levels in the basements, as walls and floors represent a major source of indoor
thoron.
Ln (Radon concentration)
1 2 3 4
fre
que
ncy
0
4
8
12
16
Ln (Thoron concentration)
1 2 3 4
fre
que
ncy
0
5
10
15
20
25
Figure 5.4 Histogram showing the distribution of the natural logarithms of radon and thoron concentrations in workplaces with a normal distribution fitted to the data. Geometric means of radon and thoron concentrations were 7.4 and 5.8 Bq m-3, with a geometric standard deviation of 2.2 and 2.0 Bq m-3, respectively.
The natural logarithms of the overall radon and thoron concentrations in workplaces
appeared to be normally distributed (Figure 5.4). In other words, indoor radon and thoron
concentrations followed a log-normal distribution, which perhaps indicates there are several
random variables influencing indoor radon and thoron levels in a multiplicative rather than
additive manner. The log-normal distribution of radon concentrations has been widely found in
indoor environments whether workplaces (Oikawa et al., 2006; Papachristodoulou et al., 2010)
or dwellings (Langroo et al., 1991).
115
5.3.2 Diurnal variation
Time of the day
0.00 3.00 6.00 9.00 12.00 15.00 18.00 21.00 24.00
No
rma
lise
d r
ad
on c
once
ntr
atio
n
0.0
0.5
1.0
1.5
2.0
2.5
Time of the day
0.00 3.00 6.00 9.00 12.00 15.00 18.00 21.00 24.00
No
rma
lise
d tho
ron c
once
ntr
atio
n
0.0
0.5
1.0
1.5
2.0
Figure 5.5 The variation in radon and thoron concentrations during the day.
The diurnal behaviour of radon and thoron concentrations in workplaces is illustrated
in Figure 5.5. The concentrations of radon and thoron in all workplaces were averaged on a
three-hourly basis. During normal work hours, between 09:00 am and 05:00 pm, the
concentration of radon was reduced to its lower value and remained relatively low over the
entire period. This concentration tended to rise after 05:00 pm, until it reached its maximum
value in the early morning. In fact, most workplaces are air-conditioned, which is usually
operated during the occupied hours; the air-conditioner acts as a ventilation system, and its
influence on reducing indoor radon concentrations is well-known (Marley and Phillips, 2001).
The frequent opening of doors and windows during the occupied period may also contribute in
lowering the level of indoor radon by increasing the air exchange in the room (Doi et al., 1994;
Yu et al., 2000). Reductions in the indoor radon concentration during the working hours have
been observed in a number of workplaces (Lee et al., 2013; Vaupotič et al., 2012).
Radon concentrations during and after the working hours were compared with the
average over the total time (Figure 5.6). The mean radon concentration for all workplaces
during the occupied hours was 9.6 ± 11.6, while it was 10.5 ± 11.3 Bq m-3 over the total time.
Although the concentration during the working hours was less than the average over the whole
time at most sites, no statistical significance was observed between the periods. However, a
comparison between the radon concentrations during and after working hours showed a
significant difference (p < 0.05) in the radon level between the two periods. Among workplaces,
radon concentrations in offices during the occupied time were statistically lower (p < 0.05) than
after-hours concentrations.
In some locations, such as the library and car park, the concentrations of radon during
and after work hours were equal to the average over the total time. This is possibly due to the
type of ventilation system used in those places. In the library, work hours start in the early
116
morning (7:00 am) and extend until late (around 10:00 pm); consequently, the air conditioners
are switched off only for a few hours, whereas the car park is ventilated naturally all the time.
Workplace category
Office
Classroom
Library
Laborato
ry
Research R
oom
Basement
Stora
ges
Confind a
rea
Car park
Main h
all
No
rma
lise
d r
ad
on c
once
ntr
atio
n (
Bq
m-3
)
0.0
0.5
1.0
1.5
2.0
Radon-working hours
Radon-off hours
Average radon per day
Workplace category
Office
Classroom
Library
Laborato
ry
Research R
oom
Basement
Stora
ges
Confind a
rea
Car park
Main h
all
No
rma
lise
d tho
ron c
once
ntr
atio
n (
Bq
m-3
)
0.0
0.5
1.0
1.5
2.0
Thoron-working hours
Thoron-off hours
Average thoron per day
Figure 5.6 Normalised radon and thoron concentrations during and after work hours.
The findings suggest that attention should be paid to radon concentrations during work
hours, which represent the actual exposure time, when radon levels are measured in
workplaces. As the mean radon concentration during the working hours was less than over the
total time, the average radon concentration can be estimated incorrectly. Therefore, averaging
the radon concentration in the workplace without considering the concentration during the
actual exposure time (i.e. working hours) can result in overestimating radon concentrations
and, subsequently, the effective dose.
In contrast, the diurnal variation in the thoron concentration was less observed in all
workplaces (Figure 5.6). The mean thoron concentrations during the working hours and over
the total time were 8.2 ± 10.9 and 8.2 ± 1.4 Bq m-3, respectively. The statistical analysis
indicated no difference (p = 0.984) between the thoron concentrations during and after work
hours, nor over the total time.
117
5.3.3 The influence of the measurement heights
Figure 5.7 Radon and thoron concentrations in the air at 5 cm and 120 cm above the floor for each sampling site.
Radon Thoron
Co
nce
ntr
atio
n (
Bq
m-3
)
0
10
20
30
40
50
60
70
Ground Zone
Breathin Zone
Figure 5.8 Box-whisker plot showing the overall radon and thoron concentrations at the ground and breathing zones for all location.
At each site, the radon and thoron concentrations were measured at two different
heights: 5 cm (ground zone) and 120 cm above the floor (breathing zone) (Figure 5.7). In some
sampling sites, the radon and thoron concentrations were measured at one height only due to
118
access constrains where missing concentrations in Figure 5.7. A comparison between the radon
and thoron concentrations at two different heights for all locations is shown in Figure 5.8, which
indicates the average radon concentrations near the surface of the floor and at the breathing
zone were nearly consistent at all sites. The statistical analysis revealed no significant
differences (p = 230)* between the radon concentrations of the two zones. Conversely, the level
of thoron close to the ground was significantly higher (p < 0.05) than its value at breathing
height, indicating the influence of distance on the thoron concentration. As thoron has a
relatively short half-life compared to radon, most of it decays before reaching the breathing
zone. Care, however, should be taken to distinguish between thoron and thoron progeny. Long-
lived thoron progeny, 212Pb (T1/2 = 10.6 hours), may distribute differently than thoron.
5.3.4 The influence of floor levels
Floor level
Basement Ground First Second Third and above
Ra
do
n c
once
ntr
atio
n (
Bq
m-3
)
0
10
20
30
40
50
60
70
Floor level
Basement Ground First Second Third and above
Tho
ron c
once
ntr
atio
n (
Bq
m-3
)
0
10
20
30
40
50
60
70
Figure 5.9 Box-whisker plot showing radon and thoron concentrations categorised by the floor-level.
As the measurements in this study were conducted in multi-story buildings, the
influence of the floor level on the concentration of radon and thoron was investigated. Figure
5.9 shows the radon and thoron concentrations characterised by the floor level, in that the level
of radon decreased as it moved towards the upper floors, which is not the case for thoron in our
data. The basements and ground floors had significantly (p < 0.05) higher radon concentrations
than other floor levels, with no significant difference between the basements and ground floors
(p = 0.383). In the case of thoron, only the basements were statistically different (p < 0.05) from
other floors.
* Typographic error. The value should be 0.230—Note: footnotes were added to the thesis, not present in the published paper.
119
5.3.5 The role of flooring on indoor radon and thoron concentrations
A)
Flooring type
Concrete Tile Vinyl Carpet
Ra
do
n c
once
ntr
atio
n (
Bq
m-3
)
0
10
20
30
40
50
60
70B)
Flooring type
Concrete Tile Vinyl Carpet
Tho
ron c
once
ntr
atio
n (
Bq
m-3
)
0
10
20
30
40
50
60
70
Figure 5.10. Box-whisker plot showing A) radon and B) thoron concentrations characterised by the
flooring type.
The effect of flooring on the levels of indoor radon and thoron was examined in this
study by categorising the data according to the flooring type, as can be seen in Figure 5.10.
There were four types of flooring: carpet, tile, vinyl sheet and concrete. A one-way ANOVA on
the log-transformed data showed a significant variation in radon concentration with the floor
type (p < 0.05). A post hoc test, Fisher's least significant difference (LSD), identified that the
radon concentrations in rooms with concrete floors were higher than in rooms with other
flooring types. The level of thoron was also higher in concrete rooms, indicating the effect of the
flooring type on the indoor thoron concentration. This is due to the existence of uranium and
thorium contents in concrete (Beretka and Mathew, 1985). These results suggest that covering
the floor with carpet or vinyl sheets can help mitigate indoor radon and thoron levels.
5.4 Conclusion
The concentrations of radon and thoron in various workplace categories have been
measured at Queensland University of Technology, Brisbane, Australia. The average radon and
thoron concentrations for all workplace categories were 10.5 ± 11.3 and 8.2 ± 1.4 Bq m-3,
respectively. Among workplaces, the concentration of radon in basements and confined areas
was significantly higher than in other locations, while the highest thoron level was found in
storage areas. Poor ventilation is the most likely cause of increasing the radon level in those
locations. During the working hours, radon concentration tended to be lower than at other
times of the day. This can be attributed to the ventilation systems, including air conditioners
and natural ventilation, which normally operate during working time. The diurnal variation was
less observed in the case of thoron as the change in its concentration during and after the
working hours was insignificant. The findings suggest that care should be taken when
120
measuring radon concentrations in workplaces, as its concentration during the working hours
may considerably differ from that of the entire day.
The influence of other parameters on indoor radon and thoron concentrations was
investigated from different point of view. At each site, the concentrations of radon and thoron
were measured at two heights, 5 cm and 120 cm above the floor. The level of radon showed no
difference between the two heights, while the change in thoron concentration was significant,
indicating the impact of the distance from the source on the thoron concentration. The floor
level and flooring type affected the level of indoor radon and thoron concentrations which are
elevated in basements and on the ground floor level, as well as in rooms with concrete flooring.
Acknowledgements
The authors are thankful for the support provided by Colin Phipps for arranging access to
the sampling sites.
121
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Chapter 6 A broad energy germanium detector for
routine and rapid analysis of naturally occurring
radioactive materials
Sami H. Alharbi*, Riaz A. Akber
*School of Physical and Chemical Sciences, Queensland University of Technology, 2 George
Street, Brisbane, Qld 4000, Australia
Journal :
Journal of Radioanalytical and Nuclear Chemistry (In Press)
125
Statement of Contribution of Co-Authors for Thesis by Published Paper The authors listed below have certified that:
1. they meet the criteria for authorship in that they have participated in the conception,
execution, or interpretation, of at least that part of the publication in their field of
expertise;
2. they take public responsibility for their part of the publication, except for the
responsible author who accepts overall responsibility for the publication;
3. there are no other authors of the publication according to these criteria;
4. potential conflicts of interest have been disclosed to (a) granting bodies, (b) the editor
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5. they agree to the use of the publication in the student’s thesis and its publication on the
Australasian Research Online database consistent with any limitations set by publisher
requirements.
In the case of this chapter:
A broad energy germanium detector for routine and rapid analysis of naturally occurring
radioactive materials
Contributor Statement of contribution
Sami Alharbi
Signature:
Date: 31/03/2016
Identified concept for the research paper, prepared samples of
analysis, conducted experimental work, analysed and interpreted the
data, wrote the manuscript
Riaz Akber Assisted with editorial process, aided in concept development
Helped in experimental design, reviewed the manuscript.
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126
Abstract
The analysis of naturally occurring radioactive materials using a broad energy germanium
detector is reported, along with a description of the rapid measurement of 226Ra and the
determination of 227Ac in environmental samples. The detector calibration procedure, sample
preparation, and the necessary correction factors have been described. The detector was found
useful for the routine analysis of natural radionuclides, with the ability to detect low photon
emission peaks at lower gamma energies. The results showed that the direct determination of
226Ra via the 186 keV peak is feasible and the determination of 227Ac using gamma spectrometry
is reliable.
Keywords: BEGe; NORM; 226Ra; 227Ac; Rapid measurement
6.1 Introduction
The information obtained from individual radionuclides of naturally occurring
radioactive material (NORM) has become a powerful tool for many scientific disciplines, such as,
radiometric dating (Ugur et al., 2003, Aba et al., 2014), atmospheric studies (Baskaran, 2011,
Doering and Akber, 2008), soil erosion and sediment deposition rates (Matisoff, 2014, Zapata,
2003). Selections of measurement methods and sample preparations are dictated by a project’s
specific needs, the required accuracy and precision of results, and by the availability of
resources. Analytical methods for NORM radioisotopes are generally time-and resource-
consuming—and, therefore, expensive. Gamma-ray spectrometry is one of the most effective
measurement techniques for identifying radionuclides and quantifying the activity levels of
materials in a non-destructive manner (Engelbrecht and Schwaiger, 2008). For example,
Vesterbacka et al. (2009) compared various methods of detecting radioisotopes by alpha, beta
and gamma spectrometry and found gamma spectrometry to be superior in terms of the speed
of analysis.
Over the past few decades, several types of gamma detectors have been developed for
various applications. The broad energy germanium (BEGe) detection system (manufactured by
Canberra Industries Inc.) represents the latest technology that is commercially available for
routine gamma analysis. According to the manufacturer, the BEGe detector has a resolution
equivalent to the conventional planar detector at lower gamma energies and comparable to the
coaxial detector at higher energies. Han and Choi (2010) tested different types of germanium
detectors. In their study, the BEGe detector had the highest efficiency at lower gamma energies
than other detectors. This may be beneficial for analysing environmental samples in which a
number of key radionuclides emit gamma-rays in low-energy regions.
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The BEGe detector has been extensively used in nuclear physics (e.g., in neutrino-less
double beta decay experiments (Aalseth et al., 2011, Budjáš et al., 2013)) and in medical science
(e.g., in lung counting measurements (Webb and Kramer, 2001)). However, the use of BEGe
detectors in environmental studies is still limited. This paper examines the potential
implications of BEGe detectors for the routine measurement of NORM. The detector was also
employed to investigate two other topics in the context of NORM: the rapid measurements of
226Ra and the determination of 227Ac.
6.1.1 Routine analysis of NORM
Routine measurement of NORM targets radionuclides in the uranium, actinium and
thorium decay series as well as potassium-40. Some of radionuclides in the natural series
cannot be measured directly with gamma spectrometry as they do not emit gamma-ray or their
emitted gamma-ray is very weak. In such cases, these radionuclides can be determined through
their daughter radionuclides assuming a state of equilibrium between the daughter and parent
radionuclides. For example, 238U emits a very weak gamma line (0.064%) at 49.55 keV, which
makes the direct measure of 238U through that line impracticable. Therefore, the activity of 238U
is typically determined through the gamma-rays of its daughter nuclides, such as, 234Th, 234mPa,
214Pb or 214Bi.
6.1.2 Rapid measurements of 226Ra
226Ra is an important natural radionuclide that has many applications in environmental
and dating studies (Voltaggio et al., 1995, Shuktomova and Rachkova, 2011, Condomines and
Rihs, 2006, Paytan et al., 1996). It has a long half-life of 1600 years, and it decays to 222Rn (half-
life of 3.8 days) by emitting alpha particles. This decay is accompanied with gamma-rays at
186.1 keV, with a low emission probability (3.6%) (Ekström and Firestone, 2004).
226Ra can be measured in environmental samples using gamma spectrometry (Murray et
al., 1987, Dowdall et al., 2004, Landsberger et al., 2013). In the gamma-ray spectrometry
technique, 226Ra is commonly determined through its progenies: 214Pb (at 295.2 and 351.9 keV)
and 214Bi (at 609.3, 1120.3 and 1764.5 keV) (Al-Ghamdi et al., 2016, Khandaker et al., 2012).
This is because the 186.1 keV peak of 226Ra interferes with the 185.7 keV (57.2%) peak of 235U,
the 186.1 keV (0.009%) peak of 230Th, and the 186.2 keV (1.76%) peak of 234Pa. The latter two
peaks are usually ignored due to the extremely low intensity of 230Th and the low branching
ratio of 234Pa. However, this procedure is time consuming because it requires a delay period of
at least three to four weeks to achieve secular equilibrium between 226Ra and its daughters,
214Pb and 214Bi.
In order to overcome the delay time, a number of methods for the rapid determination
of 226Ra have been attempted. Some researchers have adopted the method of deconvolution of
the overlapping peak at 186 keV (Saïdou et al., 2008, Zhang et al., 2009). However, this
128
approach is unlikely to provide accurate results due to the small distance between the
overlapped peaks as demonstrated by Dowdall et al. (2004). Another approach of determining
226Ra is by direct measurement of the186.1 keV peak after subtracting the contribution of 235U
(e.g. (De Corte et al., 2005, Justo et al., 2006, Dowdall et al., 2004, Johnston and Martin, 1997,
Völgyesi et al., 2014)). The contribution of 235U can be deduced through the natural activity ratio
of 235U/238U using 238U daughter, 234Th (Dowdall et al., 2004) or using an independent method
for measuring 235U concentration (Johnston and Martin, 1997), which would add extra cost to
the analysis. The main issues with the former approach are the long waiting time for secular
equilibrium of 234Th, which can be guaranteed after 240 days. Whereas the problem with the
latter approach is the alterations that could occur to the 235U/238U ratio as a result of certain
nuclear activities (Danesi et al., 2003, Yoshida et al., 2000, Sansone et al., 2001). A different
technique was developed by Johnston and Martin (1997) utilising the 222Rn ingrowth for the
rapid analysis of 226Ra in water samples. Nevertheless, their method leads to an overestimate of
the 226Ra concentration as a result of a systematic error.
Alternatively, the contribution of 235U can be obtained from its gamma lines at 143.8,
163.3 and 205.3 keV (De Corte et al., 2005, Justo et al., 2006). For example, Justo et al. (2006)
used 235U peak at 143.8 keV to directly measure 226Ra from 186.1 keV peak. However, no
comparison has been made to assess the accuracy of their results. Moreover, in another study
conducted by De Corte et al (2005), the 143.8 and 205.3 keV peaks were examined and
demonstrated to have unsatisfactory results due to large uncertainties associated with it. The
present study investigates the feasibility of using 235U peaks (143.8, 163.3 and 205.3 keV) for
the rapid determination of 226Ra.
6.1.3 Determination of 227Ac using gamma-ray spectrometry
The actinium isotope 227Ac is a member of the 235U series derived from the decay of
231Pa. It is a highly toxic radionuclide that can pose radiological health concerns to both humans
and the environment. The effective dose coefficients for inhalation and ingestion intakes of 227Ac
are more restrictive than those for 226Ra (ICRP, 2012). Although the activity of 227Ac in
environmental samples is expected to be extremely low, as the natural abundance of the whole
235U series is only 0.0072 of the 238U, it may be elevated to a higher level through certain human
activities. In an oil and gas industry, a level of 2.5 Bq g-1 of 227Ac was reported in scale samples
(Kolb and Wojcik, 1985). This level is 25 times higher than the exemption level recommended
by the International Atomic Energy Agency (IAEA, 2014).
The measurement of 227Ac is associated with difficulties due mostly to its low
abundance. The current analytical techniques available for the determination of 227Ac are ion
exchange with alpha spectrometry (Martin et al., 1995, Bojanowski et al., 1987), delayed
coincidence counting (Shaw and Moore, 2002), liquid scintillation counting (Kossert et al.,
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2015), and gamma-ray spectrometry (Van Beek et al., 2010). All of these techniques, with the
exception of gamma-ray spectrometry, require complicated chemical separation and extraction
processes. The present work uses a BEGe detector to establish analytical techniques for
measuring 227Ac using gamma-ray spectrometry in soil samples.
6.2 Material and methods
6.2.1 Routine analysis of NORM
In the analysis of NORM with gamma spectrometry, a number of key radionuclides are
routinely measured. These radionuclides are listed in Table 6.1. This table has been generated
using the nuclear data given in the Table of Isotopes (Ekström and Firestone, 2004). In the
literature, several authors have commented on the photon peaks behaviour during
measurements. These comments are discussed in the paragraphs below and also summarised in
Table 6.1.
In the 238U series, seven nuclides, namely, 234Th, 234mPa, 230Th, 226Ra, 214Pb, 214Bi, and
210Pb, emit detectable gamma-rays. In practice, the direct daughters 234Th and 234mPa are most
preferable for such measurements (Yücel et al., 2009, Huy and Luyen, 2004, Lenka et al., 2009,
Yücel et al., 1998) due to their short half-lives of 24.1 d and 1.17 min, respectively. Moreover,
these radioisotopes are the early members of the 238U series, which means that they require less
time to achieve secular equilibrium with their parent than other radionuclides in the series.
234Th emits several gamma-lines, at 63.3, 92.4 and 92.8 keV. The latter peaks are doublets and
interfere with fluorescence x-rays, such as Kα1 of thorium series radionuclides (228Ac) at 93.3
keV. This makes it difficult to deconvolute. Therefore, this peak was avoided, although some
authors found this peak provided acceptable results (Kaste et al., 2006, El-Daoushy and
Hernández, 2002).
The peaks at 63.3 (4.80 %) and 1001 keV (0.84 %) are more practical for estimating the
activity concentration of 238U. The gamma line at 63.3 keV suffers from self-absorption and from
interference with 232Th at 63.8 keV (0.26%) and 231Th at 63.9 keV (0.02%). However, this self-
absorption can be corrected (Huy et al., 2013), and the contributions of 232Th and 231Th to the
peak 63.3 keV in environmental samples is normally negligible, unless the sample has higher
232Th or 235U content (Yücel et al., 2009). In this work, the interference is corrected by
subtracting the contributions of 232Th, such that the activity of 234Th (A234) is given as:
𝐴234 = [𝐶63
𝜀63 𝑚− ( 𝐼232 𝐴232)] .
𝑓𝑠
𝐼234 (6.1)
where C63 is the count rate of the 63 keV peak, A232 is the activity of 232Th (as determined
through its daughter, 228Ac), 63 is the calibration efficiency of the 63 keV peak, m is the mass of
the sample, fs is the correction factor of the self-absorption, and I232 and I234 are the photon
130
emission intensities of 63.8 and 63.3 keV peaks of 232Th and 234Th, respectively. Equation. 1
assumes a secular equilibrium between 238U and 234Th. This assumption may not be valid under
some special cases where 238U has been separated from the remainder of the series during the
past four months.
This research also employed the 1001 keV (0.84%) peak of 234mPa to estimate the
activity concentration of 238U. Despite having low gamma emissions, this peak has been
successfully used to determine 238U in environmental samples (Yücel et al., 1998). It also
exhibits potential interference with a very weak 228Ac gamma line at 1000.7 keV (0.005%). This
interference was ignored, since the contribution of 228Ac to the 1001 keV peak is negligible.
However, overlooking this contribution in materials rich in thorium may cause an error of 0.05
to 2.98% (Yücel et al., 2009), which is likely to be well within the range of experimental
uncertainties.
Neither the weighted averages of 214Pb and 214Bi nor the natural isotopic activity ratios
of 235U/238U were used to estimate the activity concentration of 238U. In the first case, the
equilibrium between these isotopes and 238U cannot be guaranteed in samples, because of the
differences in the chemical properties and environmental behaviours of the various long-lived
isotopes in between. The second approach is not always viable, because, in some parts of the
world, the ratio 238U/235U has been altered due to nuclear contamination (Danesi et al., 2003,
Yoshida et al., 2000, Sansone et al., 2001). In addition, the natural abundance of 235U is very low
(0.72%), which could increase the measurement’s uncertainty.
Other important radionuclides in the 238U series that have many applications in
environmental studies are 230Th, 226Ra and 210Pb. 230Th emits a number of weak gamma
radiations, such as 67.7 keV (0.38%) and 143.9 keV (0.049%). Only the 67.7 keV peak is
employed for gamma spectrometry (Lozano et al., 2001, Yücel et al., 2010). The higher efficiency
of the BEGe detector at lower gamma-rays makes the determination of 230Th through the 67.7
keV peak possible. The measurement of 226Ra is discussed in further detail in the next section.
The 210Pb nuclide has a single gamma-line at 46.5 keV (4.25%). It should be noted that both the
46.5 and 67.7 keV peaks are affected by self-absorption, which was corrected for each sample.
The actinium series, which was named after the first element discovered in the series,
227Ac, is headed by 235U. Within this series, four nuclides (235U, 227Th, 223Ra, and 219Rn) are
applicable to gamma spectrometry. The measurement of this series is often problematic due to
its low abundance in nature. The major peak of 235U at 185.7 keV (57.2%) interferes with the
186.2 keV (3.5%) peak of 226Ra. This interference can be corrected by subtracting 226Ra
contribution at the peak. The portion of 226Ra can be obtained from its daughters, 214Pb and
214Bi, in equilibrium. Care must be taken, so, 222Rn, the gaseous daughter of 226Ra and the direct
parent of 214Pb and 214Bi, loss not occur in these samples. Other peaks of 235U at 143.8 (10.96%),
131
163.3 (5.08%) and 205 keV (5.01%) can be used as alternative measures of 235U (Pöllänen et al.,
2003, De Corte et al., 2005), despite the probability for true summing with the 19.6 keV gamma
line (Garcıa-Talavera et al., 2001).
In the 232Th series, several radioisotopes, particularly 228Ac, 224Ra, 212Pb, 212Bi and 208Tl,
have strong gamma emissions. 232Th itself has a very weak gamma line at 63.8 kev (0.27 %),
which interferes with 234Th and 231Th. It is more easily determined through its short-lived
daughter, 228Ac, which has strong gamma lines at 911.2 (25.8%), 968.9 (15.8%), and 338.3
(11.3%) keV assuming secular equilibrium. 228Ac is also used to determine the activity
concentration of 228Ra, which emits gamma-rays that difficult to be detected due to their low
intensities.
Another important radionuclide in the 232Th series is 228Th. Unfortunately, this isotope
has a very weak gamma line at 215.98 keV (0.26%), which may be useful in thorium-rich
samples. In this study, the activity concentration of 228Th was determined through its 215.98
keV peak, as well as through its daughters, 224Ra, 212Pb, 212Bi and 208Tl. The gamma line of 224Ra
at 241.0 keV (4.1 %) interferes with the 242.0 keV (7.0%) line of 214Pb, which makes the use of
its daughters, 212Pb, 212Bi, and 208Tl, more practical. 212Pb is typically measured using its
strongest gamma line of 238.6 keV (44%). The other peak at 300.1 keV (3.3%) interferes with
other peaks from the actinium series (231Pa and 227Th) and thus is not recommended.
The gamma lines of 212Bi are not frequently used, as its strongest peak at 727.3 keV
(6.6%) interferes with the 726.9 keV (0.64 %) peak of 228Ac. Other 212Bi gamma lines at 1620.7
(1.5 %) and 39.9 keV (1.5 %) are weak, and the latter suffers from serious self-absorption.
These peaks can be used after correcting the results for interference and self-absorption.
However, the 208Tl peaks at 2614.53 (99 %), 583.2 (85 %), and 510.8 (25.8 %) were employed
instead. The accuracy of these peaks may be affected by potential interference at the latter peak,
with the annihilation peak occurring at 511 keV and a high probability of true coincidence
summing at the other two peaks (Garcıa-Talavera et al., 2001, Şahiner and Meriç, 2014).
The naturally occurring radionuclide 40K can also be quantified using gamma
spectrometry. The gamma line at 1460.8 keV (11%) was used to determine the activity
concentration of 40K. This gamma line interferes with the 1459.14 keV (0.80%) peak of 228Ac,
and the spectral interference can cause serious errors of up to 100% or more in the measured
activity of 40K, depending on the 232Th/40K ratio (Lavi et al., 2004). However, the contribution of
228Ac at the 1460.8 keV peak was corrected as described by Lavi et al. (2004), and the 40K
analysis was performed after that correction. The error in the activity concentration of 40K
caused by 228Ac can be calculated by multiplying the intensity ratio (I228Ac / I40K) at the peak by
the activity concentration ratio (232Th/40K). The uncertainty in the measured activity
132
concentration of 40K due to peak interference should be corrected accordingly. An example is
shown in Section 2.8.
Table 6.1 Gamma energy lines of key natural radionuclides (Ekström and Firestone, 2004)
Isotope Half-life Energy
(keV)
Emission
(%)
Note
238U series
234Th 24.1 d 63.29 4.80 Self-absorption is pronounced - Interferes with 232Th at
63.83keV(0.267%) 234mPa 1.17 min 1001.03 0.84 Potential interference with 1000.7 keV (0.005%) peak of 228Ac 230Th 7.54 ×104 a 67.67 0.38 Self-absorption is pronounced 226Ra 1600 a 186.10 3.50 Overlapping with 185.7 keV (57.2 %) peak of 235U 214Pb 26.8 min 351.92
295.21
35.80
18.50
214Bi 19.9 min 609.31
1120.29
1764.49
1238.11
44.80
14.80
15.36
5.86
A probability of true coincidence summing effect
A probability of true coincidence summing effect
A probability of true coincidence summing effect 210Pb 22.3 a 46.54 4.25 Self-absorption is more pronounced
232Th series
228Ac 6.15 h 911.20
968.97
338.32
964.77
463.01
26.60
16.20
11.27
5.11
4.44
A probability of true coincidence summing effect
Interfering with 223Ra (338keV, 2.79%)
228Th 1.912 a 215.98 0.26 Can be useful in thorium rich samples 224Ra 3.66 d 240.99 3.97 Interferes with 241 keV (7%) peak of 214Pb 212Pb 10.64 h 238.63 43.30 212Bi 60.55 min 39.86
727.33
1.09
6.58
Self-absorption is more pronounced
Interferes with 726.8 keV (0.64%) peak of 228Ac 208Tl 3.053 min 2614.53
583.19
510.77
860.56
99.00
84.50
22.60
12.42
All peaks are affected by the true coincidence summing
potential interference with 214Pb at 580.15 keV (0.35%)
interference with the annihilation peak at 511 keV
235U series
235U 7.038×108 a 185.71
143.76
163.36
205.31
57.2
10.96
5.08
5.01
Interferes with 226Ra.
Interferes with 223Ra(143.1keV,3.22%) and 230Th( 143.87keV,
0.049%) and a probability of true coincidence summing effect.
Potential interference with 228Ac at 204.03 keV (0.12%) and 228Th at 205.93 keV (0.0204%) in thorium rich samples
227Th 18.72 d 235.97
50.13
256.25
12.3
8.0
7.0
Self-absorption is more pronounced
223Ra 11.44 d 269.46
154.21
144.23
13.7
5.62
3.22
Interferes with 271.23 keV (10.8%) peak of 219Rn and 270.24
keV (3.43%) peak of 228Ac
Interferes with 153.98 keV (0.74%) of 228Ac
219Rn 3.96 s 271.23
401.81
10.80
6.24
Interferes with 223Ra and 228Ac
40K 1.28×109 a 1460.83 11.00 Interferes with 1459.14 keV (0.80 %) peak of 228Ac
133
6.2.2 Rapid measurements of 226Ra
The activity concentration of 226Ra was determined directly from the 186 keV peak by
subtracting the contribution of 235U from the peak. The 235U portion was estimated from its
gamma lines at 143.8, 163.3, and 205.3 keV. Although these gamma lines may be affected by
true coincidence summing with the 19.6 keV line (Garcıa-Talavera et al., 2001), the use of
standard calibration sources identical to those of the sample can minimise this effect (Gilmore,
2008). The activity concentration of 226Ra (ARa) is given by the following formula:
𝐴𝑅𝑎 = [𝐶186
𝜀186 𝑚 − ( 𝐼𝑈 𝐴𝑈)] .
𝑓𝑠
𝐼𝑅𝑎 (6.2)
where C186 is the count rate of the peak at 186 keV, AU is the activity of 235U, 186 is the
calibration efficiency of the 186 keV peak, and IU and IRa are the gamma emission intensities of
the 186 keV peak of 235U and 226Ra, respectively.
The activity concentration of 235U obtained from the peak at 143 keV was corrected for
the interference caused by 223Ra and 230Th, as follows:
𝐴𝑈(143 𝑘𝑒𝑉) = [𝐶143
𝜀143 𝑚 − ( 𝐼223𝑅𝑎 𝐴223𝑅𝑎) − ( 𝐼230𝑇ℎ 𝐴230𝑇ℎ)] .
𝑓𝑠
𝐼𝑈 (6.3)
where C143 is the count rate of the peak at 143 keV, A223Ra and A230Th are the activity
concentrations of 223Ra and 230Th, 143 is the calibration efficiency of the 143 keV peak, and I223Ra
and I230Th are the gamma emission intensities of the 143 keV peak of 223Ra and 230Th,
respectively. The 223Ra activity was measured using its daughter 219Rn at the 401.8 keV (as the
269.5 keV peak has more interferences), while the activity of 230Th was determined at the 67.7
keV peak.
The activity of 226Ra was also determined indirectly through its daughters, 214Pb and
214Bi, after three weeks of waiting time. The results were then compared with those of the direct
method. For the reference materials IAEA-434 and RGU-1, the measurements were repeated
twice (after 1 day and 30 days of preparing the sample) to examine the influence of the elapsed
time. The results were then compared with the IAEA values. The quoted uncertainties in this
method were two standard deviations and arise from the counting statistic, efficiency
calibration and peak interference corrections.
6.2.3 Determination of 227Ac using gamma ray spectrometry
227Ac decays through the emission of low-energy beta particles (98.62%) to 227Th and
alpha particles (1.38%) to 223Fr. This decay is associated with extremely low gamma intensity,
which makes the direct measurement of 227Ac via gamma spectrometry very challenging.
Instead, in this research, 227Ac was determined by measuring its daughters, 227Th, 223Ra, and
219Rn, after a delay period of at least four months. The delay period was dictated by the half-life
134
of 227Th, which is 18.7 days. Some samples may have already reached equilibrium state between
227Ac and 227Th; in that case, the waiting time is not required. The direct daughter, 227Th, emits a
number of detectable gamma-rays (Table. 6.1). The strongest line is 12.3% at 235.97 keV. This
line is useful to use, but is also likely to interfere with the 212Pb gamma line at 238.63 keV
(43.3%), especially in samples rich in thorium. The next gamma line of 227Th occurs at 50.13 keV
(8.0%), where the self-absorption is more pronounced. Another peak that can be used is the
256.25 keV (7%) peak, which is in close proximity to the 214Pb peak at 258.79 keV (0.55%).
223Ra and its short-lived daughter, 219Rn, have very useful gamma lines that can be used
to determine 227Ac. The peak of 223Ra at 269.46 keV (13.7 %) has been used after deconvoluting
the spectral interference due to the 270.24 keV (3.43 %) peak of 228Ac and the 271.23 keV
(10.8 %) peak of 219Rn (Van Beek et al., 2010). Desideri et al. (2008) used the gamma peak of
219Rn at 401.81 keV (6.6 %) to determine 227Ac in waste samples. 219Rn has another gamma line
at 271.23 keV (10.8 %); however, it interferes with the two gamma lines emitted from 223Ra and
228Ac.
227Ac was determined through its daughters 227Th (at 50.13 and 256.25 keV), 223Ra (at
269.46 keV), and 219Rn (at 401.81 keV). The 235.97 keV peak of 227Th was deliberately not used
due to its complex interference and the 269.46 keV peak was corrected for the contributions of
other radionuclides (i.e., 219Rn and 228Ac). The results were compared with the expected values
for 227Ac, as well as with the activity of 235U.
6.2.4 Sample preparation
Thirteen samples of different types were prepared in a simple and reproducible way. Six
of these samples are reference materials with known activity concentrations and seven are
representative samples. It should be noted that the activities given for the IAEA-434 sample
were corrected for decay and ingrowth, since the equilibrium is disturbed in the 238U series. The
representative samples were selected to cover a range of activity concentrations (from a few
tens of Becquerel to a few mega Becquerel per kilogram) and to represent different radioactive
equilibrium conditions. A description of all samples is provided in Table 6.2.
The samples were dried overnight at 110oC, crushed and sieved through a 2 mm sieve to
obtain homogeneity. Samples TI and TII were initially in clay boulder form. They were coarsely
ground in mortar pestle to generate a sample that was sieved. Each sample was then weighed,
compressed in a polyester petri dish (6 cm in diameter and 1.2 cm in height) using a 10 kpa
hydraulic press and sealed airtight with a silicone adhesive to retain radon. It has been found
that placing the reference material RGU-1 in a sealed cylindrical plastic container with a
diameter of 6 cm can reduce radon loss to 0.8% (Scholten et al., 2013).
135
Table 6.2 Sample descriptions
Sample code Description Bulk Density (g cm-3)
Reference materials
IAEA-434 Certified reference material-Phosphogypsum 0.91
RGU-1 Certified reference material- uranium ore 1.31
RGTh-1 Certified reference material- thorium ore 1.30
RGK-1 Certified reference material- potassium sulphate 1.35
SS Spiked SiO2 1.39
SG Spiked grass 0.25
Representative samples
CHI Soil sample 1.12
CHII Soil sample 1.10
TI Uranium tailing 1.59
TII Uranium tailing 1.59
BC Black concentrate 2.09
CS Contaminated soil* 2.36
MO Monazite rich with thorium 3.51 * Contaminated due to a past mineral sand processing operation
6.2.5 Analytical technique
The BEGe detector (BE3820) used in this study is manufactured by Canberra and is
made of a high-purity germanium (HPGe) crystal housed in a 100 mm thick cylindrical graded
lead shield (Model 747, Canberra). The detector has good resolution, with a Full Width at Half
Maximum (FWHM) of 0.59 keV at 122 keV and 1.77 keV at 1332 keV and an optimum efficiency
across a wide range of gamma energies between 3 keV and 3 MeV.
The counting time was set at 48 hrs to achieve better statistics. The net count at the peak
of interest was obtained using Genie 2000 software supplied by Canberra. The activity
concentrations of the natural radionuclides in each sample were calculated via direct
comparison with a calibration standard with the same geometry. The calibration standard was
prepared by diluting certified reference materials supplied by New Brunswick Laboratory
(NBL) in silica powder. The diluted materials contain uranium and thorium ores. Reagent grade
potassium chloride supplied by Ajax Chemicals (Australia) was also added to the diluted
materials. A blank sample with the same matrix and composition as the standard was prepared,
using identical geometry to make background corrections.
The comparative method is highly recommended for environmental analysis (Gilmore,
2008). In this method, the effects of geometry, self-absorption, and coincidence summing are
cancelled out in the standard and the sample. However, in environmental samples, some
variations may exist mostly in the density, mineralogy and activity strength. In such situations,
correction factors are required for both self-absorption and random summing. Corrections to
the true summing may not be required, since the standard and the sample have identical
geometries and are analysed for the same nuclides.
136
The calibration data were obtained empirically using the prepared calibration standard.
The counts per photon emission of the calibration standard as a function of photon energy are
plotted in Figure 6.1(A). The results of the reference materials are plotted in the same figure as
well to check the performance of the detector. These values were obtained using the same
geometry, petri dishes. The measured counts per emission of the calibration standard and
reference materials were consistent. The results indicate a higher efficiency of the detector at
low gamma energy. The overall uncertainty in the calibration data is estimated to be 4% and it
arises from the uncertainty in the source activity and counting statistics.
The empirical calibration of the BEGe detector was compared with an HPGe coaxial
detector after preparing the calibration standard material in Marinelli Beaker geometry. Figure
6.1 (B) shows a superior efficiency of the BEGe detector at lower gamma energy and
comparable values at higher gamma energy. Furthermore, 39.86 and 67.67 keV peaks of 212Bi
and 234Th were only detected with the BEGe detector.
The analytical performance of the laboratory was tested using two statistical
parameters: relative bias and u-score. The relative bias identified the potential difference
between the analyst’s results and the IAEA assigned values using the following relation:
𝑅𝑒𝑙𝑎𝑡𝑖𝑣𝑒 𝑏𝑖𝑎𝑠 =(𝐴𝑛𝑎𝑙𝑦𝑠𝑡 𝑣𝑎𝑙𝑢𝑒−𝐴𝑠𝑠𝑖𝑔𝑛𝑒𝑑 𝑣𝑎𝑙𝑢𝑒)
𝐴𝑠𝑠𝑖𝑔𝑛𝑒𝑑 𝑣𝑎𝑙𝑢𝑒× 100 (6.4)
The u-scores examined whether the analyst’s results differed significantly from the
certified values, taking into account their uncertainties. The u-scores were calculated using the
following equation (Brookes et al., 1979):
𝑢𝑠𝑐𝑜𝑟𝑒 =(𝐴𝑛𝑎𝑙𝑦𝑠𝑡 𝑣𝑎𝑙𝑢𝑒−𝐴𝑠𝑠𝑖𝑔𝑛𝑒𝑑 𝑣𝑎𝑙𝑢𝑒)
√[(𝐴𝑛𝑎𝑙𝑦𝑠𝑡 𝑢𝑛𝑐𝑒𝑟𝑡𝑎𝑖𝑛𝑡𝑦)2+(𝐴𝑠𝑠𝑖𝑔𝑛𝑒𝑑 𝑢𝑛𝑐𝑒𝑟𝑡𝑎𝑖𝑛𝑡𝑦)2] (6.5)
The analysed and assigned uncertainties are quoted as two standard deviations. In order
to determine the significance of the comparison of the measured and expected values, the u-
scores were compared with the critical values listed in the t-test table. The u-test parameter of
1.96 for a confidence interval level of 95% was applied. Thus, the results were considered to
pass the test when u-score 1.96.
137
(A)
Energy (keV)
100 1000
cp
s/p
ho
ton e
mis
sio
n
0.00
0.02
0.04
0.06
0.08
0.10
0.12
0.14
RG-set
IAEA-434
SS
SG
Calibration standard
(B)
Energy (keV)
100 1000
cp
s/p
ho
ton e
mis
sio
n
0.00
0.02
0.04
0.06
0.08
0.10
0.12
0.14
BEGe_Petri dish
HPGe_Marinelli beaker
Figure 6.1 (A) Count per emission versus the gamma energy for the reference materials using the BEGe detector. (B) A comparison between the detection efficiency of a BEGe detector and an HPGe coaxial detector. The results are for a calibration standard in petri dish geometry (50 g) in BEGe and Marinelli Beaker geometry (1000 g) in HPGe-coaxial.
6.2.6 Gamma self-attenuation
The phenomenon of self-attenuation occurs when gamma-rays pass through different
materials, and it depends on both the gamma energy of the incident photon and the density and
composition of the material. The effect is more pronounced at low gamma energies, which are
associated with some natural radionuclides, such as 210Pb, 234Th, and 230Th (Pilleyre et al., 2006).
In the comparison method used to determine the activity concentration, the measured sample is
ideally required to have a geometry, density, and composition identical to those of the
calibration standard. This approach ensures that the self-attenuation effect and the other effects
in the standard are similar to those of the sample and, thus, cancel each other out. However,
variations in density and composition do exist. In order to cover a wide range of environmental
samples, one option may be to prepare a large number of calibration standards with different
densities and compositions, which is costly and time consuming. Alternatively, a correction
factor can be worked out and applied to the results.
There are several methods to estimate the correction factor of the self-attenuation
effect. In this study, a practical approach using the direct transmission of gamma-rays was used.
This method was proposed by Cushal (1983) and it is based on measuring the transmission of
gamma-rays through a sample by positioning a point source above the sample. The self-
attenuation correction factor (fs) for a sample relative to the calibration standard is given as
(Cutshall et al., 1983):
𝑓𝑠 =ln(𝐼 𝐼0)⁄
𝑠𝑎𝑚𝑝𝑙𝑒
ln(𝐼 𝐼0)⁄𝑠𝑡𝑎𝑛𝑑𝑎𝑟𝑑
×(𝐼0−𝐼𝑠𝑡𝑎𝑛𝑑𝑎𝑟𝑑)
(𝐼0−𝐼𝑠𝑎𝑚𝑝𝑙𝑒) (6.6)
where I and Io are the intensities with and without attenuation, respectively, expressed in count
rates. The correction factors are limited to gamma energies emitted from the available sources;
241Am, 137Cs, and 60Co (their activities ranging between 8 to 40 kBq). For other gamma energies,
138
the correction factors were interpolated from a self-absorption curve thus produced. The values
of fs for different samples used in this study are listed in Table 6.3†.
6.2.7 Coincidence summing
Coincidence summing occurs when two or more photons are recorded almost
simultaneously by a detector. If the incident photons are emitted by the same nuclide, the
summing is true, and if they originate from different nuclides, the summing is random. True
coincidence summing depends on the decay scheme of the radionuclide, as well as on the
geometry of the sample. Coincidence summing can introduce substantial error into the
measurement of activity concentrations, which requires correction. In the present study, the
summing errors in the standard and the sample are cancelled out, since the standard contains
reference materials with the same nuclides and geometry as the sample.
On the other hand, random summing is count-rate-dependant, and as long as the
activities of the sample and the calibration standard are comparable, it will be cancelled out. For
environmental samples, the count rate is typically low, which means negligible random
summing. However, if the activity of the sample is significantly higher than that of the standard,
random summing should be taken into account. This study has adopted an approach, which is
based on a method described by Wyttenbach (1971). The random summing correction factor
(fR) is calculated as follows (Wyttenbach, 1971):
𝑓𝑅 =1
𝑒−𝑊𝑓 (
𝑇𝑡
−1) (6.7)
where T is the real time, t is the live time, and Wf is the Wyttenbach factor = (2τ/θ)—where τ
represents the resolution time and θ is the sum of the fixed conversion time and the mean
channel-dependent conversion time of the memory cycle of the multi-channel analyser. The Wf
factor was determined experimentally using an approach given in (Wyttenbach, 1971) by
placing a 137Cs point source at a fixed position (100 mm) above the detector to yield an
uncertainty of less than 1% in the net area of the 661.6 keV peak. Another source, 241Am, was
introduced at a position above the 137Cs source and was moved closer to the detector through a
stepwise approach to vary the dead time. The ratio of the count rates at the peak (661.6 keV)
was plotted against the ratio of the real time to the live time, and the gradient was Wf. The
correction factors presented in Table 6.3 were given relative to the calibration standard. The fR
values for reference materials are equal to unity, as they have the same matrices and
† Other methods have been proposed to estimate the self-attenuation correction factor. For example,
theoretical and simulation approaches have been developed by some researchers (Dababneh et al., 2014;
McMahon et al., 2004; Pilleyre et al., 2006). However, this technique requires precise knowledge of the
sample geometry and composition, as well as of the detector characteristics. (Note: footnotes were added
to the thesis, not present in the published paper)
139
comparable activities with the calibration standards. This was confirmed by the experiment as
well. The fR values of CHII and TII samples were similar to that of CHI and TII, respectively.
Table 6.3 Correction factors for self-absorption (fs) and random summing (fR); material densities (g cm-3) are given between the brackets. A brief description of samples is given in Table. 2.
Correction Energy (keV) Nuclide SG CHI TI BC CS MO factor (0.25) (1.10) (1.59) (2.09) (2.36) (3.51) fs 46.54 210Pb 0.69 0.98 1.41 3.39 5.51 26.50 50.13 227Th 0.70 0.98 1.39 3.05 4.88 23.92 63.29 234Th 0.71 0.99 1.34 2.19 3.34 17.33 67.67 230Th 0.72 0.99 1.33 2.00 3.00 15.80 143.76 235U 0.77 0.99 1.19 1.34 1.89 2.02 186.00 226Ra+235U 0.79 0.99 1.14 1.25 1.56 1.85 205.31 228Th 0.80 0.99 1.12 1.22 1.44 1.78 215.99 228Th 0.81 0.99 1.12 1.20 1.39 1.75 238.63 212Pb 0.81 0.99 1.10 1.17 1.29 1.69 295.21 214Pb 0.83 0.99 1.07 1.11 1.10 1.57 351.92 214Pb 0.85 0.99 1.04 1.06 1.07 1.48 401.81 219Rn 0.86 0.99 1.02 1.06 1.07 1.41 609.31 214Bi 0.89 0.99 1.00 1.04 1.07 1.22 911.21 228Ac 0.93 1.00 1.00 1.00 1.06 1.18 1001.03 234mPa 0.94 1.00 1.00 1.00 1.06 1.17 1460.83 40K 0.97 1.00 1.00 1.00 1.06 1.14 2614.53 208Tl 1.00 1.00 1.00 1.00 1.05 1.09 fR 1.000 1.001 1.004 1.001 1.035 1.205
6.2.8 Uncertainties
All uncertainties in this paper have been computed to two standard deviations
stemming from the statistics of counting only. Other sources of uncertainties including detection
efficiency, peak interference correction factors, and the uncertainties in the certified activities
were also taken into account. The latter uncertainty was specified by the supplier of the
standard materials. The individual uncertainties were combined to give the overall
uncertainties according to ISO (1995). There are a number of uncertainties assumed to be
negligible due to their fairly low values. These include uncertainties in the acquisition time, the
half-life, the sample weight, and in the sample preparation procedure. Uncertainties of
interference correction that occur at some peaks are also considered. For example, the
associated uncertainty with the activity concentration of 226Ra (u(ARa)) determined through the
rapid analysis, i.e. eq.6.2, is calculated as follows:
If: 𝐴𝑅𝑎 = [𝐶186
𝜀186 𝑚 − ( 𝐼𝑈 𝐴𝑈)] .
𝑓𝑠
𝐼𝑅𝑎 (equation. 6.2), then:
u(𝐴𝑅𝑎) = √(𝐶186 𝑓𝑠
𝜀186 𝑚 𝐼𝑅𝑎 )
2[(
𝑢(𝐶186)
𝐶186 )
2+ (
𝑢(𝜀186)
𝜀186 )
2] + (
𝐴𝑈 𝐼𝑈𝑓𝑠
𝐼𝑅𝑎 )
2[(
𝑢(𝐴𝑈)
𝐴𝑈 )
2] (6.8)
where u(AU), u(C186), and u(186) are the standard uncertainties for the activity concentrations of
235U, the count rate, and the efficiency at the 186 keV peak. Note that u(ARa) represents
140
uncertainty to one standard deviation. In this paper, the reported values are to two standard
deviations, i.e. 2 u(ARa).
6.3 Results and discussion
6.3.1 Routine analysis of NORM
The activity concentrations of the key radioisotopes were first determined for the
reference materials. The measured activity concentrations of radionuclides in the reference
materials were compared with certified values as shown in Figure 6.2. In general, the measured
values were in excellent agreement with the certified values (Table 6.4). The results of the u-
score statistic revealed no significant difference between the analyst’s and assigned values
(Table 6.4), such that the calculated u-scores were lower than 1.96 at a 95% confidence interval.
The relative biases of the reference materials were within a deviation range of -5.50% to 4.39%.
The activity concentrations of 235U and its daughters were only reported for the RGU-1
sample, as the expected values were given by IAEA. 235U was measured using its gamma lines at
the 185.7 keV peak. 227Th was measured using its gamma lines at 50.13 and 256.25 keV, while
its short-lived progeny, 223Ra and 219Rn, were measured through their gamma lines at 269.46
and 401.81 keV, respectively. The 269.46 keV peak was corrected for the interference resulting
from 228Ac and 219Rn. The results of 235U and its daughters, 227Th, 223Ra, and 219Rn, agreed with
the recommended values within the range of statistical accuracy.
The RGTh-1 standard material was used to check the activity measurements of several
radionuclides in the 232Th series (Figure 6.2(A)). The activity of 228Ac is often used to estimate
the activity concentrations of 232Th and 228Ra as well. The measured and expected values of
228Ac indicate excellent agreement with a u-score of 0.28 and a relative bias of 1.6%.
Interestingly, the 228Th activity could also be measured directly through its weak gamma line
(0.263 %) at 215.98 keV. Although the measured activity agreed with the expected value, the
uncertainty at this peak was relatively high (Figure 6.2(A)). The other short-lived daughters in
the series (namely, 212Pb, 212Bi and 208Tl), had activities comparable to the recommended values.
The activity concentrations of 40K in the reference materials, RGK-1, SS and SG, were
measured and compared with the given values (Figure 6.2(C)). The measured activities of 40K
were found to be in excellent agreement with the assigned values (the relative bias for the RGK-
1 was -1.93 % with u-score value of 0.23).
In Figure 6.2(B), ratios between the measured and expected activities of radionuclides
in the 238U decay chain in reference materials fell between 1.01 ± 0.08% and 1.08 ± 0.20% for
SS, and between 0.87 ± 0.12% and 1.05 ± 0.15% for SG. The corresponding values for the 232Th
series radionuclides, the measured activities of the SS and SG materials were between 0.91 ±
0.10% and 1.17 ± 0.11% of the expected values, respectively.
141
RGU-1
234Th
234mPa
230Th
226Ra
210Pb
235U
227Th
223Ra
Activi
ty c
once
ntr
atio
n (
Bq
kg
-1)
200
5000
Measured
Expected
± 2SD
SS and SG - 238
U series
Radionuclide
234Th
234mPa
230Th
226Ra
210Pb
Activi
ty r
atio
0.0
0.5
1.0
1.5
2.0
SS
SG
Expected
RGTh-1
228Ac
228Th
212Pb
212Bi
208Tl
Activi
ty c
once
ntr
atio
n (
Bq
kg
-1)
2500
3000
3500
4000
Measured
Expected
± 2SD
SS and SG - 232
Th series
Radionuclide
228Ac
228Th
212Pb
212Bi
208Tl
Activi
ty r
atio
0.0
0.5
1.0
1.5
2.0
SS
SG
Expected
(B) Activity ratios of several radionuclides in the uranium and thorium series for the spiked samples (SS and SG) prepared from the NBL standards. The continuous line represents a ratio of 1:00 and the error bars are the uncertainty to 2.
IAEA-434
234Th
234mPa
230Th
226Ra
210Pb
Activi
ty c
once
ntr
atio
n (
Bq
kg
-1)
0
200
600
800
1000
Measured
Expected
± 2SD
40K
Sample
SS SG RGK-1
40K
activi
ty c
once
ntr
atio
n (
Bq
kg
-1)
0
500
15000 Measured
Expected
± 2SD
142
Figure 6.2 (A) The measured activity concentration (dots) compared to the certified values (continuous line) for radioisotopes in the reference materials used in the present study, RGU-1, RGTh-1, and IAEA-434. The comparison is for uranium, thorium, and actinium radioisotopes. The uncertainty (2) in the certified values is given as dashed lines and in the measured values as error bars.
The results of the representative samples are presented in Table 6.5. In all samples, the
activity of 234Th at the 63.29 keV peak agreed with that of 234mPa obtained from the 1001.03 keV
peak. Both peaks were able to provide reliable estimations for 238U in most of the analysed
samples. However, the 63.29 keV peaks in the two highest density samples, CS and MO, did not
provide results consistent with those obtained from 234mPa. This discrepancy could be related to
the larger self-absorption effect, because of the large densities of these samples, occurring at the
63.29 keV peak than at the 1001.03 keV peak.
The 230Th activity was measured via its gamma line at 67.67 keV. Despite its lower
probability of gamma emission and higher self-absorption at this peak, the activities obtained
appeared to align well with those of other radionuclides in the series. This was true for samples
for which the equilibrium in the series was guaranteed (i.e. when the activities of 238U, 230Th,
226Ra, and 210Pb were equivalent).
The activity concentration of 226Ra was calculated as the weighted average of its short-
lived progeny 214Pb and 214Bi, after allowing four weeks for secular equilibrium. The activity of
226Ra was consistent with its daughter isotope 210Pb in all samples, indicating equilibrium in the
lower part of the series with no excess 210Pb activity.
Table 6.4 U-scores and relative bias for reference materials
Reference material Nuclide Activity concentration (Bq kg-1) Bias U-Score
IAEA Measured (%)
IAEA-434 238U 120 ±22 119 ± 8 -0.59 0.03
230Th 211 ± 18 209 ± 28 -0.73 0.05
226Ra 778 ± 124 773 ± 16 -0.67 0.04
210Pb 701 ± 120 732 ± 17 4.39 0.37
RGU-1 238U 4940 ± 60 4832 ± 150 -2.17 0.67
230Th 4967 ± 286 0.54 0.09
226Ra 4990 ± 82 1.02 0.49
210Pb 5073 ± 182 2.70 0.69
235U 228 ± 4 221 ± 10 -2.94 0.62
227Th 232 ± 12 1.58 0.29
223Ra 226 ± 4 -0.95 0.38
219Rn 215 ± 28 -5.50 0.44
RGTh-1 228Ac 3250 ± 180 3302 ± 36 1.60 0.28
228Th 3229 ± 254 -0.63 0.07
212Pb 3295 ± 22 1.38 0.25
208Tl 3317 ± 46 2.05 0.33
RGK-1 40K 14000 ± 800 13730 ± 440 -1.93 0.23
(c) A comparison of 40K measured activity concentrations and referenced values for the reference samples SS, SG, and RGK-1.
143
For the upper part of the series, an equilibrium state between 226Ra and its parent 238U
was observed in most samples (Table 6.5). However, for the TI, TII and BC samples, the activity
concentrations for the lower part of the series (i.e., 226Ra and its progeny) were higher than
those for the upper part. This implies that the equilibrium states of these samples were
disturbed suggesting that during past industrial processing uranium has been extracted from
bulk materials.
The reported activities of radionuclides in the 235U series followed the same equilibrium
pattern as those in the 238U series (Table 6.5). The 235U/238U ratio remained close to the natural
value of 0.046 in all samples. The 235U/226Ra ratio was altered in disequilibrium samples. This is
evidence that 235U atoms were also removed with the removal of 238U atoms and left the series.
In the 232Th series, equilibrium was present in all analysed samples (Table 6.5). Moreover, the
direct measurement of 228Th from its weak gamma line at 215.98 keV (0.26%) appeared to be
possible with the BEGe detector. In most samples, the measured activity of 228Th through the
215.98 keV peak was comparable to that obtained from its short-lived daughters, 212Pb, 212Bi,
and 208Tl. However, in samples with low thorium concentrations (100 Bq kg-1), such as CHI and
CHII, the 215.98 keV peak was unreliable. This is most likely related to the low probability of
gamma emissions within that line, along with other factors (e.g., self-absorption).
In the spectrum of samples with high count rates, additional interferences have been
observed. For example, the 238.63 keV peak of 212Pb overlapped with the 235.97 keV peak of
227Th and the 241 keV doublet peaks of 214Pb and 224Ra. This is because, at high count rates, the
resolution is degraded, which may broaden the peaks (Debertin and Helmer, 1988) and cause
adjacent peaks to overlap. However, the 238.63 keV peak was successfully corrected for this
interference. Furthermore, the 510.77 and 583.19 keV peaks of 208Tl were avoided due to the
complex interference among the annihilation peak at 511 keV, the 214Pb peak at 580.15 keV
(0.35%) and the 228Ac at 583.4 keV (0.11%). Instead, the 2614.5 (99 %) and 860.6 keV (12.4 %)
peaks were used. Likewise, the 235U peak at 143.8 keV overlapped with two 228Ac peaks at
145.85 (0.16%) and 141.02 keV (0.05 %). This occurred in addition to the contributions of 230Th
and 223Ra at the same peak. The correction for such interference can be risky due to the large
number of interferences and will probably introduce additional errors to the final results.
The spectral interferences that occurred at the 63, 143, and 270 keV peaks of 234Th, 235U and
223Ra, respectively, were successfully corrected in most samples. The contributions at the
interfered peaks were due mainly to radionuclides in the 232Th series, including 232Th at 63 keV
and 228Ac at the other two peaks. In the CS and MO samples, however, the results obtained from
the overlapped peaks were unreliable. These samples exhibited higher thorium contents, such
that the 232Th/238U ratio was greater than five in the CS and seven in the MO sample. However,
144
the unreliability of the correction may not be related to the 232Th/238U ratio. For example, in the
case of RGTh-1 sample, correction could be successfully applied despite the 232Th/238U ratio
exceeding 40. Other factors, such as, random summing and self-absorption would have been
responsible for this effect. These factors are discussed below.
The activity strengths of MO and CS samples are higher than the activity of the calibration
standard by a factor of 30 or more. This has an impact on the comparative method commonly
used for NORM samples analysis. The probability of random summing, which is count-rate
dependant, in these samples will be much greater than it is in the calibration standard. As a
result, there will be a gain or loss in the number of events at the peaks of interest. Another
observation in the CS and MO samples is the elevated spectral background, which originated
from Compton scattering. This can reduce the signal-to-noise ratio. Most of the overlapped
peaks lay in the region dominated by Compton scattering. Moreover, the self-absorptions of the
CS and MO samples were relatively large. Although, the random summing and self-absorption in
these samples have been corrected, the large values of such factors may add complexity in a way
that is not understandable.
The activity concentration of 40K was reported only for TI and TII samples. The level of
40K in these samples was within the global population weighted average (420 Bq kg-1)
(UNSCEAR, 2008). In the CHI and CHII samples, the activity of 40K was below the detection limit.
This is due to the natural formation of the underlying geology of the area from which the
samples were collected, which consists of basaltic rocks known to have low potassium content
and consequently low 40K. In other samples, in which the 232Th/238U ratio is high (i.e., MO and
CS), the 1460 keV peak was due mostly to 228Ac interfering with 40K. The error in 40K activity
caused by this interference was greater than 100%.
145
Table 6.5 The measured activity concentration of several radionuclides of NORM (Bq kg-1) in various samples used in the present study.
Series Nuclide CHI CHII TI TII BC CS MO Soil Soil U-tailing U-tailing Concentrate Contaminated soil Monazite 238U 234Th 1032 ± 56 666 ± 38 2006 ± 102 1976 ± 100 1295 ± 68 3460 ± 164 N.A 234mPa 946 ± 146 551 ± 102 1929 ± 228 1866 ± 222 1223 ± 150 6407 ± 566 34887 ± 3120 230Th 927 ± 74 567 ± 50 8557 ± 468 8599 ± 470 2346 ± 156 6639 ± 392 34652 ± 3420 226Ra 956 ± 50 555 ± 32 6347 ± 106 6337 ± 106 13833 ± 204 6242 ± 106 34194 ± 602 210Pb 984 ± 42 552 ± 26 5319 ± 194 5380 ± 196 15221 ± 540 6427 ± 250 33080 ± 1838 232Th 228Ac 83 ± 4 95 ± 4 167 ± 6 171 ± 6 83 ± 6 33763 ± 1208 248796 ± 5096 228Th BDL BDL 155 ± 32 55 ± 20 79 ± 20 35264 ± 2320 255584 ± 17658 212Pb 78 ± 4 96 ± 4 165 ± 6 169 ± 6 63 ± 2 35930 ± 2384 233136 ± 9324 212Bi 82 ± 10 95 ± 8 166 ± 14 156 ± 12 83 ± 8 33334 ± 1272 234724 ± 9192 208Tl 85 ± 4 96 ± 6 164 ± 10 163 ± 10 79 ± 6 33449 ± 704 235187 ± 6510 235U 235U 45 ± 4 31 ± 4 115 ± 8 119 ± 8 63 ± 6 312 ± 30 N.A 227Th 55 ± 6 29 ± 4 285 ± 18 239 ± 18 601 ± 12 240 ± 20 N.A 223Ra 44 ± 10 27 ± 10 338 ± 68 299 ± 60 646 ± 130 N.A N.A 219Rn 43 ± 10 34 ± 8 288 ± 34 267 ± 32 634 ± 66 337 ± 36 1902 ± 178 40K BDL BDL 575 ± 44 598 ± 44 BDL BDL BDL BDL = Below Detection Limit N.A = Not Applicable
146
6.3.2 Rapid measurements of 226Ra
The results of the rapid determination of 226Ra for the reference materials RGU-1 and
IAEA-434 after 1 day and 30 days of preparing the samples are illustrated in Figure 6.3(A). The
measured 226Ra activities obtained from 186 keV peak corrected through various gamma lines
of 235U and through the natural 235U/238U ratio fell within the statistical uncertainty of IAEA
values at all times. The relative bias between the measured 226Ra obtained through the rapid
approach and the IAEA values ranged from – 6.2% to 3.6% and the u-score values were found
satisfactory (u<1.96).
On the other hand, the results of 226Ra obtained from its progeny, 214Pb and 214Bi,
appeared to be dependent on the delay time required for the secular equilibrium. The activity
concentration of 226Ra, measured after a day of preparing the sample, was lower than that after
30 days. The discrepancy would depend on the loss of any 222Rn that has emanated from the
grains and released from the interstitial space during sample preparation. The results
demonstrate that measuring 226Ra soon after preparing the sample is possible using the 186 keV
peak, with results equivalent to those obtained from 214Pb and 214Bi after a waiting time of 30
days. Without this delay period, the latter method may underestimate the results. Of the
alternative suggested for rapid 226Ra measurement, the method based on 235U/238U ratio
assumes the natural activity concentration ratio of 0.046 between the two uranium isotopes.
This value may not may not be true for depleted or enriched uranium samples. The methods
based on the direct measurement of 235U peaks are not influenced either by the 235U/238U ratio
or the state of equilibrium between the uranium or actinium series.
The results of the measurements of 226Ra from the reference materials SS and SG are
plotted in Figure 6.3(B). Most of the observed data points are scattered fairly close to the line of
equality, indicating good agreement between the measured and expected values of 226Ra, with
comparable results across the examined methods. The results of the gamma line at 163.3 keV in
the SG sample were higher than the expected values by more than 30%; however, the result was
consistent with the statistical error.
147
RGU-1
Method
(a) Progeny (b) 143 keV (c) 163 keV (d) 205 keV (e) 235U(f) 238U/235U ratio
Activi
ty c
once
ntr
atio
n (
Bq
kg
-1)
3500
4000
4500
5000
5500
6000
1-2 days
After 30 days
Expected value
± 2SD
IAEA-434
Method
(a) Progeny (b) 143 keV (c) 163 keV (d) 205 keV (e) 235U(f) 238U/235U ratio
Activi
ty c
once
ntr
atio
n (
Bq
kg
-1)
400
600
800
1000
1200 1-2 days
After 30 days
Expected value
± 2SD
Figure 6.3 (A) The 226Ra activity concentration measured in the IAEA standards, RGU-1 and IAEA-434, using (a) 222Rn progeny gamma lines and the 186 keV peak after correcting it for the 235U contribution, calculated from 235U individual gamma lines at (b) 143, (c) 163, (d) 205 keV, (e) the combined average from these lines, and (f) from the 238U/235U ratio compared to the certified values.
148
Method
(a) Progeny (b) 143 keV (c) 163 keV (d) 205 keV (e) 235U(f) 238U/235U ratio
Re
lative
to
the
exp
ecte
d v
alu
e
0.0
0.5
1.0
1.5
2.0
SS
SG
Expected
Figure 6.3 (B) The 226Ra activity measured using different approaches compared to the expected values in the spiked samples.
The direct measure of 226Ra activity using the 186 keV peak was also applied to the
representative samples. The measured activity of 226Ra from each method was compared to the
activities obtained from 226Ra progeny, 214Pb and 214Bi, at equilibrium (Table 6.6). The ratios of
the activity of 226Ra, as measured through several gamma energy lines of 235U, to those of 214Pb
and 214Bi were close to unity for most samples.
The results for the 226Ra obtained via the 143 keV peak in the MO and CS samples were
unsatisfactory. Those samples were rich in 232Th; hence, two peaks of 228Ac at 145.85 (0.16%)
and 141.02 keV (0.05%) became pronounced and overlapped with the triplet peak at 143 keV
due to 235U, 223Ra, and 230Th. The complexity of the interference, along with the high Compton
continuum background and the large self-absorption factor, would lead to enormous errors at
this peak. Recall the low natural abundance of 235U and the weak gamma intensity at the 143
keV peak; the same is true for the 205.3 keV peak in the MO sample where it interferes with the
204.0 keV (0.12%) 228Ac peak. These results are consistent with those reported by De Corte et
al. (2005), where the use of 143 and 205.3 keV peaks yielded systematically high results.
The rapid determination of 226Ra appeared to be unreliable in the case of the MO sample.
The difference in the activity concentrations measured through the 226Ra progeny and measured
149
directly via 235U gamma-lines was significant, with the exception of the measurements taken at
the 163.3 keV peak. This peak resulted in an activity ratio of 1.17 ± 0.12% and was free of
interference. This is most likely due to the random summing and Compton continuum
backgrounds that result from the high count rates of this sample.
The uncertainty of the rapid 226Ra measurement is estimated to be 25% or lower. This
uncertainty arises mainly from counting statistics and calibration efficiencies. It is directly
related to the low count rates of 235U peaks, which are associated with low natural abundance,
as well as the low probability of emission of the 235U gamma lines employed in the direct
method. In addition, the interference that occurred at the targeted peaks can introduce another
source of error. This can be seen clearly in the high uncertainty associated with the results of
the 143 keV peak that interferes with 223Ra, 230Th, and in some cases, 228Ac.
The findings confirm the potential for using the gamma lines of 235U at 143.8, 163.3 and
205.3 keV for the rapid determination of 226Ra. This approach was found to be equivalent to the
conventional method of measuring 226Ra using gamma spectrometry (i.e., through 226Ra progeny
in equilibrium). However, the usefulness of the rapid approach appeared to be limited by the
systematic errors arising from the large differences in activity strengths between the sample
and the calibration standard. Due to these errors, the Compton scattering backgrounds, the
random summing in samples with high count rate spectra are likely to be elevated. High count
rates allow greater spectral interference to occur and cause complexity in correcting
interference that occurs at the peak of interest.
Table 6.6 Ratios of 226Ra activity as measured using different approaches to the ratio obtained via its progeny
Sample 143 163 205 235U 235U/238U
(keV) (keV) (keV) (weighted mean) Ratio
CHI 0.92 ± 0.10 1.00 ± 0.10 0.96 ± 0.10 0.97 ± 0.10 0.97 ± 0.10
CHII 0.93 ± 0.10 0.95 ± 0.10 1.32 ± 0.14 1.14 ± 0.12 1.05 ± 0.10
TI 1.13 ± 0.11 1.18 ± 0.08 1.09 ± 0.08 1.14 ± 0.10 1.17 ± 0.11
TII 1.11 ± 0.10 1.18 ± 0.08 1.07 ± 0.08 1.13 ± 0.10 1.19 ± 0.11
BC 1.22 ± 0.08 1.27 ± 0.08 1.28 ± 0.08 1.27 ± 0.08 1.27 ± 0.08
CS 4.26 ± 0.42 0.96 ± 0.12 1.08 ± 0.08 1.28 ± 0.11 1.17 ± 0.24
MO 7.03 ± 0.32 1.17 ± 0.12 5.44 ± 0.20 2.71 ± 0.14 0.27 ± 0.04
6.3.3 Determination of 227Ac using gamma ray spectrometry
The activity concentration of 227Ac was obtained as a weighted average of its daughter
radionuclides, 227Th, 223Ra, and 219Rn, after a waiting time sufficient for equilibrium. The
measured 227Ac activity concentration was compared with those of 235U, 238U, and 226Ra, as
presented in Table 6.7. The 227Ac/235U activity ratio was close to unity in the samples exhibiting
the equilibrium condition. In these samples, the 227Ac/238U and 227Ac/226Ra activity ratios
corresponded to the natural value of 0.046. On the other hand, the 227Ac /235U ratio was greater
150
than 2 in the IAEA-434, TI, TII, and BC samples. This confirms the disequilibrium state of the
235U series. The activity ratio of 227Ac/238U in these samples was also changed from the natural
value of 0.046, while the 227Ac/226Ra ratio remained the same. In case of samples TI, TII, and BC,
the removal of element uranium would have resulted in this behaviour. These samples have
been taken from the product and waste streams of mineral processing plants. However, the
exact reason of the IAEA-434 sample behaviour is not clear.
Figure 6.4 shows a plot of the mean activity concentration of 227Ac against the measured
activity of 235U in all samples. An excellent correlation was found between the two results in
samples that maintained equilibrium, as can be seen in the data points distributed closely to the
line of equality in Figure 6.4. When the equilibrium state was lost, the activity of 227Ac differed
significantly from that of 235U, as indicated by the outliers in Figure 6.4.
It is worth noting that under certain conditions, the use of 227Th and its daughters to
determine the activity concentration of 227Ac may be misleading due to the differences in their
chemical properties. In an acidic pH condition, for example, the activity concentration of 227Ac,
as measured through 227Th, was found to be incorrectly estimated due to the mobility of 227Th in
that condition (Patra et al., 2013).
Table 6.7 Activity ratios of 227Ac to 235U, 238U and 226Ra
Sample 227Ac/235U 227Ac/238U 227Ac/226Ra IAEA434 4.514 ± 1.108 0.208 ± 0.050 0.032 ± 0.008 RGU-1 1.062 ± 0.226 0.046 ± 0.008 0.046 ± 0.008 SS 0.859 ± 0.178 0.040 ± 0.008 0.039 ± 0.008 SG 1.068 ± 0.136 0.049 ± 0.014 0.050 ± 0.006 CHI 1.137 ± 0.150 0.052 ± 0.012 0.055 ± 0.006 CHII 0.961 ± 0.182 0.047 ± 0.012 0.053 ± 0.008 TI 2.855 ± 0.274 0.145 ± 0.024 0.045 ± 0.002 TII 2.332 ± 0.226 0.126 ± 0.022 0.038 ± 0.002 BC 10.876 ± 1.338 0.536 ± 0.088 0.049 ± 0.002 CS 0.918 ± 0.052 0.060 ± 0.012 0.040 ± 0.002 MO n.a 0.055 ± 0.010 0.056 ± 0.006
151
235U Activity (Bq kg
-1)
10 100 1000
22
7A
Activi
ty (
Bq
kg
- 1)
10
100
1000
IAEA-434
RGU-1
SS
SG
CHI
CHII
TI
TII
BC
CS
MO
Line of equality
Figure 6.4 The activity concentrations of 227Ac versus 235U in various samples examined during the present study. For most samples, the values lie along the line of equality. Samples such as TI, TII, and BC show the 227Ac:235U ratio 1. This behaviour is most likely due to extraction of uranium from these samples.
6.4 Conclusion
The BEGe detector was successfully employed for routine analyses of NORM samples.
The detector was found to be efficient for measuring radionuclides with low gamma energy and
intensity that have many applications in the field of environmental radioactivity, such as 46.5,
63.3, and 67.7 keV, of 210Pb, 234Th, and 230Th, respectively. Analysis of the same standard using a
BEGe and an HPGe coaxial detector demonstrated the superiority of the BEGe detection system
in the 20 – 200 keV range. A comparison of the activity concentrations of the analysed
radionuclides in reference materials measured by BEGe detector and assigned by the supplier
showed a good agreement with a bias range between -5.5% and 4.4% and the u-score values
were less than 1.96. Care should be exercised when using the comparative method to calibrate
the detector as the large difference in the densities and activity strengths between the sample
and the calibration standards can lead to unreliable results.
The findings demonstrated the feasibility of the rapid determination of 226Ra through
the direct measure of the 186 keV peak after subtracting the contribution of 235U at the peak.
The gamma lines at 143.8, 163.3, and 205.3 keV were succeeded, for most samples, in
152
estimating the activity concentration of 235U. The validity of this approach has been confirmed
by a statistical comparison with reference materials. It revealed that the activity concentrations
of 226Ra measured using the rapid approach agreed with the certified values to within two
standard deviations. The results suggested the use of this approach to save time in determining
the activity concentration of 226Ra in environmental samples and accomplish the analysis within
two days. However, the reliability of this approach may be limited by the large variations in the
source strength and density between the sample and the comparator, which may resulting in
systematic errors.
The activity concentration of 227Ac was successfully determined in the environmental
samples through its daughters using gamma-ray spectrometry. The results were consistent with
the expected values, as well as with the values for the parent nuclide, 235U, in samples where
secular equilibrium was attained. The drawback of this approach is, perhaps, the delay period
required for the secular equilibrium state (227Th, T1/2 = 18.7 d; 223Ra, T1/2 = 11.4 d). However,
this equilibrium may already exist in some samples.
Acknowledgements
The authors would like to thank Dr Jamie Trapp for preparing the calibration standard.
153
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Chapter 7 Summary and Conclusion
7.1 Conclusion
Despite the marked improvement in the radiation protection system, NORM is still an
evolving subject, and many challenges have to be overcome. This thesis critically reviews the
current regulation of NORM, identifies emerging challenges and provides a better
understanding of how the measurements and monitoring of NORM should be carried out to
adequately demonstrate compliance with regulatory regimes. In particular, the measurements
used in quantifying the activity concentration of radionuclide members of NORM in solid
materials, as well as the release of radon from material surfaces and its concentration in indoor
air, were investigated.
The current radiological protection system of NORM was comprehensively reviewed. A
number of regulatory issues were highlighted and discussed, including: the absence of a
complete framework; the interference between planned and existing exposure situations; the
lack of a homogeneous approach; the ubiquity of NORM; the source of low doses; societal
attitudes; measurement challenges; waste and residues, and the ethical approach. The absence
of a complete framework, along with the interference that might occur between planned and
existing exposure situations, has created inconsistency in the regulation of NORM. This has led
to confusion among users of NORM guidelines. More clarity may be required to eliminate such
confusion. In addition, societal attitudes toward natural sources of radiation have influenced the
decision in terms of radiation protection against NORM, resulting in less attention being paid to
it. The impact of the measurement (which is necessary to implement and comply with the
regulation) on the management of NORM is obvious. Improving NORM measurement techniques
is still needed to address the challenges associated with it.
The review of the regulation of NORM has also shown that each industry involving
NORM has its own merits. It is vital from a regulation point of view, therefore, to characterise
exposure to NORM and its potential pathways in each industry on a case-by-case basis.
Implications of the implementation of the new radon and thoron policies were also discussed
from economic and practical perspectives. Further steps need to be taken by national
authorities to develop and implement a national radon protection strategy to control indoor
radon exposure. Characterising such exposure cannot be accomplished without quantifying
radon levels in the air, and its exhalation flux from surfaces.
The measurement of radon flux density from surfaces using the activated charcoal
canister technique was examined. The distribution of the adsorbed 222Rn in the charcoal bed of
the measured canister was found to be slightly different from that in the standard canister, and
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became more uniform with longer exposure times. This is because calibration standards were
prepared by a uniform mixing of 226Ra in the charcoal matrix, while in the field 222Rn was
adsorbed more readily in the charcoal layer closer to the exposed surface. A correction factor
was worked out by differences in the counting geometries between the uniformly distributed
standards and the measured canisters with non-uniform distribution. Its value for one day of
exposure, which recorded the lowest degree of uniformity, was at 14% or less. The influence of
exposure time on the measured 222Rn activity flux was also investigated. The measured value of
the activity flux was found to decrease as exposure time increased. The reduction was
attributed to the back diffusion of 222Rn from the canister through a canister-soil interface, and
possibly the absorption of water molecules by the activated charcoal. These findings introduced
limitations to the methodology, which need to be taken into account when interpreting the
measurements.
The levels of indoor radon (222Rn) and thoron (220Rn) were measured in various
buildings representing workplaces in Brisbane CBD, Australia. The average 222Rn and 220Rn
concentrations for all workplaces were 10.5 ± 11.3 Bq m−3 and 8.2 ± 1.4 Bq m−3 respectively. The
highest 222Rn concentration was found in basements and confined areas, while the maximum
220Rn level was detected in storage rooms. Poor ventilation in these locations is most likely the
main cause of the increased concentration levels. Diurnal variation in 222Rn level was observed
in the workplaces-concentrations were lower during working hours due to the use of air
conditioners and natural ventilation during this time. Changes in 220Rn concentration were
insignificant during and after working hours. The study extended to investigating the
parameters that could influence indoor 222Rn and 220Rn concentrations. 222Rn was more
influenced by floor level and flooring type, while 220Rn was most affected by the distance from
the source of origin. The findings suggest that care should be exercised when measuring 222Rn
concentrations in workplaces, as its concentration during and after working hours differ
considerably. The study demonstrated that indoor 222Rn and 220Rn levels are influenced by the
parameters examined in this study.
The suitability of the BEGe detector for routine analysis of NORM was investigated. The
detector was found to be useful for the activity concentration measurement of natural
radionuclides and was capable of detecting peaks with low probabilities of gamma emissions at
lower energy regions, such as 46.5, 50.1, 63.3 and 67.7 keV, of 210Pb, 227Th, 234Th and 230Th,
respectively. The detector was used for the rapid determination of 226Ra through the direct
measure of the 186 keV peak. Although this approach succeeded in most cases, its reliability is
limited by the systematic errors arising from the large variations in activity strength and
density between the sample and the calibration standard. The outcome of this study is
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encouraging, and the detector has a promising future in a variety of applications in the field of
environmental radioactivity.
7.2 Future Direction
This thesis has shown the importance of understanding the measurement and
regulation of NORM, and has opened new horizons for future research projects on the topic. It
has been shown that the regulation of NORM within the relevant industries has to be treated on
a case-by-case basis. Information about NORM in non-nuclear industries is scarce, so
investigation of different types of industries and industrial processes would be beneficial in
assessing the radiological impact of NORM on a wide scale. The regulatory challenges associated
with NORM that have been identified in this thesis should be addressed and taken into account
in future recommendations and policies. Additionally, the influence of measurement issues (e.g.,
the cost of the analysis, uncertainties of the measurement, the background radiation of
detectors, and the state of radioactive equilibrium) on adequate compliance with NORM
regulations needs further investigation. A particularly important example is the development of
a cost-effective method for measuring certain radionuclides of NORM in specific environmental
media.
The findings of each measurement technique investigated in this thesis highlight a
number of further works that are yet to be explored:
The study of measuring radon activity flux density using an activated charcoal canister
outlined the limitations of the method owing to exposure time and the distribution of
radon in the charcoal bed. A study of the effects of the charcoal type and charcoal grain
size on radon adsorption would provide a better understanding of the adsorption
efficiency of the charcoal.
The measurements of radon and thoron in workplaces provide the average
concentrations of these gases but do not determine radon and thoron progeny
distribution in indoor air. A comprehensive approach that includes a time series of
radon and thoron, as well as their progeny, with suitable modelling and analysis
techniques would allow for a better radiological assessment of exposure to radon and
thoron in indoor environments.
The BEGe detector was used only for the measurement of solid NORM samples.
Investigation of the use of the detector for the analysis of filters and water samples
would be beneficial for a comprehensive analysis of NORM encountered in industrial
situations.
The results of these research projects will significantly improve the measurement and
regulation of NORM.
