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Working Report 2016-02
Monitoring Radiation Effects in the Environment
Karen Smith
April 2016
POSIVA OY
Olkiluoto
FI-27160 EURAJOKI, FINLAND
Phone (02) 8372 31 (nat.), (+358-2-) 8372 31 (int.)
Fax (02) 8372 3809 (nat.), (+358-2-) 8372 3809 (int.)
Working Report 2016-02
Monitoring Radiation Effects in the Environment
Karen Smith
RadEcol Consulting Ltd
April 2016
Working Reports contain information on work in progress
or pending completion.
MONITORING RADIATION EFFECTS IN THE ENVIRONMENT
ABSTRACT
Olkiluoto Island, in southwest Finland is currently home to two operational nuclear
power plants, with a third under construction, and a repository for low and intermediate
level waste. The island is also the site of a spent nuclear fuel repository, currently under
construction following the construction licence granted in 2015. Posiva, the company
responsible for the management of spent nuclear fuel of its owners in Finland, has
undertaken a long-term monitoring programme on the island to derive data from the
environment and geosphere in support of safety assessments and to evaluate the impacts
of construction activities. The current monitoring programme is valid for the
construction phase of the repository and will be updated before the operation phase
begins. A pre-existing radiation monitoring programme is also in place as a result of the
nuclear power plants before Posiva starts its own radiological baseline study in 2016.
Within the latest monitoring programme, an item of "monitoring radiation effects in the
environment" was included. This topic was recognised as very challenging, therefore an
initial task of reviewing radiation effects monitoring approaches was set. This report is
intended to meet the objectives of this task. A review has been undertaken of biomarker
techniques that may be used to detect radiation effects in organisms and to identify
benefits and limitations in their application and ultimately to inform on ways in which
this radiation effects monitoring objective could best be met.
Whilst a number of biomarker techniques are available that allow the effects of
genotoxic pollutants on wildlife to be evaluated, few, if any, can be considered to
specifically measure the effects of radiation exposure. It is also evident that many
biomarker analyses are affected by individual variability in response, and the influence
of other stressors on individuals. This is of particular importance when considering the
application of biomarkers within field monitoring programmes where plants and
animals will be exposed to an array of stressors, both natural and anthropogenic.
In light of the uncertainties associated with biomarker techniques and their application
to field monitoring situations, the addition of biomarker analysis within Posiva’s
environmental monitoring programme around Olkiluoto is not considered proportionate
to the risks posed by any releases that may occur under normal operating conditions. It
is therefore recommended that radiation effects on wildlife be evaluated from biota dose
rates derived on the basis of activity concentrations in environmental media and
comparison of those dose rates against radiation effects data and/or derived screening
criteria that are based on those effects data.
Keywords: Radiation effects, wildlife, monitoring.
SÄTEILYN VAIKUTUSTEN MONITOROINTI YMPÄRISTÖSSÄ
TIIVISTELMÄ
Lounais-Suomessa sijaitsevassa Olkiluodossa on tällä hetkellä toiminnassa kaksi
ydinvoimalaitosta ja kolmas on rakenteilla, lisäksi saarella on käytössä oleva matala- ja
keskiaktiivisen jätteen loppusijoitustila. Olkiluoto on myös valittu käytetyn
ydinpolttoaineen loppusijoituspaikaksi ja omistajiensa käytetyn ydinpolttoaineen
loppusijoituksesta vastaava Posiva Oy toteuttaa alueelle loppusijoitus- ja
kapselointilaitosta vuonna 2015 saadun rakentamisluvan mukaisesti. Posiva on
toteuttanut alueella jo vuosia monitorointiohjelmaa, jonka puitteissa kerätään aineistoa
sijoituspaikan geosfäärin ja ympäristön pitkäaikaisturvallisuuskriittisistä ominaisuuksista ja kehityskuluista, sekä seurataan rakennustöiden ympäristövaikutuksia.
Tämänhetkinen ohjelma on laadittu ajalle ennen loppusijoituslaitoksen käyttövaihetta ja
tullaan päivittämään ennen käyttövaiheen aloittamista. Saarella toimivien voimalaitosten osalta alueella on jo käynnissä säteilyntarkkailuohjelma ennen kuin Posiva
aloittaa oman ympäristön radioaktiivisuuden perustilaselvityksensä vuonna 2016.
Viimeisimmässä monitorointiohjelman päivityksessä mukaan sisällytettiin myös tehtävä
"säteilyn vaikutusten monitorointi ympäristössä". Tämä aihe todettiin erityisen
haastavaksi ja asiaa päätettiin käsitellä toteuttamalla selvitys menetelmistä, joilla
säteilyn aiheuttamia vaikutuksia ympäristössä voitaisiin monitoroida. Tämän raportin
tarkoituksena on täyttää kyseisen tehtävän tavoitteet. Raportissa tehdään katsaus
biomarkkeritekniikoihin joita voitaisiin käyttää havainnoidessa säteilyn vaikutuksia
eliöissä, tunnistetaan menetelmiin liittyviä etuja ja rajoituksia ja lopulta tehdään
yhteenveto siitä miten säteilyn vaikutusten seurannan tavoite parhaiten saavutettaisiin.
Vaikka on olemassa useita biomarkkeritekniikoita, joilla voidaan arvioida genotoksisten
saasteiden vaikutuksia luonnolle ja eläimistölle, vain harvoja niistä voidaan pitää
sellaisena että niillä voitaisiin edes jossain määrin erityisesti mitata säteilyaltistuksen
aiheuttamia vaikutuksia. Moniin biomarkkerianalyyseihin vaikuttavat myös yksilötason
erot vasteessa, kuten myös muut, yksilötasolla vaihtelevaa kuormitusta aiheuttavat
ympäristötekijät. Tämä on erityisen tärkeää huomioida jos biomarkkerimenetelmien
soveltamista käytännön monitorointiohjelmiin harkitaan, koska luonnossa kasvit ja
eläimet ovat jatkuvasti alttiina laajalle kirjolle sekä luonnollisia että ihmisen aiheuttamia
kuormitustekijöitä.
Biomarkkeritekniikoiden käytännön soveltamiseen liittyvien epävarmuuksien valossa,
niiden käyttämistä osana Posivan ympäristön monitorointiohjelmaa ei nähdä
tarpeellisena suhteessa loppusijoituslaitoksen normaaliin käyttöön mahdollisesti
liittyvien radioaktiivisten päästöjen suuruusluokkaan. Täten suositellaan että säteilyn
vaikutuksia luonnolle ja eläimistölle arvioidaan aktiivisuuspitoisuuksiin ympäristön eri
osissa perustuvien annosnopeuksien pohjalta, verraten annosnopeuksia säteilyn
vaikutuksista tehtyihin tutkimuksiin tai niihin perustuviin seulontakriteereihin.
Avainsanat: Säteily, vaikutukset, eläimistö, monitorointi.
1
TABLE OF CONTENTS
ABSTRACT
TIIVISTELMÄ
1
INTRODUCTION ..................................................................................................... 2
1.1
2
AN INTRODUCTION TO BIOMARKERS ................................................................ 4
2.1
3
Report outline ................................................................................................... 3
Indicator species .............................................................................................. 6
RADIATION EFFECT BIOMARKERS ..................................................................... 8
3.1
Mechanisms for radiation damage to organisms.............................................. 8
3.2
Biomarkers of genetic damage......................................................................... 9
3.2.1
Comet assay ............................................................................................. 9
3.2.2
Micronucleus assay................................................................................. 10
3.2.3
Frequency of symmetrical chromosome translocations (double strand
breaks)................................................................................................................... 12
3.2.4
DNA mutation frequencies, repair genes and proteins ........................... 14
3.3
Biomarkers of oxidative damage .................................................................... 15
3.4
General health and reproduction biomarkers ................................................. 16
4 APPLICATION OF BIOMARKER TECHNIQUES TO DETECT RADIATION
EXPOSURE AND/OR EFFECTS FROM ENVIRONMENTAL CONTAMINATION ....... 17
4.1
Biomarker responses to radioactively contaminated environmental media
under laboratory exposure conditions ....................................................................... 17
4.2
Biomarker responses following in situ exposure to radioactivity .................... 18
5 UNCERTAINTIES IN THE APPLICATION OF BIOMARKERS TO EVALUATE
RADIATION EFFECTS ON BIOTA............................................................................... 21
6
DISCUSSION AND CONCLUSION....................................................................... 23
6.1
Summary ........................................................................................................ 24
REFERENCES ............................................................................................................. 25
2
1
INTRODUCTION
In December 2012, Posiva Oy submitted an application for a construction licence for a
spent nuclear fuel repository on Olkiluoto Island, in southwest Finland. The island is
currently home to two operational nuclear power plants, with a third under construction,
and a repository for low and intermediate level waste. The licence submission was
supported by a long-term monitoring programme, undertaken by Posiva, since 2003.
The aims of the monitoring programme have been (Posiva, 2012):

To provide information to underpin the selection of a final disposal site for spent
nuclear fuel and, following selection, to confirm the suitability of the site for
disposal and support the planning and design of the facility.

To observe changes in the host rock during construction activities of the
ONKALO rock characterisation facility and in the surface environment, that may
affect the long-term safety of spent nuclear fuel disposal, or the assessment of it.

To derive data on the properties of the site to support assessment of the longterm impacts of spent fuel disposal and on the environmental impact during
construction and operation of the disposal facility.
Within the overall monitoring programme, a surface monitoring campaign has been
undertaken, which will continue throughout the construction and operation phases. The
surface monitoring programme is focussed on fulfilling the monitoring targets set for
the environmental impacts arising from the extensive construction activities related to
repository construction and operation, especially excavations and rock piling activities.
The environment section also provides data on land-use changes that may possibly
affect the results of other monitoring or research activities ongoing at the site. The
environment section also produces data to support biosphere modelling of the storage
and transport of radionuclides in and between environmental media and biota (Posiva,
2012).
The monitoring programme and associated schedule have been revised on a regular
basis. Within the latest plan for monitoring at Olkiluoto (Posiva 2012), an item of
"monitoring radiation effects in the environment" was included within the
environmental monitoring section of the programme. The item originated from a list of
monitoring activities presented in the 2008 environmental impact assessment report
(Posiva, 2008). This topic was recognised as very challenging, therefore an initial task
of reviewing radiation effects monitoring approaches was set (Posiva, 2012). This report
is intended to meet the objectives of this task, and is based on an understanding of the
current situation at Olkiluoto with regard to radiation sources and construction
activities, and releases of radioactivity that may arise from the spent nuclear fuel
repository as a result of planned activities.
In terms of radiation effects on organisms, two broad approaches can be taken. One
approach is to apply biomarker techniques that provide a measure of the impact of
exposure on organisms at various levels of biological organisation. The alternative
approach is to interpolate effects from knowledge gained from experimental (laboratory
and field) studies on the effects of radiation and measured radioactivity concentrations
3
in the environment. The purpose of this report is to provide a brief review of biomarker
techniques that may be used to detect radiation effects in organisms and to identify
benefits and limitations in their application and ultimately to inform on ways in which
this radiation effects monitoring objective could best be met.
1.1
Report outline
The remainder of this report is structured as follows:

Section 2 provides an introduction to biomarkers and the benefits and limitations
of their application to field studies. Indicator species are also discussed.

Section 3 provides a summary of mechanisms by which radiation can cause
biological damage in organisms and biomarker techniques that can be applied to
evaluate effects at different levels of biological organisation identified. Key
issues in the application and interpretation of these techniques are also
discussed.

Section 4 provides an overview of the application of the principal biomarker
techniques, specifically to evaluate radiation effects, under both laboratory and
field conditions.

Section 5 provides a brief discussion of key uncertainties associated with the
application of biomarker techniques to evaluate radiation effects on the
environment.

Section 6 then discusses the findings of the review, in the context of the current
situation at Olkiluoto and presents an alternative approach for evaluating
radiation effects, consistent with that employed by Posiva for evaluating
radiation effects on people.
The review of biomarker techniques for application to radiation effects monitoring
presented here is not intended to be exhaustive. Rather, the intention has been to
identify possible techniques, including examples that may help illustrate the benefits
and limitations to their application to field monitoring programmes and, from this,
suggest an approach by which Posiva could evaluate the effects of radiation on biota.
4
2
AN INTRODUCTION TO BIOMARKERS
Biomarkers are functional measures of the exposure of an organism (plant or animal) to
stressors and have been defined as (Depledge, 1996):
“biochemical, cellular, physiological or behavioural variations that can be
measured in tissue or body fluid samples or at the level of whole organisms that
provide evidence of exposure to and/or effects of, one or more chemical pollutants
(and/or radiations)”.
The use of biomarkers has been advocated as a useful addition to risk assessment
procedures (Hagger et al., 2006, 2009). Indeed, biomarkers of genetically relevant
damage in organisms have been suggested as a means of addressing uncertainties in
ecological risk assessments for radiation by providing a meaningful measure of
biological damage (Ulsh et al., 2003).
It is widely accepted that there is a continuum of events induced during exposure to
pollutants (Dillon and Lynch, 1981; Depledge and Fossi, 1994; Decaprio, 1997). As a
healthy organism is exposed to an increasing pollutant load, the response of that
individual changes from a normal function, through a reversible phase to an irreversible
phase preceding death. Ideally, biomarkers are aimed at identifying the initial reversible
changes such that measures can be taken to avoid permanent physiological damage.
Biomarkers are thus intended to provide a more sensitive indicator of environmental
stressors than death. There are, however, a number of limitations associated with the
application of biomarker techniques, including (Forbes and Forbes, 1994):

Lack of consistent response;

Specificity of response to a species or taxonomic group;

Seasonal variation;

Low levels of precision; and

Lack of ecological relevance.
Variation in response to a contaminant can also occur depending upon the age and sex
of an individual, the reproductive status of that individual and its nutritional status
(Michelmore & Chipman, 1998).
There are three broad categories of biomarker: exposure, effect and susceptibility
biomarkers (Hagger et al., 2012). Biomarkers of exposure provide qualitative or
quantitative estimates of exposure of an organism to stressors, but the response
measured may not reflect the degree of adverse effects on that organism or on higher
levels of organisation such as communities or populations. Biomarkers of effect are
specifically associated with a toxicants mode of action in causing biological damage and
provide a qualitative link between the degree of biomarker response and the degree of
adverse effects. Biomarkers of susceptibility indicate inherent or acquired ability of an
organism to respond to the challenge of exposure to a specific toxicant.
5
In a risk assessment context, there can be considerable challenges in linking observed
biomarker responses to individual pollutants within a multi-stressor environment
(Mitchelmore & Chipman, 1998). Such stressors could relate to anthropogenic
substances in the environment (i.e. chemical contaminants) or result from natural
environmental stressors such as competition for food and habitats, presence of
predators, or environmental variables. Environmental conditions such as pH, soil
moisture content or temperature can affect exposure (van Gestel, 1997). Seasonality can
also affect an organism’s response to environmental contaminants, for example, as a
result of changes in diet and hormonal status (Mitchelmore & Chipman, 1998). Natural
ecosystems are therefore complex with multiple interactions occurring both between
organisms within communities and between those organisms and their environment.
In order to apply biomarkers in a risk assessment context, knowledge is required as to
the normal limits of an organism’s response in order to ascertain the onset of
anthropogenic stress. Where multiple contaminants (or additional environmental
stressors) are present in the environment, there is the potential for stressor interactions
to occur resulting in additive, antagonistic or synergistic effects on an organism.
Furthermore, the presence of some contaminants can induce effects in organisms similar
to those induced by radiation exposure. For example, metals and organic contaminants
can induce free radicals, resulting in genetic instability and mutation (Salbu, 2009) with
such responses occurring in the absence of radiation. The activation of some enzymes
can also give rise to genotoxic impacts similar to those occurring as a result of exposure
to genotoxic agents; heat stress has also been found to induce DNA damage in fish
Mitchelmore & Chipman, 1998).
The potential for multiple environmental stressors to induce similar effects in organisms
complicates the interpretation of any detectable biomarker response observed under
field conditions. A positive response in a biomarker assay may not therefore indicate
radiation exposure, unless that assay is induced solely as a result of exposure to ionising
radiation. Differentiating between radiation impacts caused by anthropogenic sources of
radiation and natural radiation exposure further complicates effects analysis
(Mitchelmore & Chipman, 1998). For example, ultraviolet (UV) radiation has been
shown to alter oxidative status through reactive oxygen species (ROS) production and
increase stress protein responses in crustaceans, at exposure levels consistent with
natural solar radiation (Hollmann et al., 2015). Natural UV irradiation can also increase
the toxicity of contaminants in the environment (Kim et al., 2009).
The non-specificity of many biomarker techniques can limit their application in the
screening of chemicals and potentially polluted environments. However, the high
sensitivity of many techniques can also be considered favourable in environmental
monitoring applications, but the causes of any observed effect should be investigated,
including potential variability as a result of diet, temperature and other environmental
stressors as well as variations linked to individuals (age, sex and reproductive status)
(Mitchelmore & Chipman, 1998). It should also be noted that some chemicals can act as
inhibitors, potentially reducing an organisms biomarker response to a given
concentration (or activity) of a contaminant (Mitchelmore & Chipman, 1998).
There is rarely a unique single signature for a stress response to an environmental
contaminant. As such, a suite of biomarkers, operating at different levels of biological
6
organisation (e.g. sub-cellular, cellular, whole organism), is often adopted in a weight of
evidence approach and to achieve a more robust evaluation of the hazard (Hagger et al.,
2006, 2012). Such an approach allows for the discrimination between clean/healthy and
polluted/unhealthy sites, but may not provide a conclusion as to the cause of any
observed effects, whether due to radiation or other environmental stressors.
2.1
Indicator species
It is not feasible to monitor all species that may be present in an ecosystem for the
effects from contaminant exposure. Indicator species are therefore commonly selected
that provide a measure of exposure and/or effect to environmental pollutants. However,
the selection of one or two species may over or underestimate the potential for
contaminant exposure and/or effects, depending upon the ecological characteristics of
the species such as habitat and diet (Smith et al., 2007). Hagger et al. (2012) suggests
the selection of species occupying key trophic positions, in order to evaluate the
integrity of an ecosystem as a whole.
The choice of indicator species will be determined by the communities of species
present at a site of interest, nonetheless there are particular traits that can help guide
selection (Forbes & Forbes, 1994):

Sedentary nature to minimise uncertainty associated with variability in
magnitude, duration and extent of exposure and to allow effects to be linked to a
particular site.

Naturally resident in the area of interest.

Habitat occupancy and feeding habits that maximise exposure.

Abundant to ensure impacts do not occur as a result of sampling within the
population and to allow sampling at various locations.

Of a large enough size to allow tissue sampling and analysis.

Sensitive to the effects from the pollutant (or pollutants) of interest.

Easy to sample.
Indicator species selection may also be driven by the biomarker analyses to be
undertaken, for example due to particular cell-type requirements of a technique or due
to a technique having been developed and optimised for particular species (or types of
species). Blood haemocytes and erythrocytes are closely exposed to environmental
agents through their role in the transport of toxicants within the body of an organism
and in various defence mechanisms (Lacaze et al., 2010). Blood can be harvested nondestructively from many species and a number of biomarker techniques have therefore
been developed for use on blood samples.
In aquatic environments, bivalve molluscs, such as mussels are commonly used as test
species. Many species are sedentary and inhabit sediments that commonly accumulate
7
pollutants. The filter feeding habits of mussels also exposes them to large volumes of
water and suspended sediments. Furthermore, many species are considered to be
broadly tolerant of changes in biotic and abiotic parameters (Franzellitti et al. 2010) and
are easily accessible and often available in sufficient numbers to support monitoring
programmes (Ericson et al., 2002). Haemocytes are also relatively easy to sample from
the adductor muscle of molluscs.
Fish blood cells consist of around 97% erythrocytes, and therefore provide a relatively
homogenous biological test material, leading to the selection of fish as indicator species in
some monitoring programmes (Mitchelmore & Chipman, 1998). Worms are also
commonly used as bioindicator species (Wilding et al. 2006; Hagger et al., 2012),
providing test organisms both in aquatic and terrestrial environments. Indeed, earthworms
have been suggested as being one of the most suitable representatives of soil animals used
for the assessment of soil pollution (Misra et al. 2005; Lourenço et al., 2012).
8
3
RADIATION EFFECT BIOMARKERS
Organisms are continuously exposed to ionising radiation from natural background
sources, including solar (UV) radiation with anthropogenic sources such as the nuclear
industry, hospitals and research facilities potentially increasing exposure. Radiation, both
natural and anthropogenic, is associated with a number of deleterious effects. Effects can
include production of ROS, DNA mutation and immunosuppression (Hollmann et al.,
2015). It is important to note, however, that studies on effects in organisms provide an
integrated analysis of the impact of all mutagenic activity, which may result from
exposure to radiation and/or genotoxic chemicals (Hebert & Luiker, 1996).
The following section provides an overview of the mechanisms for radiation damage in
organisms and of some of the more common biomarker techniques that have been
applied to evaluate exposure and/or effect of genotoxic pollutants, including radiation,
on organisms. Examples of the application of the biomarker techniques to detect
genotoxic damage, including as a result of exposure to radiation, are given. Examples of
the technique being used to evaluate genotoxicity from in situ radiation exposure are
discussed in Section 4.
3.1
Mechanisms for radiation damage to organisms
Ionising radiation can cause biological harm through both direct and indirect damage.
Direct damage occurs as a result of radiation affecting atoms within key molecular
structures such as DNA (Hagger et al., 2012) and may give rise to chromosome
aberrations such as sister chromatid exchange1 or single / double strand breaks, gene
mutations and/or cell death (Wright & Coates, 2006). Indirect damage occurs when
radiation interacts with a cells cytoplasm, splitting water molecules which results in the
formation of toxic substances and free radicals that can harm a cells organelles and
affect cellular function (Hagger et al., 2012; Salbu, 2009). The major adverse
consequences of radiation exposures are attributed to DNA damage in irradiated cells
that has not been correctly repaired by metabolic processes (Coates et al., 2004).
Radiation-induced bystander effects, where non-irradiated cells respond as if irradiated,
can also occur (Coates et al., 2004). Such effects occur as a result of non-irradiated cells
being in close proximity to irradiated cells and receiving stress signals from those cells
(Wright & Coates, 2006; Ilnytskyy & Kovalchuk, 2011). Bystander effects that may be
exhibited include altered production of stress proteins or free radicals; cell death,
proliferation or differentiation; and/or the induction of mutations and chromosome
aberrations, among others (Coates et al., 2004). Protective adaptive responses may also
occur whereby bystander cells subsequently exposed to radiation are more radioresistant than those that have not been subject to bystander signals (Ilnytskyy &
Kovalchuk, 2011). Both direct and indirect effects of radiation exposure can persist
through multiple cell generations (Wright & Coates, 2006).
Whilst the bystander effect can either promote greater radio-resistance or increased
radio-sensitivity upon subsequent exposure of previously non-irradiated cells, it should
1
The exchange of genetic material between two identical sister chromatids
9
be born in mind that other contaminants can have a similar effect by sensitising cells
(Mothersill & Seymour, 2010). Non-contaminant stressors can also affect cells in a
similar way. For example, prior heat stress has been shown to affect the response to
radiation (Mitchel et al., 1999). Such sensitising agents may be present at the same time
as ionising radiation exposure or may have occurred at a time prior to radiation
exposure. The response of an organisms cells (and consequently higher levels of
biological organisation) to ionising radiation in an environmental context is not,
therefore, straightforward. Even if radiation exposure were to occur in isolation from
other stressors (natural or anthropogenic), both positive (adaptive) responses and
negative effects (e.g. genetic damage) could occur, depending upon the exposure
conditions.
Information of biological effects in non-human organisms resulting from low-dose
chronic exposure radiation is limited (Copplestone et al., 2008; Salbu, 2009). Of the
information available on chronic effects, most is focussed on fish, mammals and
terrestrial plants (Salbu, 2009). Ionising radiation is associated with the induction of
double strand breaks and subsequent chromosome exchange aberrations, which have
been suggested to be the most detrimental of DNA effects and the most resistant to
repair (Ulsh et al., 2003; Hagger et al., 2012). Single strand breaks can also occur, but
are more readily repaired. Furthermore, single strand breaks can be induced by a wide
range of toxic agents, including heavy metals and pesticides whereas double strand
breaks are more commonly associated with radiation induced damage (Ulsh et al.,
2003). With repair mechanisms being limited, accumulation of strand breaks can occur
as a result of chronic exposure to radiation. The measurement of strand breaks and
chromosome aberrations have therefore been proposed as potentially sensitive
biomarkers of radiation exposure (Ulsh et al., 2003). The application of biomarkers in
environmental studies has however been limited; the techniques required for analysis
are both specialist and costly (Ulsh et al., 2003).
3.2
3.2.1
Biomarkers of genetic damage
Comet assay
The comet assay, otherwise known as the single-cell gel electrophoresis assay, is a rapid
and sensitive method of identifying DNA damage, repair and cell death in different cell
types (Klobučar et al. 2003; Hagger et al., 2012) and has been widely applied in
genotoxicity studies (Rank & Jensen, 2003). Unlike many genotoxicity assays, the
comet assay does not require cells with large chromosomes and/or cells that are
mitotically active2; cells from any tissue can be used so long as nuclei are present
(Ralph & Petras, 1998; Wilson et al., 1998; Pavlica et al., 2001). The technique is used
to identify single or double strand breaks in DNA, with double strand breaks being
considered to be more biologically relevant as they can lead to more persistent damage
(Hagger et al., 2012). In the majority of published articles on the application of the
comet assay, animal species have been used as bioindicator organisms, but can also be
applied on plant tissues (Ventura et al., 2013).
2
Cells are undergoing division of the nucleus whereby chromosomes are separated into two
identical sets of chromosomes that each end up forming their own nucleus.
10
The comet assay uses an electric field to draw negatively charged fragments of DNA
through a gel with the extent of migration (a comet ‘tail’) being directly related to the
degree of DNA damage within cells; a greater tail indicates a greater extent of DNA
damage. A single-cell suspension is required as the basis to the technique. Two versions
of the assay can be performed. The alkaline comet assay is used to detect single-strand
breaks, a neutral comet assay can be used to detect double-strand breaks. The assays are
able to detect breakages as low as one break per chromosome within an individual cell
(Michelmore & Chipman, 1998).
The comet assay has been used in a number of studies to evaluate the effect of various
anthropogenic contaminants on a range of biota, both aquatic and terrestrial. For
example, Ralph & Petras (1998) tested amphibian tadpoles as an indicator species for
genotoxicity of freshwaters through transplantation of caged tadpoles to a number of
potentially contaminated water bodies. Impacts were analysed using the alkaline
(single-strand break) comet assay, which was found to be sensitive in detecting different
degrees of pollution associated with the water bodies in which they were placed.
Freshwater mussels have also been used as indicator species, with individuals being
transplanted downstream of municipal wastewater outlets (Pavlica et al., 2001). A
significant increase in DNA damage in haemocytes was observed in transplanted
mussels relative to a control population. It was noted, however, that appreciable DNA
damage was observed in some cells derived from the control population, indicating the
individual variability that can be observed. Klobučar et al. (2003) also applied the
comet assay under field conditions to evaluate levels of genotoxicity in different
polluted sites using transplanted freshwater mussels.
In terms of radiation effects evaluation, Kumar et al. (2014) looked at the comet assay
response in mussel haemocytes following acute single dose gamma irradiation (2 to 10
Gy) of mussels. A dose dependent response was observed with effects being observed
in all exposure groups with response reducing over time. The maximum response was
observed 24 hours post-exposure, reducing through 48 and 72 hours. The reduction in
observed response was considered to relate to either cellular repair mechanisms or to the
loss of cells due to cell death. The comet assay has also been applied within a suite of
biomarkers to evaluate genotoxic impacts in a bat population inhabiting an abandoned
thorium mine (Meehan et al., 2004). Examples of the technique being used to evaluate
genotoxicity from in situ radiation exposure are discussed in Section 4.
Whilst the comet assay has been applied to evaluate of radiation damage in organisms,
the technique is a non-specific biomarker of genotoxicity, responding to a wide range of
environmental contaminants (Mitchelmore & Chipman, 1998) and different forms of
environmental stress (Klobučar et al., 2003). The sensitivity of the assay can be affected
by individual variability in factors such as reproductive status, sex and age (AlAmri et
al., 2012). Even in unexposed groups, individual variability can be high (Wilson et al.,
1998), which may reduce the overall sensitivity of the assay to detecting pollutantinduced genotoxic damage.
3.2.2
Micronucleus assay
Micronuclei are masses of DNA that resemble small nuclei located in the cytoplasm of
cells, rather than being within the nuclear membrane (Hebert & Luiker, 1996).
11
Micronuclei form during cell division in cells that have been subject to exposure to
chemical mutagens, including radiation, resulting in chromosome breaks or spindle
dysfunction and a bi-nucleated appearance in cells that have undergone cellular division
(Figure 3-1). Micronuclei can be detected using light microscopy, with the number of
cells containing micronuclei within a cell population being related to the extent of
cytogenetic damage (i.e. damage to genetic material within a cell).
Use of the micronucleus assay is now common place in the field of genetic toxicology
(Hagger et al., 2012; AlAlmari et al., 2012). The assay is sensitive and rapid and can be
performed across a wide range of terrestrial and aquatic organisms with little formal
training (Hagger et al., 2012). The technique is however limited to detecting damage as
a result of chromosomal breakage or lagging chromosomes during cell division. Other
forms of damage, such as sister chromatid exchanges, are not detectable. Furthermore,
the technique is not specific to the effects of radiation and, like the comet assay, results
can be influenced by seasonal, sex and age variations and the health and nutritional
status of individuals (Hebert & Luiker, 1996; AlAmri et al., 2012). The effect of season
alone can be considerable (Hebert & Luiker, 1996).
The micronucleus assay has been used in a wide range of studies, both in the laboratory
and the field to assess the cytogenetic impact of exposure of different organisms to
pollution. For example:

Jaeschke et al. (2011) studied the incorporation and effects of different forms of
tritium in marine mussels, with the micronucleus assay being used with
haemolymph samples to gauge the level of genotoxicity.
Figure 3-1. Formation of micronuclei during cell division (adapted from Hagger et
al., 2012).
12

Anbumani & Mohankumar (2012) exposed fish to gamma radiation (5 Gy
delivered at a dose rate of 0.002 Gy/min) and micronuclei formation was
measured in erythrocytes in the days following exposure. A significant increase
in micronuclei was recorded from the third day, post-exposure. The exposure,
whilst protracted, was nonetheless high and much greater than would be
expected to occur in the environment under most circumstances.

Klobučar et al. (2003) transplanted mussels from a clean site to polluted sites
and applied the micronucleus assay in combination with the comet assay to
identify genotoxicity, again in haemolymph samples. The formation of
micronuclei was linked to the level of pollution although differences were not
observed in all polluted sites relative to the control.
3.2.3
Frequency of symmetrical chromosome translocations (double strand
breaks)
DNA double strand breaks are considered to be one of the most critical effects
following the exposure of organisms to ionising radiation, potentially resulting in cell
death and/or genomic instability (Gerić et al., 2014).
The frequency of symmetrical chromosome translocations (Figure 3-2) in peripheral
blood lymphocytes has been proposed as a biomarker of cumulative radiation exposure,
due to its relative specificity to radiation exposure and the potential for non-lethal
sampling of lymphocytes (Ulsh et al., 2003). As noted previously, whilst other
genotoxic agents can also induce strand breaks, they are more commonly associated
with single strand breaks as compared with radiation-induced double strand breaks
(Ulsh et al., 2003). Nonetheless, double strand breaks can occur as a result of exposure
to non-radiation stressors, with cancer-treating pharmaceuticals, routinely entering the
environment from hospital discharges and patient excreta, being associated with such
genotoxic effects (Gerić et al., 2014).
The background frequency of double strand breaks in humans has been shown to
accumulate with age. The incidence of chromosome aberrations in lymphocytes has
therefore been suggested as a means of determining the extent of unknown exposure to
radiation in people, and also in biota inhabiting radionuclide-contaminated
environments (Ulsh et al., 2003). The dose rate over the period of exposure does
however affect the frequency of chromosome aberrations for a given total dose rate
(Ulsh et al., 2003), which complicates the interpretations of biomarker response.
13
Figure 3-2. Formation of double strand breaks (note, only viable options are
illustrated). Adapted from Ulsh et al. (2003).
Double strand breaks can be detected by labelling individual pairs of homologous
chromosomes3 in unique colours. Fragments from labelled chromosomes that are
translocated to other non-labelled chromosomes can then be detected using a fluorescent
microscope. The approach to labelling chromosomes (fluorescence in situ hybridisation,
FISH) is not however a simple process, requiring the denaturing and reannealing4 of
DNA. It is also possible for all chromosomes within a cell to be labelled in different
colours for maximum sensitivity in detecting aberrations throughout the entire
chromosome population of a cell through a technique known as multiplex FISH
(mFISH) (Ulsh et al., 2003). Whilst potentially a sensitive biomarker for radiation
induced damage in cells, detection of double strand breaks using FISH is costly and
requires specialist training in cytogenetics and chromosome microdissection in order to
construct the appropriate probes/labels (Ulsh et al., 2003). It is also necessary to
identify suitable target species, i.e. those with a few large and easily recognisable
chromosomes. Species should also have relatively long life-spans to allow the
accumulation of aberrations, thus allowing measurable responses under chronic, lowlevel radiation exposure. Ulsh et al. (2003) suggest a suite of suitable species be
selected for biological monitoring applications.
3
Maternal and paternal chromosomes that pair up during meiosis.
4
The process by which two single strands of DNA combine to form double-stranded DNA.
14
An alternative approach to detecting double strand breaks is the γ-H2AX foci assay.
H2AX is a gene that contributes to the structure of DNA. Double strand breaks in
chromatin initiate the phosphorylation of the gene resulting in γ-H2AX, which is an
indicator for the induction of cellular double strand break repair mechanisms (Kinner et
al., 2008). The regions containing γ-H2AX can then be detected using phospho-specific
antibodies and immunofluorescence microscopy and serve as markers of the induction
and repair of double strand breaks (Kinner et al., 2008). Phosphorylated H2AX can be
detected after only a few minutes in cells exposed to ionising radiation, with the
occurrence of γ-H2AX reaching a maximum around thirty minutes later, and each
occurrence indicating a double strand break (Kinner et al., 2008). The extent to which γH2AX can be detected following induction of double strand breaks and their subsequent
repair is not, however, clear; the assay may therefore be most appropriate for detecting
current effects rather than as a marker for long-term chronic exposure to genotoxic
stressors such as radiation.
As with the FISH assay, responses to non-radioactive environmental contaminants have
been reported (Gerić et al., 2014). The assay has also been shown to have a lower
detection limit than the likes of the comet and micronuclei assays (Gerić et al., 2014). As
noted previously, the neutral comet assay can be employed to detect double strand breaks.
3.2.4
DNA mutation frequencies, repair genes and proteins
Upon exposure to ionising radiation, DNA damage response pathways are activated and
key proteins involved in these pathways can be monitored (AlAmri et al., 2012).
RAD51 is one such protein. RAD51 assists in the repair of DNA double strand breaks,
but not single strand breaks and is, hence, more specific to radiation exposure than
proteins associated with single strand break repair mechanisms (AlAmri et al., 2012).
The approach requires the isolation of the RAD51 DNA sequence in test species and
quantitative analysis of its mitochondrial RNA expression, using polymerase chain
reaction (PCR)5.
Additional DNA repair proteins and genes have been used as biomarkers of radiationinduced damage in organisms. One such gene is p53, which plays an essential role in
regulatory pathways of DNA repair (Rhee et al., 2013). Studies on the regulation of p53
in larvae of fish have however indicated that this is not a particularly sensitive
biomarker of radiation exposure, with altered regulation of p53 only being observed in
individuals exposed to 8 Gy or more of gamma irradiation during acute exposure
conditions (Rhee et al., 2013). Rhee et al. (2013) also studied the RAD51 response in
the same study. The RAD51 response was found to be the most sensitive marker of
gamma radiation exposure, being stimulated at a dose of 4 Gy.
The frequency of mitochondrial DNA mutations has also been investigated as a potential
biomarker for radiation exposure. Under laboratory conditions, Wilding et al. (2006)
exposed adult worms to chronic (55 day) gamma irradiation at a range of dose rates (0,
5
PCR is a molecular biology technology used to amplify a single copy or a few copies of a piece
of DNA across several orders of magnitude to create thousands of copies of the selected DNA
sequence. PCR is a specialist technique, requiring sequence data for the genetic material of
interest to allow primers to be developed for the target protein, gene or DNA fragment.
15
1.4 and 8.5 mGy/h). Mutations in mitochondrial DNA were then analysed, using PCR. At
the higher dose rate, a significant increase in mitochondrial DNA mutation frequency was
observed relative to the control, but no increase was observed following exposure to a
dose rate of 1.4 mGy/h relative to the control. Whilst the technique detected mutation
frequency following chronic exposure to radiation, it was concluded that it was not
sufficiently sensitive at the dose rates likely to be encountered in the environment to
promote its use as a biomarker of radiation-induced effects (Wilding et al., 2006).
DNA mutation frequencies, repair genes and protein markers are commonly applied in
combination with biomarkers of oxidative damage (see for example Rhee et al., 2013).
3.3
Biomarkers of oxidative damage
When organisms are exposed to ionising radiation, free radicals and ROS are induced
that can have detrimental cellular effects, such as damage to proteins and nucleic acids,
reduction in membrane integrity and potentially cell death (Salbu & Skipperud, 2006).
The presence of ROS can therefore be measured as an indicator of exposure to
radiation, but also to other contaminants such as heavy metals and organics (Misra et al,
2005; Salbu & Skipperud, 2006; Salbu, 2009; Kaddissi et al., 2012; Won & Lee, 2014).
Cellular responses to mitigate against the damage caused by increased oxidative stress,
such as the production of superoxide dismutase, catalase and oxidised glutathione can
also be measured as a means of identifying the effects of low dose radiation (Salbu &
Skipperud, 2006; Salbu, 2009; Kaddissi et al., 2012). It is again important to note that
these responses are not specific to radiation exposure, but can rather be triggered by a
range of environmental contaminants as well as natural environmental stressors such as
UV radiation exposure (Misra et al., 2005; Krapp et al., 2009; Dahms & Lee, 2010). For
example, the antioxidant capacity of arctic marine-sea ice amphipods (Gammarus
wilkitzkii) was shown by Krapp et al. (2009) to be markedly increased during summer
months as compared to winter. This, it was suggested, may result from a higher level of
oxidative stress being experienced in the summer when individuals are exposed to UV
radiation.
Kaddissi et al. (2012) undertook a study of the relative chemical toxicity and
radiotoxicity of uranium (as U-233 and depleted uranium) on oxidative stress
biomarkers in crayfish. Results indicated that glutathione S-transferase activity was
induced as a result of the radiotoxicity of uranium rather than the chemical toxicity
whereas a range of other biomarkers were either indiscriminate or were more closely
linked with uranium chemical toxicity. Whilst these results may suggest that glutathione
S-transferase could provide a useful biomarker of exposure to low dose exposure to
ionising radiation, this biomarker can also be triggered by a range of other
environmental contaminants, including polychlorinated biphenyl (PCB) (Pérez-López et
al., 2002) and mixed industrial effluents (Ahmad et al., 2000).
Won & Lee (2014) also investigated glutathione S-transferase response to acute gamma
radiation exposure of copepods. A range of other oxidative stress markers were also
evaluated, including the production of antioxidant enzymes such as superoxide
dismutase, catalase and glutathione reductase. More general health indicators were also
evaluated (see section 3.2.6 below). Superoxide dismutase, glutathione reductase and
16
catalase were found to increase in concentration in a dose-dependent manner.
Glutathione-s-transferase was less responsive. The exposures experienced by the test
population (10 to 100 Gy) were however substantially greater than those that are likely
to occur under normal situations in the environment.
Si et al. (2013) also studied the effects of radiation on antioxidant production. Zebrafish
embryos were exposed to carbon ion irradiation (1, 3 and 7 Gy) and a range of
antioxidants (superoxide dismutase, glutathione and catalase) measured. In this
instance, antioxidant mechanisms were reduced following irradiation. A range of
antioxidants were also analysed for in larvae of the fish Kryptolebias marmoratus,
following gamma irradiation. Acute exposures were again employed (1 – 10 Gy) and
antioxidant generation was observed only in larvae exposed to doses of 8 Gy and above.
3.4
General health and reproduction biomarkers
Reproduction is one of the key endpoints of concern with regard to the exposure of
organisms to ionising radiation. As such, a number of researchers have studied the
effects of exposure to ionising radiation on reproduction endpoints in biota. For
example, Zeman et al. (2008) studied the effects of uranium on the number of neonates
produced per female daphnid following 21 days of exposure to uranium concentrations
ranging from 25 to 100 µg/L (relating to internal dose rates calculated at 2.1 to 13
µGy/h at 21 days of exposure). Egg dry mass was also recorded. A significant decrease
in egg dry mass was reported at uranium concentrations above 25 µG/L and an EC10
(the concentration causing a 10% effect across the study population) was calculated at
14 µg/L for the neonate production endpoint. Whilst dose rate was calculated in the
study, the effects observed were linked by the authors to the chemotoxicity of uranium
rather than the radiotoxicity, which was supported by other literature data.
Reproduction was also an endpoint studied by Won & Lee (2014) in copepods exposed
to gamma radiation, with the focus being on egg production. Oxidative stress (described
above), mortality and growth retardation were also studied. Copepods were acutely
exposed to radiation (10 to 100 Gy) and days until egg sacs were produced following
radiation exposure reported. In non-exposed individuals, egg sacs were produced after 1
day, whereas in those exposed to 100 Gy of gamma radiation, the average was extended
to 4 days. The production rate of eggs was also significantly affected with an 80%
reduction in egg production being observed in individuals exposed to 100 Gy as
compared to controls. The ability of nauplii to grow to the copepod stage was also
impaired, with an increase of around 1 day being observed at a dose of 20 Gy and
significant retardation and death being observed in exposures greater than 30 Gy.
Whilst reproductive biomarkers have been shown to be responsive to radiation exposure,
and have a high degree of ecological relevance, measurement of reproductive effects
under field conditions can be complex. Exposure of individuals to contaminated media
obtained from the field within a laboratory or controlled field setting could however be
feasible. For example, Lourenço et al. (2012) exposed worms to contaminated soils by
placing the soils of interest into containers that were kept at the site from which soils were
obtained. To avoid issues with resilience and tolerance etc. cultured worms were added to
each soil. The approach taken allowed worms to be retained within a controlled area,
allowing juvenile production to be evaluated as a marker of reproductive capacity.
17
4
APPLICATION OF BIOMARKER TECHNIQUES TO DETECT RADIATION
EXPOSURE
AND/OR
EFFECTS
FROM
ENVIRONMENTAL
CONTAMINATION
Biomarker techniques are often employed to evaluate biological responses to
contaminants under carefully controlled laboratory conditions, and largely in response
to single stressor exposures. Whilst this approach can be useful in linking biological
effects to specific contaminants, responses under natural field conditions are likely to
vary considerably as compared to those in the laboratory due to the presence of
additional stressors, both natural and anthropogenic. Soil conditions, affecting the
bioavailability of contaminants and, hence, exposure, will also vary under field
conditions where factors such as temperature and moisture are not controlled. This
section therefore reviews studies where biomarker techniques have been employed to
evaluate the effects of ionising radiation under field conditions. Two general approaches
to evaluating effects are reported: collection of contaminated environmental media from
the field and testing of biomarker responses under laboratory conditions; and in situ
exposure of biota and subsequent biomarker analysis. In all the examples identified, a
suite of biomarker techniques were employed.
4.1
Biomarker responses to radioactively contaminated environmental
media under laboratory exposure conditions
Two examples of biomarker techniques being employed to evaluate the effects of
environmental contamination have been identified whereby contaminated media were
collected from the field and indicator species exposed to the sediments within a
laboratory setting.
Laurenço et al. (2011) studied a suite of biomarker responses in earthworms exposed to
contaminated soil from an abandoned uranium mining site that was known to be
contaminated both with radionuclides and metals. A standard reference soil was used as
a control. Biomarkers employed included the comet assay and flow cytometry (to detect
the frequency of cell populations in blood samples and, hence, indicate
immunotoxicity). Metal and radionuclide body burdens were also analysed. Results
showed that worms exposed to the contaminated soil had suppressed populations of
cells responsible for immune responses and increased populations of those responsible
for metabolic and regulatory processes. Genotoxicity, measured using the comet assay,
was also significantly higher in the population of worms exposed to the contaminated
soil. Whilst the study showed the relevance of using multiple biomarker techniques to
identify effects of the contaminated soil on earthworms, it was not intended to identify
the causal factor; rather, a combined response from mixed contamination was intended.
The results therefore indicate how useful the techniques employed can be in identifying
sub-cellular impacts of contaminants. Additional analysis would be required however to
link the observed effects to one or other contaminant present.
A study has also recently been undertaken in the UK (Hagger et al., 2012). A previous
study (Hagger et al., 2008) had shown that significant increases in genetic damage were
evident in a range of species inhabiting estuaries that were known to have a range of
contaminants present, including heavy metals, petroleum-derived organics and
pesticides. Due to the range of contaminants present in the estuaries, the causative
18
agents could not be ascertained. A follow on study (Hagger et al., 2012) was therefore
undertaken to investigate whether biomarker techniques could determine which
contaminants were causing damage. The estuaries investigated were:

Ravenglass Estuary in northwest England, UK, associated with elevated
concentrations of radionuclides as a result of discharges from the Sellafield
reprocessing plant; and

Ribble Estuary, also in northwest England, UK, that receives discharges from
the Springfield’s nuclear fuel production site and is associated with a wide range
of non-radioactive contaminants, including tri-butyl tin (TBT) and heavy metals
(Hagger et al., 2012).
A site in eastern Scotland was also investigated to provide a control exposure.
Lugworms (Arenicola marina) were exposed over a ten day period to sediment that had
been collected from each site. Coelomic fluid was then harvested for micronuclei and
comet assay evaluation. Results of the comet assay indicated that worms exposed to
sediment from Ravenglass Estuary had significantly more DNA damage than those
exposed to Ribble Estuary or control site sediments. No significant difference was
observed in results of the micronucleus assay between worms exposed to the different
sediments. Multivariate analysis was employed to investigate clusters in the responses
at each site and clear differences were observed between all three sites. Relationships
between the biomarker responses and contaminants present in Ribble Estuary sediments
were not readily distinguishable due to the mix of contaminants present. A relationship
was however observed for Ravenglass Estuary sediments between plutonium
concentration and biomarker responses. No relationship was demonstrated with other
radionuclides present at the site. The study concluded that chemical and radionuclide
analysis is vital at sites that are to be subject to biological monitoring to provide
tentative identification of associations between observed effects and pollutants present.
It was further concluded that further research is required to establish the relationships
between biomarkers responses and contaminants giving rise to exposure.
4.2
Biomarker responses following in situ exposure to radioactivity
Of the studies identified below that have evaluated radiation effects on populations
exposed to radiation under field conditions, the majority relate to uranium or thorium
mining situations. The dose rates / activity concentrations associated with these
exposure situations are likely to be much higher than those likely to be encountered in
the environment from exposures resulting from planned activities.
Laurenço et al. (2012) evaluated genotoxicity and cytotoxicity endpoints in earthworms
that had been exposed in situ to uranium mining wastes at an abandoned mine in
Portugal. Soil contamination has occurred as a result of mine tailings and from the
application of sludge from a residual water treatment pond that is periodically removed
and spread on the land. The sludge contained high levels of heavy metals and
radionuclides from the uranium decay chain and low levels of radionuclides from the
thorium decay chain. A suite of biomarker techniques was employed to evaluate effects
of the soil contamination on worms, including comet, flow cytometry and reproduction
assays. Exposure to the contaminated soil resulted in complete inhibition of worm
19
cocoon production with no juveniles being detected whereas the rate of production of
juveniles in controls was at least 30%. Significant detriment was also reported for
coelomocytes (free cells) of worms exposed to the contaminated soils, as measured by
the comet assay. Flow cytometry, used to evaluate cell frequency and cell proliferation,
indicated effects on the immune function of exposed worms. DNA content of cells was
also affected, with the impact increasing with exposure duration. All biomarkers
therefore showed a positive response in worms exposed to the contaminated soil.
In a second study, gene expression was also evaluated in worms exposed to the
contaminated soil (Laurenço et al., 2013a). Results indicating that the expression of
some genes involved in key physiological functions associated with the response to
oxidative stress were altered. Wood mice inhabiting the area were also captured and p53
gene expression analysed in combination with genotoxicity, as measured by the comet
assay (Laurenço et al., 2013b). Results again indicated a loss of DNA integrity. The
sensitivity of p53 gene expression was variable depending upon the tissues analysed.
The causative agents for the effects observed were not investigated and it is not
therefore possible to determine whether effects occurred as a result of radiation
exposure and/or chemical toxicity.
Meehan et al. (2004) studied DNA damage in a bat population inhabiting an abandoned
thorium mine in South Africa. The bats were known to hibernate in the mine during
winter and were thus exposed continuously during this time to relatively low dose
radiation. The bats also roosted in the mine throughout the year. External dose rates in
two chambers of the mine were measured at 20 µGy/h and 100 µGy/h. A control
population (external dose rate < 2 µGy/h) was also evaluated. Bats from each location
were sampled and peripheral blood extracted for analysis using both the comet and
micronucleus assays. Results of the micronucleus assay indicated that micronuclei were
present in all exposure groups, with the frequency of occurrence increasing with dose
rate. Results of the comet assay also showed dose-dependent DNA damage, with
statistically significant differences being observed between all exposure groups. The
study concluded that both assays were sensitive enough to detect DNA damage induced
by prolonged exposure to low dose radiation, although dose rates were notably higher
than would be expected from planned exposure situations.
AlAmri et al. (2012) applied both the RAD51 analysis and comet assay to evaluate
effects of contaminated sediments in the Ravenglass Estuary on marine mussels. As
noted previously, the estuary is subject to historical and current input of radioactivity
from the Sellafield reprocessing plant. Effects of in situ exposure of mussels to
contamination in the estuary was evaluated using the comet and RAD51 assays.
Elevated RAD51 was recorded in mussels sampled from the estuary compared with a
control site. The increased RAD51 response in the estuary, compared with the reference
site, was considered to have arisen from the presence of alpha radionuclides, known to
be efficient in inducing DNA double strand breaks. DNA damage within mussel
haemocytes, as measured using the comet assay, was also significantly higher, but could
not be directly linked to ionising radiation exposure due to the presence of other
contaminants such as heavy metals and organics within the estuary.
In addition to the examples provided above, it should be noted that a UK funded project
(TREE) is currently underway to evaluate radiation effects in populations of animals
20
inhabiting the Chernobyl exclusion zone (see www.ceh.ac.uk/tree). Techniques that will
be employed during this study include single and double strand comet assay,
chromosome aberration measures, ROS, anti-oxidant capacity and micronucleus assay.
21
5
UNCERTAINTIES IN THE APPLICATION OF
EVALUATE RADIATION EFFECTS ON BIOTA
BIOMARKERS
TO
Whilst much effort has been made in recent years to evaluate the potential effects of
radiation on wildlife, there remain considerable knowledge gaps. This is particularly the
case with regard to low dose, chronic exposures typical in the environment of organisms
influenced by radioactive releases (Garnier-Laplace et al., 2003).
No single analytical approach can provide an appraisal of the range of effects that may
arise with exposure to contaminants (Hebert & Luiker, 1996). Long-term chronic
exposure to environmental stress, including anthropogenic contaminants, rarely results
in rapid and observably detrimental effects. Rather, and as noted by Moore et al. (2004),
the impact will be gradual, subtle and often difficult to decipher from the process and
effects of natural environmental change. The fluctuation in conditions within an
environment greatly affects the ability to attribute and effect to a cause. For example,
Klok and Kraak (2008) have demonstrated that the influence of natural pressures across
a gradient of pollution sampling stations can give rise to an unequal drop in density
and/or biomass that is not attributable to the pollutant stressor.
Organisms in their natural environment are not exposed to ionising radiation in
isolation; numerous additional stressors may be present, including additional
anthropogenic substances and natural stressors. Factors such as hot or cold
temperatures, desiccation, nutritional deprivation and pathogens can affect biomarker
responses of individuals (Holmstrup et al., 2010). Radiation is also naturally present in
the environment. Most organisms are also exposed to UV radiation, which has been
shown to be an important environmental stressor to organisms, with exposure leading to
lipid, protein and DNA damage and the production of ROS (Dahms & Lee, 2010).
Exposure of biota to anthropogenic sources of radiation may therefore be
indistinguishable from those caused by exposure to UV radiation and other naturally
occurring radioactivity in the environment. This is most likely to be the case where
chronic low-level radiation effects are being investigated. The presence of other
pollutants may also give rise to a different biomarker response than would be observed
in response to radiation exposure in isolation. For example, Bao et al. (2012) found that
chronic low-dose internal exposure of rats to cadmium induced an adaptive response to
radiation by promoting metallothionein6 production that provided protection from the
subsequent effects of gamma irradiation, and reduced micronuclei formation in
peripheral lymphocytes. Attributing any positive biomarker responses to radiation
exposure will therefore be complex in the presence of additional environmental
contaminants.
Natural variation in response is also an important factor; exposure can be affected by the
physiological status and age of an organism (Moore et al., 2004). For example, DNA
adducts7 can occur at reasonable concentrations in populations that have not been exposed
6
Proteins that bind metals and, as a result, reduce metal toxicity within the body of an
organism.
7
Altered forms of DNA that occur following exposure to carcinogenic substances. 22
to contaminants and varies considerably over season (Hebert & Luiker, 1996). The natural
occurrence of such markers may prevent association with contamination or may result in
false positive results from the incorrect interpretation of biomarker responses.
Furthermore, environmental contaminants are rarely homogenously distributed through an
environment and individuals within a population are very unlikely to be equally exposed.
As such, variation in biomarker responses would be expected to result from spatial
heterogeneity in contamination. Adaptive responses may also occur, masking effects in
exposed individuals or individuals may develop tolerance as a result of prolonged
exposure to low levels of contamination, with tolerance potentially affecting the response
of individuals to additional stressors (Marquis et al., 2009).
The choice of species for which biomarker assessments are undertaken is an important
consideration; not all species are as sensitive to pollution exposure as others. For
example, Rodriguez-Cea et al. (2003) looked at the sensitivity of a range of freshwater
fish species (European eel, European minnow and brown trout) to three different
chemical contaminants using the micronucleus assay. Both laboratory and field
experiments were undertaken with the finding that eel and minnow species had a low
sensitivity to environmental pollutants; micronuclei were not increased following
laboratory exposure to contaminants and similarly those tested from polluted rivers did
not exhibit higher micronuclei formation than those from controls. Brown trout were
however found to be responsive, both in the laboratory and in the field. It should be
noted however, that the sensitivity of species to different environmental stressors may
also be variable; whilst one species may prove sensitive to certain chemotoxicants, the
response to ionising radiation exposure (or to other stressors) may be lower than that
observed in other species.
Finally, it should be noted that no measureable effect does not guarantee that no harm is
occurring, effects may be different from those being measured or be at such a low level
that they are not readily distinguishable between sites.
23
6
DISCUSSION AND CONCLUSION
Whilst a number of biomarker techniques are available to detect effects in organisms
exposed to genotoxic pollutants, including radioactivity, few examples are evident
where the techniques have been applied under field conditions to monitor the impact of
radiation on exposed plants and animals. Where such studies have been made (e.g.
AlAmri et al., 2012; Hagger et al., 2012; Laurenço et al., 2011, 2012, 2013a, 2013b),
deciphering the cause of any observed effects has been difficult due to additional
stressors present in the environment, both natural and anthropogenic. One study
(Meehan et al., 2004) did link chronic external exposure to radiation to genotoxic
effects in bats as measured using the comet and micronucleus assays. The dose rates
experienced by the bats were however high (20-200 µGy/h) compared to what might be
expected under most of the theoretical environmental exposure situations calculated
within the safety assessments, including the current and foreseeable conditions at
Olkiluoto. Furthermore, whilst an increase in biomarker response was observed as dose
rate increased, effects were also observed in a control population (dose rate less than 2
µGy/h).
The island of Olkiluoto is home to two operational nuclear power plants, with a third
under construction, and a repository for low and intermediate level waste. Discharges
from the nuclear power plants have resulted in some detectable quantities of
radionuclides in environmental media and biota (Pere et al., 2015) although the major
anthropogenic source of radionuclides is attributed to Chernobyl fallout. Measureable
activity concentrations of the naturally occurring radionuclides Be-7 and K-40 are also
present. The presence of radioactivity in media and biota on the island is however low
and unlikely to give rise to any decipherable impact on plant and animal populations,
particularly when issues around natural variation in exposure and response are taken
into account, along with the potential presence of other contaminants that may have
arisen from past and present activities on the island, including construction.
Furthermore, Finland is subject to considerable temperature fluctuations with the
seasons and considerable variations in daylight hours. Both UV radiation and
temperature fluctuations have been linked with variation in biomarker responses
(Dahms & Lee, 2010; Holmstrup et al., 2010).
Few studies are evident in which biomarker techniques have been applied to evaluate
impacts of radiation on biota under field conditions and at exposures that would be
consistent with those that could arise from operational and post-closure discharges to
the environment from a spent fuel repository (excluding accident scenarios). Indeed,
dose rates calculated for biota (Posiva, 2014) in support of the BSA-2012 assessment
(Posiva, 2013) for the construction licence application for the Olkiluoto repository were
several orders of magnitude lower than dose rates considered to pose a risk to
populations and lower than would be expected to arise from natural background
exposure. Of the field studies that have been undertaken, it has been difficult to link any
observed biomarker responses to radiation exposure due to the presence of other
contaminants, such as heavy metals.
With the considerable uncertainties associated with the application of biomarker
techniques to evaluate radiation effects under field conditions, the addition of biomarker
analysis within the environmental monitoring programme around Olkiluoto is not
24
considered proportionate to the risks posed by any releases that may occur under normal
operating conditions. The costs of implementing such a programme are likely to be
considerable and any effects identified would be very difficult to link to radiation
exposure. Furthermore, monitoring radiation effects in wildlife is inconsistent with the
approach taken for people whereby, in recognition of the small amounts of radioactivity,
monitoring of population health was not deemed necessary (Posiva, 2008). Evaluation
of radiation effects on people is therefore based on the evaluation of radioactive releases
to derive concentrations in environmental media and throughout the human food chain,
allowing doses to be evaluated.
In light of the uncertainties identified above in relation to the application of biomarker
techniques under field conditions, it is recommended that a similar approach to that
used for human effects is adopted to allow radiation effects on wildlife to be evaluated.
Potential effects could be evaluated from derived (or preferably measured) activity
concentrations in environmental media and/or biota and dose rates calculated for
comparison against compiled radiation effects data within the FREDERICA database
(www.frederica-online.org) and/or application of screening criteria. Such an approach is
not only consistent with the approach adopted for evaluating effects on people, but is
also an established approach to evaluating the risk of harm arising from existing and
planned exposure situations (ICRP, 2014). Furthermore, use of a calculational approach
allows a multitude of interest species to be evaluated and allows for maximum use of
international knowledge on the effects of radiation.
To support a calculation approach, it is further recommended that the radionuclide
activity concentrations continue to be monitored in a range of environmental media
(soils, sediments, water) and within types of biota to establish a baseline prior to
operational activities and to evaluate any releases of radioactivity that may occur during
operations. Results of environmental radioactivity analysis will help ensure appropriate
representation of radionuclide behaviour, consistent with local conditions.
6.1
Summary
Whilst a number of biomarker techniques are available that allow the effects of genotoxic
pollutants on wildlife to be evaluated, few, if any, can be considered to specifically
measure the effects of radiation exposure. Furthermore, it is evident that many biomarker
analyses are affected by individual variability in response, and the influence of other
stressors on individuals. This is of particular importance when considering the application
of biomarkers within field monitoring programmes where plants and animals will be
exposed to an array of stressors, both natural and anthropogenic. In light of the
uncertainties and the likely cost implications, the addition of biomarker analysis within
Posiva’s environmental monitoring programme around Olkiluoto is not considered
proportionate to the risks posed by any releases that may occur under normal operating
conditions. It is therefore recommended that radiation effects on wildlife be evaluated
from biota dose rates derived on the basis of activity concentrations in environmental
media and comparison of those dose rates against radiation effects data and/or derived
screening criteria that are based on those effects data.
25
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