environmental biodosimetry: a biologically relevant tool for ecological risk assessment and...

Journal of Environmental Radioactivity 66 (2003) 121–139www.elsevier.com/locate/jenvrad

Environmental biodosimetry: a biologicallyrelevant tool for ecological risk assessment and

biomonitoring

B. Ulsh a,∗, T.G. Hinton b, J.D. Congdon b, L.C. Dugan c,F.W. Whicker a, J.S. Bedford a

a Department of Radiological Health Sciences, Colorado State University, Fort Collins, CO 80523, USAb Savannah River Ecology Laboratory, Drawer E, Aiken, SC 29802, USA

c Biology and Biotechnology Research Program, Lawrence Livermore National Laboratory, 7000 EastAvenue, PO Box 808, L-452, Livermore, CA 94551, USA

Received 1 July 2000; received in revised form 14 May 2001; accepted 15 May 2001

Abstract

Biodosimetry, the estimation of received doses by determining the frequency of radiation-induced chromosome aberrations, is widely applied in humans acutely exposed as a result ofaccidents or for clinical purposes, but biodosimetric techniques have not been utilized in organ-isms chronically exposed to radionuclides in contaminated environments. The application ofbiodosimetry to environmental exposure scenarios could greatly improve the accuracy, andreduce the uncertainties, of ecological risk assessments and biomonitoring studies, because noassumptions are required regarding external exposure rates and the movement of organismsinto and out of contaminated areas. Furthermore, unlike residue analyses of environmentalmedia, environmental biodosimetry provides a genetically relevant biomarker of cumulative life-time exposure. Symmetrical chromosome translocations can impact reproductive success, andcould therefore prove to be ecologically relevant as well. We describe our experience in studyingaberrations in the yellow-bellied slider turtle as an example of environmental biodosimetry. 2002 Elsevier Science Ltd. All rights reserved.

Keywords: Environmental biodosimetry; Chromosome aberrations; Fluorescent in-situ hybridization; Yel-low-bellied slider turtle; Trachemys scripta

∗ Corresponding author. Present address: Medical Physics & Applied Radiation Sciences, McMasterUniversity, Unit 1280 Main Street, Hamilton, Ontario, Canada, L8S 4KI. Tel.: +1-5905-525-9140x27420;fax: +1-905-528-4339.

E-mail address: [email protected] (B. Ulsh).

0265-931X/03/$ - see front matter 2002 Elsevier Science Ltd. All rights reserved.doi:10.1016/S0265-931X(02)00119-4

122 B. Ulsh et al. / J. Environ. Radioactivity 66 (2003) 121–139

1. Introduction

Humans and other organisms are continuously exposed to ionizing radiation fromnatural background sources in the environment including cosmic radiation, 222Rn andit’s daughter products, actinides and their decay products, and from internal sourcesincluding 40K and 14C. This unavoidable exposure is not without consequence, as ioniz-ing radiation exposure is known to deliver a variety of insults to nuclear DNA.

Unfortunately, natural background is not the only source of ionizing radiation towhich organisms are exposed. Numerous sites across the US (Wolbarst et al., 2000)and the rest of the world have been contaminated with radionuclides as a result ofanthropogenic activity. Human exposures can be minimized by limiting access to con-taminated areas, but this is generally not feasible for nonhuman organisms, andresulting exposures can be significantly higher than those from natural backgroundsources.

Ecological risk assessments for sites contaminated with radionuclides usually relyon indirect methods for estimating exposures to species of concern. Either screeningcalculations, which generally compare radionuclide levels in various environmentalmedia against regulatory limits, and/or fate and transport modeling, is used to estimatedoses that these species might receive. There are many uncertainties associated withsuch estimates (Whicker et al., 2000). Similarly, most current biomonitoring programsfor sites with radionuclide contamination consist of sampling of environmental mediaand tissue residue analyses of various biota (Dickerson et al., 1994). These types ofdata can often be useful in determining which species are being exposed to certainradionuclides via internal uptake, but external exposures cannot be determined. Tissueresidue analyses also provide only a snapshot picture of current internal exposure rates.They reveal nothing about the cumulative exposure an organism has recieved, thereforemany ecologically relevant effects cannot be estimated based solely on such analyses.Use of a direct biomarker of genetically relevant damage in species of concern as ameasurement endpoint could remove much of the uncertainty associated with currentecological risk assessments and biomonitoring programs and provide a meaningfulindicator of biological damage. Furthermore, if this biological or genetic damage haspotential reproductive effects, the biomarker could be an ecologically relevant assess-ment endpoint as well.

In this paper, we discuss one such measure, the frequency of symmetrical chromo-some translocations in peripheral blood lymphocytes, which is ideally suited to serveas a biomarker of cumulative radiation exposure. The term ‘dosimetry’ is generallyused to refer to the determination of dose by the observation of a radiation-inducedeffect. The use of cytogenetic techniques to observe chromosome aberrations inhumans (and a few rodent species), is termed biodosimetry, and is well-developed andwidely applied. The application of these same techniques to organisms chronicallyexposed to radionuclides in their environment, which we call environmental biodosi-metry, could provide a sensitive and biologically relevant measurement endpoint forecological risk assessments and biomonitoring programs.

One of the most important advantages of this technique is its relative specificity toradiation exposure. Ionizing radiation is very effective in the induction of DNA double

123B. Ulsh et al. / J. Environ. Radioactivity 66 (2003) 121–139

strand breaks, and subsequent chromosome exchange aberrations. Many other genotox-icants, such as heavy metals, polycyclic aromatic hydrocarbons, pesticides and PCBs,can produce single-strand breaks, and subsequently chromatid aberrations. Cells carry-ing chromatid aberrations must then survive at least two cell divisions to becomea chromosome aberration in peripheral blood lymphocytes. However, under normalcircumstances, peripheral blood lymphocytes are terminally differentiated, and aretherefore nondividing. It is only through the actions of mitogens that lymphocytes areinduced to begin cycling in vitro. Therefore, in general, only single-strand breakingagents that act on hematopoietic stem cells have the possibility to induce chromosomeaberrations later observed in peripheral blood lymphocytes (Zaire et al., 1996). Theefficiency of this process is relatively low compared to the production of chromosomeaberrations directly in peripheral blood lymphocytes through prompt (first cell divisionfollowing exposure) double-strand breaks by ionizing radiation. The one consistentexception seems to be smoking (Littlefield et al., 1998; Moore and Tucker, 1999; Pluthet al., 2000; Tucker et al., 1997a,b; Zaire et al., 1996). However, it is possible thatthese other pollutants could interact with ionizing radiation to indirectly modify theproduction of chromosome aberrations. An interesting area of future research wouldbe the possible interactions between these other contaminants and radiation exposure.

Despite its ability to produce double-strand breaks, background radiation is notbelieved to be a major source of spontaneous chromosome aberrations (Lucas et al.,1999). Rather, the background frequency of these aberrations has been shown toaccumulate with age (Tucker et al., 1999), possibly due to free-radical damage resultingfrom oxygen metabolism, which may also be an issue in a variety nonhuman species,at least in those with multi-decadal lifespans.

2. The formation of chromosome aberrations

The interaction of ionizing radiation with DNA either directly, or indirectly throughintermediate reactive oxygen species, creates a spectrum of damage including oxidizedand methylated bases, DNA adducts, and single- and double-strand breaks (Ward,1975). Of all the products of radiation interaction with DNA, double strand breaks(DSBs) are thought to be the most detrimental and resistant to repair. There are threepossible outcomes of DSBs: (1) they can be repaired, with no lasting effect on thecell; (2) they can remain unrepaired, resulting in cell death; or (3) they can be misre-paired, leading to the formation of chromosome aberrations. In turn, chromosome aber-rations can be fatal to the cell if the aberration results in a loss of genetic material atcell division, as is the case with asymmetrical chromosome interchanges (dicentrics)(Fig. 1).

It was suggested over 30 years ago that the incidence of radiation-induced chromo-some aberrations in human lymphocytes could be used to determine the magnitude ofan unknown, accidental exposure (Bender and Gooch, 1966). Estimating dose involvesconstruction of a dose–response, or calibration curve for chromosome aberrationsagainst which aberration frequencies in exposed individuals are compared to determinethe received dose. Accidental exposures are typically acute, and occur at a known point


Fig. 1. Chromosome interchange aberrations are formed when DSBs in two (or more) chromosomes inter-act and are misrepaired. Asymmetrical aberrations involve the loss of genetic material (essential genes arerepresented by letters and numbers inside the chromosomes), and are therefore fatal to the cell. Symmetricalaberrations involve no such loss, and therefore they have the potential to be stable. If a symmetrical translo-cation occurs in a germline stem cell, translocation heterozygosity can lead to a 50% reduction in repro-ductive success (any zygote which has a deficit of essential genetic material will be nonviable). Of theviable offspring produced by translocation heterozygotes, half will be normal, and half will also be translo-cation heterozygotes.

in time. Therefore, the decline of unstable aberrations from a lymphocyte population ofan accidentally exposed person can be quantified as Y(t) � Y(0)(exp(�t / tm); where,Y(0) is the initial frequency of unstable aberrations, t is time since exposure, and tmis the mean lymphocyte lifetime. Unstable aberrations were easier to detect with earlycytogenetic techniques than were stable aberrations, therefore the frequency of dicen-trics and rings in lymphocytes of exposed individuals was the biodosimetric methodof choice. However, the use of unstable aberrations for biodosimetry precluded appli-


cation of this technique to chronic exposure situations, since the behavior over timeof unstable aberrations under chronic irradiation conditions is much more complicatedthan is the case with acute exposures. The frequency of unstable aberrations inducedby chronic irradiation at any point in time reflects a balance between continuous induc-tion caused by ongoing exposure, and deletion of cells bearing these aberrations fromthe cell population.

Not all chromosome aberrations are fatal to the cells carrying them. Symmetricalchromosome interchanges (translocations) do not result in a loss of genetic material(Fig. 1), therefore they have the potential to be stable. Unless such aberrations, in andof themselves, result in a selective disadvantage relative to other cells in the population,or they coexist in cells with unstable aberrations (Lindholm et al., 1998a), their fre-quency in the cell population is not predicted to decline. The decline and disappearanceof asymmetrical aberrations has been widely observed and is not in dispute(Bauchinger, 1995; Bauchinger et al., 1986; Buckton, 1983; Buckton et al., 1978),however the evidence on the stability of symmetrical aberrations is more equivocal.Although some authors have observed an initial decline in symmetrical translocationsimmediately following irradiation (Matsumoto et al., 1998; Spruill et al., 1996, 2000;Tucker et al., 1997a), others have not observed such a decline (Lindholm et al., 1998b;Lucas et al., 1992b,1996). Regardless of whether an initial decline in translocationfrequency was observed, all these studies found that at least a fraction of the symmetri-cal translocations remain stable over time, and the frequency of these aberrationsincreases with increasing dose. In organisms subjected to chronic exposures, such asthose received as a result of inhabiting radionuclide-contaminated environments, stablechromosome translocations should accumulate over time, therefore this type of aber-ration is best-suited to serve as a biomarker of cumulative radiation exposure. Chromo-some inversions and certain more complex aberrations are also stable, but methodssuitable for their routine measurement are only now being developed.

2.1. The dose-rate effect

There is a complication in the application of biodosimetry to chronic exposures: thedependence of the frequency of chromosome aberrations on dose-rate, in addition tototal dose. Numerous studies using diverse endpoints in humans (Bauchinger et al.,1979; Bedford and Hall, 1963; Brewen and Luippold, 1971; Liniecki et al., 1977;Lloyd and Edwards, 1983; Lloyd et al., 1999; Mitchell et al., 1979) and in a varietyof other organisms (Hall and Bedford, 1964; Sax, 1939; Searle, 1974; Wells andBedford, 1983) have shown that for sparsely ionizing radiation delivered at moderatedose-rates, there is a reduction in the effect caused by a given dose when the dose isprotracted in time (Fig. 2). At very high and very low dose-rates, the effect per unitdose appears to be independent of dose-rate.

The dose-rate effect for chromosome aberration induction is generally interpretedas arising from the requirement for more than one chromosome break to produce anexchange, and that breaks rejoin (restitution) or misrejoin (exchange) with time. If fora given dose, the dose-rate is high (a few thousands of cGy h�1 or more), so that allthe breaks are produced at the same time or within a few minutes of each other, every


Fig. 2. A lower response is observed for a given dose of ionizing radiation when the dose is administeredat a low dose-rate than when the same dose is adminsistered at a high dose-rate. When the dose-rate islowered beyond a minimum value (usually on the order of 20–60 cGy h�1), no further reduction in responseper unit dose is observed. In this low dose-rate plateau, the response is independent of dose-rate, and dependsonly on the total received dose. At very high dose-rates (thousands of cGy h�1), the response again reachesa plateau, and becomes independent of dose-rate. Most environmental exposures occur at dose-rates in thelow dose-rate plateau. To ensure their relevancy to field conditions, dose–responses curves generated in thelaboratory for use in environmental dosimetry should also use dose-rates in the low dose-rate plateau.

break that is near enough to another to have some possibility of mis-rejoining to forman exchange will have the opportunity to do so. If the same dose is delivered at alower dose-rate (hundreds of cGy h�1 to a few tens of cGy h�1), even though thesame total number of breaks are produced, some will be produced early and will haverejoined or restituted so they are no longer present to have the opportunity to misrejointo produce exchanges with breaks occuring later during the protracted dose delivery.

When each of these breaks is produced by an independent event, which for sparselyionizing radiation would be the passage of a single electron track, and these interactpairwise when one occurs within a certain range of another, then the number of poten-tially interacting break-pairs would increase in proportion to the square of the numberof breaks present (the latter being directly proportional to dose). The yield of exchangesfrom two independently produced breaks, Y2, would increase according to theexpression Y2 � bD2; where, b is a constant relating to both the number of breaksper unit dose, and the probability that two breaks within a certain distance of eachother will interact to form an exchange. As the dose-rate is reduced, the probabilitythat two independently produced breaks will be produced within a sufficient periodof time to interact, will become smaller and smaller, depending on their rate of resti-tution and the yield of exchanges from these independent events, and finally will disap-


pear altogether. However, it is known from experimental observation that the yield ofexchanges for a given dose does not disappear altogether as the dose-rate approaches0 (Catcheside et al., 1946; Hall and Bedford, 1964; Lea, 1955; Sax, 1939), and it isthought that this results because break-pairs can be produced by even the smallestpossible radiation event, which is the passage of a single electron track, in additionto being produced by multiple independent electron tracks. The production of thesewould, of course, be independent of dose-rate since each of the two breaks of a break-pair is produced simultaneously by the same electron track. The yield from this singleevent break-pair production, Y1, would increase linearly with dose according to theexpression Y1 � aD; where, a is a constant relating to the number of potentially inter-acting break-pairs per unit dose and the probability that such a break pair will yieldan exchange. The response per unit dose in this lower plateau where b has gone tozero, and all break-pairs result from single electron tracks, is independent of dose-rate. The total yield of exchanges, YT, is then just the sum of Y1 and Y2, so YT �aD � bD2; where bD2 is the dose-rate dependent term and aD is the dose-rate inde-

pendent term.The dose-rate effect has important implications for environmental biodosimetry.

Environments contaminated with low to moderate levels of radionuclides are typicallycharacterized by dose-rates low enough to fall into the lower dose-rate plateau region.The application of a calibration curve constructed using a dose-rate above the lowerdose-rate effect plateau will result in underestimation of the doses received by organ-isms inhabiting such environments, as illustrated in Fig. 3.

2.2. The effects of symmetrical translocations

While symmetrical translocations do not result in loss of genetic material followedby cell death, they do result in the relocation of sections of DNA which almost certainlycontain genes essential to the organism’s survival. This can potentially lead to conse-quences far more serious for the organism than the death of a limited number of cells.Symmetrical translocations have been implicated in some forms of cancer when theyoccur in somatic (nongermline) cells (Rowley, 1990). When translocations occur ingermline stem cells, they can result in a condition known as translocation heterozygos-ity, as illustrated in Fig. 1. Every cell in the offspring produced by a germline cellcontaining a symmetrical translocation will contain the translocation. Translocationheterozygotes are semi-sterile, and 50% of their gametes are nonviable. Of the viableoffspring they produce, half will be normal, and half will also be translocation hetero-zygotes.

It has been the conventional wisdom that reproduction is the most sensitive endpointwith ecological relevance for examining the impact of radiation exposure on speciesin the environment, threshold dose-rates from 1 to 10 mGy d�1 (International AtomicEnergy Agency, 1992). The potential of symmetrical translocations to lead to translo-cation heterozygosity, with the concomitant reduction in reproductive success, givesthe frequency of these aberrations direct ecological relevance. Therefore, the endpointof symmetrical translocation frequency may be more sensitive than traditional end-points such as mortality, and potentially more relevant than tissue residue analyses.


Fig. 3. Hypothetical curves showing a curvilinear dose-response relationship resulting from administrationof the dose at a dose-rate where the dose-rate-dependent component is greater than 0 (curve A), and a linearrelationship resulting from administration of the dose at a chronic dose-rate where the dose-rate-dependentcomponent has disappeared (curve B). Assume an organism receives a dose from a contaminated environ-ment, resulting in a chromosome exchange frequency, F. Using curve A as a calibration curve for such anorganism would result in an erroneously low estimate of dose (L), while the true dose received (T) wouldbe obtained by using curve B.

Chromosome inversions, referred to previously, also have similar effects on repro-duction. The goals of radiation protection of nonhuman species is maintenance of long-term population viability (International Commission on Radiological Protection, 1991),therefore, adverse reproductive effects are of paramount concern.

2.3. Detection of chromosome aberrations

Early cytogenetic and biodosimetric studies employed solid staining with dyes suchas giemsa, orcein, or crystal violet. With these dyes, all chromosomes are stained ina single color. When cells enter mitosis, the chromosomes condense and, upon staining,they are distinctly visible under bright field illumination using a microscope. Withsolid staining, only asymmetrical chromosome aberrations (rings and dicentrics) arereliably detected because they involve a visible change in chromosome morphology.Symmetrical interchanges are not visible unless they involve large alterations in chro-mosome morphology.

Development of whole-chromosome-specific molecular probe libraries containingunique DNA sequences has facilitated significant advances in the detection of sym-metrical translocations. Such probes labeled with various fluorescent tags have beenconstructed which allow individual pairs of homologous chromosomes to be paintedin unique colors. When an interchange aberration involving the painted chromosome


is present in a cell, a fragment of the painted chromosome appears translocated toanother (unpainted) chromosome, and vice versa, and both of the chromosomesinvolved appear bicolored (Fig. 4A–C). Most generally, from one to three pairs ofhomologous chromosomes are painted either in a single or in unique colors, and therest of the chromosomes are counterstained in a single background color. Symmetricaltranslocations between chromosomes painted the same color, or among unpainted chro-mosomes are not visible.

The first step in probe construction involves isolating target chromosomes frommitotic cells using either flow cytometry or microdissection techniques. This is fol-lowed by random amplification of the isolated material through the polymerase chainreaction (PCR) (Bussey, 1996; Guan et al., 1992; Lillington et al., 1992), during whichbillions of copies of the original chromosome, each containing fluorescent markermolecules, are created. This probe library contains DNA sequences specific to theoriginal chromosome of interest, as well as sequences repeated throughout the genome.To detect chromosome aberrations, irradiated cells are spread on glass slides, and boththe chromosomes in the target cells and the DNA in the probe are denatured, or meltedby heating (Fig. 5). In this process, the DNA strands are split, much like unzippinga zipper. The probe is then “hybridized” onto the target cells, when the DNA is allowedto reanneal. The unique sequences in the probe hybridize to their unique complemen-tary sequences in the target chromosome, while the probe is prevented from hybridizingto repetitive sequences throughout the genome by the addition of unlabeled cot DNA(highly enriched in repetitive sequences), which competitively binds to repetitive siteson the chromosomes as well as labeled repetitive DNA in the probe. Once hybridiz-ation is complete, the fluorescent probe is bound only to the specific sequences onthe target chromosomes, and when viewed under a fluorescent microscope, the targetchromosomes appear “painted” in a unique color distinct from the other chromosomesin the cell. This process is known as fluorescence in situ hybridization (FISH) whole-chromosome painting (Lichter et al., 1988; Pinkel et al., 1986). The molecular basisfor DNA melting and reannealing and its dependence on DNA sequence copy numberhas been reviewed on numerous occasions (Lewin, 1997).

A further significant advance has recently been made in this technology. The appli-cation of multiple probe mixtures together in processes known as multiplex FISH(mFISH) or spectral karyotyping now allows every pair of homologous chromosomesin the human genome to be assigned a unique color (Speicher et al., 1996; Speicherand Ward, 1996). The use of mFISH for aberration detection provides dramatic poten-tial increases in sensitivity, since chromosome interchanges involving any chromo-somes can be detected. At present, the analysis of each cell requires a much longertime, so the full advantage of the potential increased sensitivity will await furtherdevelopments in computerized image analysis. With single color FISH, complex aber-rations (those involving three or more breaks in two or more chromosomes) can bemisscored because they appear to be simple, and the type of aberration scored dependson which chromosomes happen to be painted (Fig. 4A–C), but with mFISH, virtuallyall chromosome exchange aberrations can be accurately resolved (Fig. 4D). A humanAG1521A fibroblast completely colored with mFISH is illustrated in Fig. 4E.


Fig. 4. FISH allows homologous chromosome pairs to be painted in a unique color. Panels A–C show ahuman AG1521A fibroblast with different homologous chromosomes painted. Depending on which chromo-somes happen to be painted, the cell could be scored as containing either an insertion (Panel A) or asymmetrical translocation (Panels B and C). The use of mFISH allows correct resolution of the complexaberration involving chromosomes 4, 6, and 13 (Panel D). Each pair of homologous chromosomes can bepainted a unique color (Panel E), allowing maximum sensitivity for aberration detection. Application ofFISH and mFISH is not limited to human cells. Panel F shows a probe for chromosome #1 of the yellow-bellied slider turtle (T. scripta) applied to a T. scripta fibroblast containing one normal chromosome, andan apparently simple symmetrical translocation (identified by arrows).


Fig. 5. During the FISH process, both the chromosome-specific probe and the target chromosomes aredenatured. The probe is applied to the chromosomes and allowed to hybridize to the specific sequences onthe target chromosomes. The probe also contains sequences repeated throughout the genome. Nonspecifichybridization to these repetitive sites is prevented by the use of unlabeled cot DNA, which is highly enrichedin the repetitive sequences. The cot DNA competes with the probe for the repetitive sites and effectivelyblocks probe hybridization at these sites.

3. Environmental biodosimetry

Most often, biodosimetric studies have been performed on mice or rats which havereceived controlled exposures in the laboratory (Boei, Balajee, de Boer et al., 1994;Boei and Natarajan, 1998; Hande et al., 1996; Tucker et al., 1997a; 1998; Xiao et al.,1999), or on humans who have been exposed accidentally (Granath et al., 1996; Little-field et al., 1998; Lucas et al., 1992b; Moore et al., 1997; Salassidis et al., 1995;Salomaa et al., 1997; Snigiryova et al., 1997; Tucker et al., 1997b) or for therapeuticpurposes (Buckton, 1983; Buckton et al., 1978; Littlefield et al., 1991). Only a fewstudies have been carried out retrospectively on humans chronically exposed to radio-


nuclides in the environment (Bauchinger et al., 1994, 1996; Pohl-Ruling et al., 1990;Pohl-Ruling and Fischer, 1983).

There are no technical limitations preventing the application of biodosimetric tech-niques to nonhuman organisms inhabiting environments contaminated with radio-nuclides, yet, to our knowledge, these techniques have not previously been used forecological applications. This is almost certainly the result of the fact that there iscurrently very little overlap between the fields of radioecology and molecular cytogen-etics. The lack of cross-disciplinary training is the most daunting challenge facing thewidespread application of these techniques to radioecological (and ecotoxicological)problems. Provided that microdissection and microscopy equipment, in addition topersonnel with cytogenetics training and a background in chromosome microdissectionare available, probes can be constructed with a relatively modest financial investment.However, individuals with the right combination of training, skills and experience arecurrently difficult to find, and the equipment to perform chromosome microdissectionand fluorescent microscopy would require a rather substantial initial investment. Oncethe probes have been isolated, an almost unlimited supply is available at minimal costthrough PCR.

The first environmental biodosimetry studies have recently been performed in ourlaboratory using yellow-bellied slider turtles (Trachemys scripta) (Ulsh et al., 2000).Fig. 4F shows our probe for T. scripta chromosome #1 applied to a fibroblast contain-ing one normal chromosome #1 and one symmetrical translocation.

3.1. Conducting environmental biodosimetry studies—step by step

In the rest of this paper, we discuss the steps necessary to conduct environmentalbiodosimetry studies, based on our experience with T. scripta. Each step is discussedin the order in which it arose in our project. We present this synopsis as an exampleof what is required to conduct these types of studies. There will no doubt be differencesin other studies, depending on the endpoint species selected, but we believe our experi-ences may reveal the sorts of issues that may be encountered.

3.1.1. Selection of endpoint speciesWhile environmental biodosimetry has the potential to be widely applicable to

numerous plant and animal species, there are certain desirable traits that candidatespecies should possess. First, a candidate species should have a suitable karyotype,that is, they should have at least a few large and easily recognizable chromosomes.Unlike humans or mice, for which whole-chromosome probes are commercially avail-able, probes for other species will have to be constructed by microdissection (or flowsorting) and PCR. For reasons discussed later, microdissection requires the presenceof large, easily identifiable chromosomes, which could eliminate species with onlysmall, indistinguishable chromosomes from consideration.

Even if limitations associated with microdissection are overcome, there is still theissue of the minimum resolution of the FISH visualization process. The minimumdetectable size of a painted fragment of DNA is approximately 11 megabases (Mb)or approximately 15 Mb for unpainted fragments (Kodama et al., 1997). Since chromo-


somes are fragmented during chromosome interchanges, species with at least a fewchromosomes larger than approximately 30 Mb would make the best candidates forenvironmental biodosimetry studies.

Demographic factors and the exact nature of the ecological impact to be studiedare also important factors in the selection of endpoint species. Organisms with rela-tively long life-spans will have the potential to accumulate significantly higher lifetimedoses than those which are relatively short-lived, and therefore they will be more likelyto show a measurable response to chronic, low-level radiation exposure, all other fac-tors being equal. On the other hand, studies of reproductive effects (such as thosecaused by translocation heterozygosity) at the population level are more easily studiedwith short-lived species with large numbers of offspring. Finally, candidate speciesshould be those with some potential to be exposed. Since many environmental contami-nants (including most radionuclides) are eventually deposited and sequestered in sedi-ments (Blaylock et al., 1987; Whicker et al., 1990), organisms which have contactwith sediments could make strong candidates. Tissue residue analyses, which measurelevels of contaminants in the tissues of various biota, may provide clues regardingwhich organisms are being exposed and what the current exposure rates from internallydeposited radionuclides might be.

It is evident that any number of plant and animal species would make strong candi-dates for environmental biodosimetry. Perhaps the best approach for ecological riskassessments and biological monitoring applications would be to select a suite of suit-able species representing diverse ecological, economic, and aesthetic values.

3.1.2. Probe constructionIn general, the largest chromosome(s) are targeted, since the greater the fraction of

the genome painted, the greater the sensitivity for detecting chromosome aberrations(Lucas et al., 1992a). If standard lymphocyte culture techniques work well with thespecies being studied, harvesting mitotic lymphocytes from peripheral blood would bethe best and most direct approach for preparing chromosomes for microdissection. Ifthis is not the case, it is easier to use fibroblasts at this stage because of the challengesin stimulating lymphocytes into mitosis. Fibroblast cell lines can be established fromembryos or from tissue samples.

3.1.3. Investigation of lymphocyte culture techniquesIn conducting environmental biodosimetry studies involving animal species, it is

preferable to use lymphocytes rather than fibroblasts (Fossi, 1994). Lymphocytes offerthe advantage of nonlethal sampling, which allows repeated blood sampling from indi-vidual organisms so the temporal behavior of the dose–response can be studied. Fur-thermore, animal welfare considerations or potential adverse impacts to endpoint popu-lations from harvesting large numbers of individuals may also favor nonlethalsampling techniques.

The challenge in using lymphocytes is stimulating them to undergo mitosis. Thesecells are usually noncycling, and they must be forced into the cell cycle by mitogenicagents. The mitogenic responses of lymphocytes from nonmammalian species is notwell characterized. Therefore, factors such as which mitogenic agents are effective for


a given species and the optimal concentrations of these agents, optimal culture tem-perature, cell cycle time, etc. will have to be determined. Of course, the availableblood volume and methods for collection may provide additional challenges.

3.1.4. Determination of background symmetrical translocation frequencyTo some extent, the sensitivity of ecological applications of biodosimetry will

depend on species-specific radiosensitivity, in particular, the number of chromosomeaberrations produced by a given dose of radiation. What may not be so obvious how-ever, is that species which are more resistant to the production of chromosome aber-rations, ironically, may be more sensitive indicators of radiation damage. This arisesfrom the fact that species which are more radioresistant may also have lower back-ground levels of symmetrical translocations, as is the case with T. scripta, which wefound to be about twice as radioresistant as humans (Ulsh et al., 2000). Backgroundfrequencies of symmetrical translocations in humans have been reported as much as30 times higher than the background we observed in T. scripta (Barquinero et al.,1999). A high background could significantly impact sensitivity, especially at the lowdoses and dose-rates likely to be encountered by organisms from radionuclide-contami-nated environments. If the factors which make a particular species more radioresistantalso depress background translocation frequencies, then the loss in sensitivity causedby higher radioresistance may be outweighed by the gain in sensitivity afforded by alower background. In any case, some estimate of background translocation frequencywill be necessary, since the organisms to be studied would have received unknowndoses. Without an estimate of background, it would be unclear how much of theobserved translocation frequency in exposed animals was due to radiation exposure,and how much was due to extraneous factors. The best way to obtain an estimate ofbackground is to score cells from numerous organisms known to be from uncontami-nated environments.

3.1.5. Investigation of dose-rate effectAs mentioned previously, it is important that the calibration curve to be used for

environmental biodosimetry be determined for low dose-rate exposures (Fig. 3). How-ever, for practical reasons, laboratory studies will almost certainly have to be conductedat dose-rates higher than those observed in contaminated environments. The applica-bility of the calibration curve can only be definitively demonstrated by identificationof the minimum dose-rate below which the reduction in effect per unit dose plateaus(Fig. 2). This involves exposing whole organisms or cell cultures to the same totaldose, but delivering the dose at a range of dose-rates from thousands of cGy h�1 toonly a few cGy h�1. For almost every organism previously studied, the lower dose-rate effect plateau begins at a rate on the order of 20–60 cGy h�1.

3.1.6. Determination of dose–response relationshipOnce an appropriate dose-rate is identified, it should be used in the construction of

the calibration curve. This involves plotting the dose–response relationship over arange of doses that the organisms are expected to receive over the course of theirlifetime. A preliminary estimate of the maximum dose which should be included in


the calibration curve can be obtained by multiplying worst-case environmental dose-rates by the maximum expected lifespan of the endpoint species. However, the mainfocus should be at the lower end of the of the dose-range, since these exposures wouldbe more typical of most environmental exposures.

3.1.7. Applying environmental biodosimetry to organisms from contaminatedenvironments

Upon completion of the preliminary steps outlined above, a variety of field studiesare possible. These could include mark-and-recapture studies of animals, since lympho-cyte sampling is nonlethal. This type of study would be particularly appropriate forbiomonitoring programs and would provide useful data for ongoing exposures. Forecological risk assessments upon which remediation decisions are to be based, anextensive, one-time sampling effort might be more appropriate. An exploration of thereproductive effects caused by radiation-induced translocation heterozygosity wouldbe particularly informative.

4. Conclusion

Environmental biodosimetry studies offer several advantages over traditionalapproaches to biomonitoring and ecological risk assessment. Unlike dose estimatesobtained by modeling, biodosimetric estimates require no assumptions regarding organ-ism movements into and out of contaminated environments (which may be difficultto verify). Furthermore, actual doses received externally, or internally via uptake ofradionuclides will be reflected in biodosimetric estimates. Another advantage ofenvironmental biodosimetry is that an estimate of cumulative lifetime dose can beobtained, rather than the snapshot picture provided by environmental sampling andtissue residue analyses. Perhaps the most significant advantage of environmental biodo-simetry, however, is its potential relevance to ecological and biological effects inexposed populations.

While environmental biodosimetry can currently serve as a useful measurementendpoint for ecological risk assessment, completing the causal chain of events betweenradiation exposure, the formation of radiation-induced symmetrical translocations, andreproductive effects through translocation heterozygosity would provide an ecologi-cally relevant assessment endpoint. Once received dose, as detected by environmentalbiodosimetry, can be linked to reproductive effects in exposed individuals, “ecologicaldosimetry” studies could be conducted on a variety of species to study effects ofradiation exposure at higher levels of biological organization.

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