dimanche 14 septembre 2008

R. I. Goncharova: Transgenerational accumulation of radiation damage

Radiat Environ Biophys (2006) 45: 167–177
DOI 10.1007/s00411-006-0054-3
Nadezhda I. Ryabokon · R. I. Goncharova
Transgenerational accumulation of radiation damage
in small mammals chronically exposed to Chernobyl fallout
Received: 5 March 2006 / Accepted: 17 June 2006 / Published online: 22 July 2006
© Springer-Verlag 2006

Abstract The purpose of this investigation has been the
analysis of the long-term development of biological damage
in natural populations of a model mammalian species,
the bank vole (Clethrionomys glareolus, Schreber),
which were chronically exposed to low doses of ionizing
radiation over 22 animal generations within 10 years following
the Chernobyl accident. The time course of the
biological end-points (chromosome aberrations in bone
marrow cells and embryonic lethality) was compared
with the time course of the whole-body absorbed dose
rate from external and internal exposure in the studied
populations inhabiting monitoring sites in Belarus with
diVerent ground deposition of radionuclides. The yield of
chromosome aberrations and, in lesser degree, embryonic
lethality was associated with the radionuclide contamination
of the monitoring areas in a dose-dependent
manner. As a main feature of the long-term development
of biological damage under low dose rate irradiation,
permanently elevated levels of chromosome aberrations
and an increasing frequency of embryonic lethality have
developed over 22 animal generations. This contrasts
with the assumption that the biological damage would
gradually disappear since in the same period of time the
whole-body absorbed dose rate decreased exponentially
with a half-value time of about 2.5–3 years. Furthermore,
gravid females were captured, and their oVspring, born
and grown up under contamination-free laboratory conditions,
showed the same enhanced level of chromosome
aberrations. Therefore the authors suggest that, along
with the biological damage attributable to the individual
exposure of each animal, the observed cellular and systemic
eVects reXect the transgenerational transmission
and accumulation, via genetic and/or epigenetic pathways,
of damage attributable to the chronic low-dose
rate exposure of the preceding generations of animals.
They also suggest that the level of the accumulated transmissible
damage in the investigated populations will
decrease in future due to the further recession of the
chronic exposure and as a consequence of selection processes.
The Chernobyl accident in April 1986 caused the deposition
of radionuclides across Europe, followed by a longterm
artiWcial increase of the radiation background [1].
In addition to the classical subject of mutagenesis after
acute radiation exposure [2], the study of the time course
of biological damage associated with chronic low-dose
radiation exposure of mammals and the endeavors to
predict biological damage in consecutive generations
have become a relevant issue. Since there is little information
on this topic [3, 4], the present work addresses
these important questions.
Starting with 1986, we were engaged in studying biological
eVects of chronic low dose radiation in natural
populations of bank vole (Clethrionomys glareolus, Schreber)
in a series of many animal generations. The bank
vole is a widespread rodent species that is used as indicator
of environmental quality. It is a convenient object for
many genetic tests, which originally have been devised for
the laboratory mouse [5]. Comparison of own and literature
data on doubling doses of acute irradiation for chromosome
injuries had shown that the sensitivity of somatic
cells of the bank vole to ionizing radiation is very similar
to the sensitivity of human lymphocytes and germ cells of
laboratory mice [6]. This Wnding conWrms that the bank
vole is a suitable model species for assessment of genetic
radiation risks in mammals.
According to the Atlas of Caesium Deposition in
Europe [1], practically the whole Belarusan territory was
contaminated by 137Cs above the level of global fallout.
A ground deposition of 137Cs equal to 37 kBq/m2 (1 Ci/
km2) was chosen to discriminate between so-called
“clean” regions and contaminated regions. We selected
monitoring sites representing large areas of Belarus with
a strong gradient of radionuclide ground deposition,
from 8 to 8,500 kBq/m2 of 137Cs, i.e., from light contamination
in “clean” regions to high contamination in the
evacuated zone. In comparison with the pre-Chernobyl
data, increased frequencies of chromosome aberrations
and genomic mutations in somatic cells, abnormal sperm
heads and embryonic losses were observed in bank vole
populations at the monitoring sites [7, 8].
The aim of the present study was to analyze the longterm
development of chromosome aberration frequency
and embryonic lethality in bank vole populations over
22 animal generations living in 1986–1996 and to compare
it with the time course of the whole-body absorbed
dose rate.
Materials and methods
Monitoring sites
The animals were collected at Wve monitoring sites in
large forestry areas with limited human activities, far
from possible sources of industrial or domestic pollution.
These sites are located at diVerent distances and directions
from the Chernobyl nuclear power plant (Fig. 1)
and represent diVerent levels of radionuclide contamination:
site 1—the Priluksky Reserve, Minsk region,
330 km NW, site 2—the Berezinsky Biosphere Reserve,
Vitebsk region, 400 km NNW, site 3—the vicinity of
Majsk village, Bragin district, Gomel region, 60 km N,
site 4—the vicinity of the evacuated Babchin village,
Khoiniki district, Gomel region, 40 km NNW and site
5—the vicinity of the evacuated Radin village, Khoiniki
district, Gomel region, 18 km N. The initial levels of
radionuclide contamination of soil at these sites were
determined earlier [9] and are shown in Table 1. The animals
were captured using live traps with bait. The Wrst
animals were captured at sites 1, 3, and 4 in September
1986, i.e., about 5 months after the period of acute irradiation
by short- and long-lived radionuclides. In the subsequent
years, the animals were as well usually collected
in the period of August–September in order to exclude
the inXuence on the data of seasonal changes in the age
structure of populations. Data obtained outside this collection
period (e.g., the data from site 1 in 1991 and 1996
[7, 8]) were not included in this study. Investigations at
sites 2 and 5 started in 1991 and 1996, respectively, i.e., 5
and 10 years after the accident.
Assessment of animal age and change of animal
For age determination, root, cusp, and height of the Wrst
mandibular molar (M2) were measured according to
Bashenina [10]. The specimens were divided into seven
age groups, which corresponded to ages of approximately
2 weeks, 1, 2, 3, and 4 months and 1 and 1.5 years.
The bank vole, like other rodent species, is known to
have a short generation time with a complete change of
about 2–3 generations per year [10]. According to data of
Rozhdestvenskaya [11] and our own estimates of animal
participation in reproduction, this holds also for the
Fig. 1 Map of Belarus with localization of the monitoring sites. The
position of the Chernobyl nuclear power plant (ChNPP) in the
neighboring Ukraine is also indicated
Table 1 Densities of radionuclide contamination of soil (kBq/m2) at Wve monitoring sites (data of April–May 1986 for 137Cs, 134Cs, 106Ru,
144Ce and data of August 1996 for 90Sr and transuranic radionuclides) according to data [9]
Site April–May 1986 August 1996
137Cs 134Cs 106Ru 144Ce 90Sr 238Pu 239,240Pu 241Pu 241Am
1 8 4 5 0 4 0.04 0.10 2.98 0.14
2 18 9 12 0 5 0.07 0.14 5.10 0.19
3 220 140 150 440 39 0.62 1.28 48.80 1.81
4 1,530 1,020 1,090 3,050 117 1.17 2.35 86.70 3.21
5 8,500 5,650 5,790 1,7200 1,200 4.90 11.00 420.00 15.00
populations studied in the present work. Due to seasonal
reproduction and short life-span, in the trapping periods
of late summer and early autumn only few animals born
in the preceding year were observed at the monitoring
sites, showing that the populations were almost completely
renewed every subsequent year by the time of
The populations living at the time of the accident, in
early spring of 1986, consisted of adult animals born in
the previous year. Due to the turnover of generations,
our investigations, beginning in September 1986, started
with the Wrst and second post-accidental generations of
bank vole. In the period from 1986 to 1996, at least 22
generations were studied.
Determination of radionuclide concentration
and estimation of absorbed dose rate
Concentrations of radionuclides in the soil and the
whole-body animal samples were assessed as described in
[9]. BrieXy, the _-spectrometry of samples was performed
in the Hydrometeorological Centre of Belarus in Minsk
using the _-spectrometer ADCAM-300 equipped with a
high-purity germanium detector GEM-30185 (EG&G
Ortec, USA). SpeciWc activities of 90Sr and transuranic
elements were determined by the staV of the Institute of
Radiobiology, National Academy of Sciences of Belarus,
using radiochemical methods and the _-spectrometer
Ortec 576A with the silicon surface barrier detector
(EG&G Ortec, USA), the liquid scintillation counter Tricarb
2700TR (Packard Company) and a gas-Xow
counter (Tesla Automat, Slovakia).
Dose rates due to internal and external exposure were
assessed in the studied specimens as described previously
[9]. The whole-body absorbed dose rate from incorporated
_-emitting radionuclides was calculated according
to the absorbed fraction model. For the dose rate attributable
to incorporated _- and _-emitters, we used the
local absorption model. The whole-body absorbed dose
was calculated as the product of the individual wholebody
absorbed dose rate and the age of the animal at the
time of capture.
Metaphase analysis of chromosome aberrations
Cytogenetic eVects in bank vole somatic cells were studied
by metaphase analysis of chromosome aberrations in
red bone marrow cells according to a standard protocol
[12] as described [7, 8]. BrieXy, visually healthy animals of
diVerent age, sex, and state of maturation were randomly
chosen for this test from numerous groups of animals
captured shortly after trapping. Colchicine at concentration
of 0.1 mg per 10 g animal weight was injected intraperitoneally
for 1.5 h to accumulate metaphases.
Animals were sacriWced by ethyl ether and cervical dislocation,
and the marrow was aspirated from the femurs
using inactivated fetal bovine serum. The marrow suspension
was incubated at 37°C for 20 min, treated with
0.56% potassium chloride for the next 20 min and Wxed
in methanol:acetic acid (3:1 v/v). Fixed cells were spread
on clean slides, Xame-dried and stained with Giemsa.
Coded slides were screened for chromatid- and chromosome-
type aberrations in approximately 100 well-spread
metaphases per specimen using standard criteria [12, 13].
Achromatic lesions (gaps) were not included in the statistical
Selected age groups
The chromosome aberrations and embryonic losses were
also recorded from selected cohorts of mature animals
with age from 2 to 4 months in order to exclude any agedependent
bias of the biological eVects.
Furthermore, two gravid females were taken to a laboratory
in Minsk, where they gave litters shortly after
capture. Mother animals and their oVspring were fed
with uncontaminated food and were studied for chromosome
aberrations in red bone marrow cells when the age
of the oVspring reached 1.5 months.
Analysis of embryonic losses
Assessment of embryonic mortality was performed in all
captured and visually healthy females at 7–22 days of
pregnancy using the conventional approaches [14]. The
content of the uteri was examined to determine the number
of implants including live and dead embryos. The
early (pre-implantation) losses were calculated as the
ratio (number of corpora lutea minus number of
implants)/(number of corpora lutea). The late (postimplantation)
losses were calculated as the ratio of the
number of dead embryos to the number of implants. The
total losses were determined as the ratio (number of corpora
lutea minus number of live embryos)/(number of
corpora lutea). All ratios are expressed in percent.
Chi-square and U tests, regression and correlation analysis
were employed as a part of the Statistica software
package (StatSoft Inc., USA). Figure 2 gives an example
of the scatter of the individual aberration frequency and
whole-body dose rate data at a given site and given season
of the study (site 4, analyzed 5 years after the accident).
In the following, these Xuctuations are represented
by the mean value and the standard deviation of the
Time course of the absorbed dose rate
The time course of the absorbed dose rates in populations
of bank vole at the monitoring sites was determined
previously [9] using numerous groups of captured
animals. It was shown that external _-irradiation and
internal _-irradiation by incorporated 137Cs and 134Cs
delivered the most prominent contributions to the
whole-body dose rate over the monitoring period, which
started 5 months after the accident with the Wrst and second
post-accident animal generations and ended 1996
after about 22 animal generations [9]. Here, we analyzed
the time course of the mean whole-body absorbed dose
rate due to internal _-irradiation and to total irradiation
speciWcally for those animals which, out of the whole
cohorts, were randomly chosen for assessment of chromosome
aberrations (Table 2; Fig. 3a, b) and embryonic
mortality (Table 3; Fig. 3c, d). The values of the wholebody
dose rates and their temporal development in these
sub-cohorts of animals were similar to those in the larger
groups of bank vole used in the earlier dosimetrical
investigations of monitored populations [9].
Fig. 2 Correlation of the chromosome aberration frequency in
bone marrow cells of bank voles inhabiting site 4, with the individual
whole-body absorbed dose rates 5 years after the Chernobyl accident.
The dose rates are due to external _- and internal _ + _-radiation
of incorporated 137Cs and 134Cs
Mean value
60 80 100 120 140
Whole-body dose rate, μGy/d
Chromosome aberrations per 100 cells
Fig. 3 Time course of the mean whole-body absorbed dose rates in
bank vole specimens captured at sites 1–5 and studied for chromosome
aberrations (a, b) and for embryonic mortality (c, d). The contributions
by internal _-irradiation from incorporated 137Cs and
134Cs are presented in a and c, the total whole-body dose rates by
external _-irradiation and internal _ + _-irradiation from incorporated
137Cs and 134Cs are shown, together with exponential approximation
curves, in b and d. Error bars are not shown for clarity
Internal â-irradiation Total irradiation
Site 1 Site 2
Site 3
Site 4 Site 5
0 6 10
Years since accident
Site 1
Site 2
Site 3
Site 5
Site 4
Years since accident
Whole-body dose rate, μGy/d
Site 2
Site 3
Site 4
Years since accident
Site 2
Site 3
Site 4
Years since accident
2 4 8 0 2 4 6 8 10
0 2 4 6 8 10 0 2 4 6 8 10
(a) (b)
Table 2 Chromosome aberrations in bone marrow cells of bank vole at Wve monitoring sites
a Pre-accident data according to [15]
b All data on total frequencies of aberrations and aberrant cells in animals at sites 2–5 in 1986–1996 were signiWcantly higher in comparison
with pre-accident data by Chi-square test
Site Year Animal generation
since the accident
absorbed dose
rate (_Gy/day)
Number of analyzed Aberrations per 100 cells § SD Aberrant
cells, %
Animals Cells Chromatid
1 1986 1–2 6.44 § 0.14 10 992 0.30 § 0.14 0.10 § 0.09 0.40 § 0.22 0.40 § 0.22
1988 5–6 3.36 § 0.02 3 310 0.64 § 0.32 0 0.64 § 0.32 0.64 § 0.32
2 1981–1983a Pre-accident – 24 2,437 0.41 § 0.12 0 0.41 § 0.12 0.41 § 0.12
1991 11–12 4.43 § 0.38 19 1,945 1.19 § 0.37 0.03 § 0.03 1.22 § 0.37b 1.22 § 0.37
1992 13–14 – 17 1,962 0.90 § 0.27 0.27 § 0.17 1.17 § 0.13 1.09 § 0.24
1996 21–22 2.41 § 0.03 8 585 1.06 § 0.54 0.11 § 0.11 1.17 § 0.53 1.17 § 0.53
3 1986 1–2 87.52 § 4.50 18 1,987 1.36 § 0.45 0.76 § 0.28 2.12 § 0.59 1.99 § 0.53
1988 5–6 76.06 § 8.90 16 1,630 1.22 § 0.49 0.52 § 0.18 1.74 § 0.57 1.51 § 0.03
1991 11–12 16.83 § 2.27 16 1,655 2.67 § 0.81 0.20 § 0.13 2.87 § 0.83 2.49 § 0.76
1996 21–22 7.02 § 0.27 11 1,121 1.86 § 0.41 0.18 § 0.10 2.04 § 0.39 1.95 § 0.38
4 1986 1–2 605.46 § 7.75 16 1,739 1.45 § 0.43 0.06 § 0.06 1.51 § 0.45 1.21 § 0.10
1987 3–4 258.57 § 11.33 36 3,675 0.79 § 0.18 0.44 § 0.15 1.24 § 0.24 1.10 § 0.22
1988 5–6 321.64 § 33.12 21 1,942 1.75 § 0.39 0.17 § 0.09 1.92 § 0.40 1.86 § 0.04
1991 11–12 95.93 § 3.69 31 3,780 1.63 § 0.34 0.31 § 0.10 1.94 § 0.40 1.84 § 0.36
1996 21–22 41.80 § 0.96 14 1,821 1.90 § 0.49 0.24 § 0.15 2.14 § 0.42 2.04 § 0.38
5 1996 21–22 274.52 § 7.27 11 492 4.74 § 0.74 1.47 § 0.61 6.21 § 0.85 5.10 § 0.78
The maxima of internal _-irradiation due to incorporated
137Cs and 134Cs were observed in the second year
after the accident (Fig. 3a, c) due to the increasing biological
availability of cesium isotopes from the biota [7-9], followed
by a decrease by one order of magnitude in the
subsequent period. In [9] it was shown that external _-irradiation
was the essential determinant for the time course of
the total whole-body dose rate; this was also true for the
animals analyzed for chromosome aberrations (Fig. 3b) or
embryonic losses (Fig. 3d). The highest values of total dose
rate were observed in the year of the accident, followed by
an approximately exponential decrease over the next
10 years (r2=0.95, P<0.05 and r2=0.91, P<0.05 for
sites 3 and 4, respectively, in Fig. 3b). The reduction factor
since the accident was about 15, corresponding to a halflife
of 2.5–3 years. These data show that after the primary
insult in the year of the accident, every successive generation
of animals was exposed to lower whole-body dose
rates of ionizing radiation than the preceding one.
The mean whole-body absorbed doses in the studied
cohorts of animals were also maximal in the year of radionuclide
deposition. The highest individual value, determined
in the Wrst and second post-accidental generations
of animals at site 4, was 73 mGy. By the end of the monitoring
period, the mean whole-body doses were about 0.3,
0.7, 12, and 25 mGy for sites 2, 3, 4, and 5, respectively [9].
At any given time of the observation period, absorbed
dose rates (Fig. 3) and absorbed doses at the diVerent sites
diVered by about two orders of magnitude, thereby representing
a strong gradient in the level of animal exposure.
Temporal development of the chromosome aberration
The time course of chromosome aberration frequencies
in the studied animals is presented in Table 2. The frequency
of chromosome aberrations in bone marrow cells
of the bank vole inhabiting the Berezinsky Biosphere
Reserve, i.e., site 2 in our study, was Wrst recorded 3–5 years
before the accident [15] and can be taken as a pre-accident
(historical) control for our data. The data of this
pre-accident control are similar to those observed at the
less contaminated site 1 (Minsk region) in the years 1986
and 1988. Before the accident, only chromatid-type aberrations
were observed, and the cells never contained multiple
aberrations. In contrast, in the post-accident period,
the chromosome anomalies consisted of both chromatidand
chromosome-type aberrations, containing single and
paired fragments in the majority of cases, but also rare
Robertsonian translocations (fusion of acrocentric chromosomes
at the centromeres). In animals inhabiting
contaminated sites, the mean frequencies of both chromosome
aberrations and aberrant cells were signiWcantly
higher (P<0.01, Chi-square test) than the pre-accident
value, and were increased in a dose-dependent manner,
by a factor of 3–7 at sites 2, 3, and 4, and of about 15 at
site 5. Also Pearson correlation analysis of the total aberration
frequencies and those of the Robertsonian translocations
at sites 2, 3 4, and 5 in 1996 (Fig. 4a, b) hints at
a relationship between mean aberration frequencies and
the whole-body dose rates at the time of capture. The
increased frequency of chromosome aberrations
observed in animals inhabiting contaminated areas
appears to remain relatively constant over the years of
the investigation (Table 2).
The data in Table 2 comprise the level of chromosome
aberrations in animals varying by age, sex and
maturity. These animals represent the studied populations,
from where they were chosen at random. Generally,
the mean age of the animals in each group did not
diVer from the usual mean age of about 3–4 months at
the season of capture. There was a single exception at site
Table 3 Embryonic lethality in bank vole populations at three monitoring sites
a Pre-accident level according to [16]
* P< 0.05 and ** P< 0.01 in comparison with pre-accident data (Chi-square test)
Site Year Animal
since the
absorbed dose
rate (_Gy/day)
Number of Mean embryonic lethality,
% (95% binomial conWdence limits)
2 1981–1983a Pre-accident – 45 249 235/4 5.62 (3.11–9.25) 1.70 (0.47–4.30) 7.23 (4.34–11.18)
1991 11–12 5.5 § 0 12 56 56/3 0 (0.00–5.21) 5.36 (1.12–14.87) 5.36 (1.10–14.87)
1992m 13–14 – 19 92 90/2 2.17 (0.26–7.68) 2.22 (0.27–7.80) 4.35 (1.20–10.76)
1996 21–22 2.51 § 0.13 8 44 39/0 11.36 (3.79–24.56) 0 (0.00–9.03) 11.36 (3.79–24.56)
3 1988 5–6 58.01 § 5.40 7 36 36/0 0 (0.00–7.98) 0 (0.00–7.98) 0 (0.00–7.98)
1989 7–8 44.38 § 3.71 30 138 130/1 5.80 (2.50–11.10) 0.77 (0.02–4.21) 6.52 (3.03–12.08)
1991 11–12 17.95 § 2.53 14 71 65/1 8.45 (3.16–17.49) 1.54 (0.04–8.28) 9.86 (4.06–19.26)
1996 21–22 6.58 § 0.07 3 15 12/1 20.00* (4.33–48.09) 8.33 (0.21–38.48) 26.67* (7.74–55.10)
4 1988 5–6 245.6 § 20.21 14 63 61/1 3.17 (0.39–11.00) 1.64 (0.04–8.80) 4.76 (0.99–13.29)
1989 7–8 265.63 § 34.53 40 201 192/3 4.48 (2.07–8.33) 1.56 (0.32–4.50) 5.97 (3.12–10.20)
1991 11–12 103.78 § 5.96 21 103 91/4 11.65* (6.27–19.76) 4.40 (1.21–10.87) 15.53* (9.15–24.00)
1996 21–22 43.47 § 0.87 11 51 44/4 13.73* (5.70–26.26) 9.09* (2.53–21.67) 21.57** (11.29–35.32)
4, 10 years after the accident, when by chance 10 of 14
animals randomly sampled for cytogenetic analysis were
at the age of one year or older, whereas in the more
numerous group of specimens captured at this site for
dosimetrical investigations [9] there was no detectable
change in age distribution. In order to exclude any possible
inXuence of variations in the age distribution on the
observed frequencies of chromosome aberration, analysis
of the time course of the chromosome aberration frequency
was limited to mature 2–4 months old animals,
whose mean values and temporal development of age
over the period of this study is shown in Fig. 5a. The
temporal pattern of the frequency of chromosome aberrations
in the 2–4 months old animals (Fig. 5b) was
almost identical with the pattern observed in all animals
(Table 2).
As is evident from Fig. 5, the frequencies of chromosome
aberrations observed in 2–4 months old bank voles
remain fairly constant over time, in spite of the approximately
exponential decrease of the whole-body dose rate.
This suggests that the induction of aberrations depends
not only on the actual exposure level. To test this, aberration
frequencies were investigated in the oVspring of
mice captured in 1988 at sites 3 and 4, which were born
and brought up in the laboratory and fed with uncontaminated
food. The chromosome aberration frequency
observed in the oVspring animals showed no signiWcant
diVerence to that observed in animals from the same sites
that had grown up in the contaminated environment,
and to the aberration frequency of their mother animals
(Fig. 6).
Embryonic mortality
Studies on embryonic losses in the monitored populations
were started 2 years after the accident. Since at sites
1 and 5 no gravid females after the stage of embryonic
implantation were captured, the analysis on embryonic
mortality is limited to sites 2, 3, and 4. Figure 7a shows
that there was no increase over the monitoring period in
the age of gravid females, which could have aVected the
values of embryonic lethality. Rather, the mean age of
Fig. 4 Correlation between mean frequency values of all types of
chromosome aberrations (a) and of Robertsonian translocation (b)
with the mean values of the whole-body absorbed dose rate in bank
vole populations inhabiting sites 2–5, 10 years after the accident.
Standard deviations of the mean values are indicated as error bars
Site 2
Site 3
Site 4
Site 5
r = 0.99
P = 0.011
0 100 200 300
Whole-body dose rate (μGy/d)
Chromosome aberrations per 100 cells
Site 5
Site 4
Site 3
Site 2
r = 0.95
P = 0.048
0 100 200 300
Whole-body dose rate (μGy/d)
Robertsonian translocations per 100 cells
Fig. 5 Time course of animal age (a) and chromosome aberration
frequency (b) in a sub-cohort of adult, 2–4 months old bank voles at
four sites. Mean data and standard deviation are shown. Pre-accident
data according to [15]
0 2 4 6 8 10
Years since accident
Age of animals (months)
Site 1
Site 2
Site 3
Site 4
Site 1
Site 2
Site 3
Site 4
0 2 4 6 8 10
Years since accident
Chromosome aberrations per 100 cells
the gravid females tended to decrease over the period of
monitoring. The frequencies of embryonic losses in the
population inhabiting the Berezinsky Biosphere Reserve
3–5 years prior to the accident [16] were used as pre-accident
control (Table 3; Fig. 7b). A comparison of pre- and
post-accident data shows that both pre- and postimplantation
embryonic mortality in populations in the
moderately contaminated site 2 and in the more highly
contaminated sites 3 and 4 remained on the level of preaccident
mortality frequencies during the Wrst years of
observation. After 5–10 years, however, a tendency
towards increasing embryonic lethality was found in all
populations studied, reaching a statistically signiWcant
increase in samples from sites 3 and 4 after 10 years (by a
factor 2–5 in comparison with pre-accident data).
In this study free-living small mammals, the bank voles,
were used as indicator organisms for the monitoring of
environmental eVects. The studied populations lived at
diVerent distances from the Chernobyl nuclear power
plant. The investigations were started 5 months after the
accident, from the 1 and 2 post-accident generations of
animals that were irradiated at doses one order of magnitude
less than animals in the acute period after the
accident. The particular subject of this study was to compare
the time course of biological damage in animals,
representing up to 22 generations living during 10 years
following the Chernobyl accident, with the time course
of the radiation exposure experienced by these animals.
The main results of this study are (a) a dose-dependent
increase in the frequencies of chromosome aberrations
and embryonic losses in animals living in contaminated
areas as compared to historic controls and to animals
living in an area with the lowest level of contamination,
(b) the fact that the chromosome aberration frequencies
remained on elevated levels (Fig. 5) and the frequencies
of embryonic losses even increased (Fig. 7) over 10 years
after the accident, although the whole-body absorbed
dose rates experienced by the animals declined almost
exponentially (Fig. 3).
Historic controls raise the concern that they might
not directly be comparable to the samples under study.
In this study the frequencies of chromosome aberrations
observed in 1986 and 1988 in animals living at the almost
uncontaminated site 1 can be used as controls, and they
were similar to those assessed in the historic control
(Table 2; Fig. 3). Thereby we can exclude that the
increases in chromosome aberration frequency seen in
the contaminated areas have reasons other than the
higher radiation exposure. Unfortunately, we have no
data from site 1 for the years 1991 and 1996 collected in
the period August–September, but the fact that all investigations
at the contaminated sites were done by the
same team guarantees that the methods of investigation
Fig. 6 Chromosome aberration frequencies in bank vole females
captured at sites 3 and 4 in 1988, and of their oVspring, born, and
raised under laboratory conditions for about 1.5 months before
analysis. Data are compared in bank vole populations living at sites
3 and 4 at the same period. Means, standard deviations of the means
and the number of analyzed animals are shown
n = 2
n = 9
n = 37
Mothers Offspring Populations
Chromosome aberrations per 100 cells
Fig. 7 a: Time course of the age of gravid bank vole females analyzed
for embryonic losses. Means and standard deviation of the
means are show. b: Time course of embryonic lethality in animals at
site 2 and at sites 3 and 4 (pooled). Means and binomial 95% conWdence
limits are indicated. Pre-accident data according to [16]
0 2 4 6 8 10
Years since accident
Age of animals (months)
Site 2
Site 3
Site 4
0 2 4 6 8 10
Years since accident
Embryonic lethality (%)
Site 2 Site 3+4
have not changed in these years. Therefore, and in line
with the unequivocal dose-rate dependency displayed in
Figs. 4, 5 and 7, we propose a radiogenic etiology of the
observed biological eVects.
Steady decrease of the whole-body dose rate over
the monitoring period of 1986-1996
At all sites investigated, whole-body absorbed dose rates
were found to decrease over about one order of magnitude
in 10 years, which is attributable to the gradual disappearance
of the radioactive fall-out material from the
biota [9]. More speciWcally, we show that the whole-body
dose rate decreased approximately according to an exponential
function with a half-life of 2.5–3 years (Fig. 3b,
d). As the observed biological damage did not decrease
in the same animals over the studied period, it is important
to exclude that the exposure levels in the later years
of the investigation, due to long-lived _-emitting transuranic
radionuclides and/or increasing 90Sr concentrations
in the studied populations, were underestimated.
Our analyses of radionuclide concentrations in animals
showed that 10 years after the accident the contribution
of transuranic elements and 90Sr to the whole-body dose
rate in the studied specimens of bank vole did not exceed
3 and 8%, respectively [9]. This contribution was estimated
to be two orders of magnitude lower than the
whole-body dose rates during the year of the accident [9].
The contribution of 90Sr to local doses in bone tissue in
the later years of the investigation was about 2–10£ lower than that of external _-irradiation. The absorbed
dose rate in the red bone marrow due to 90Sr _-radiation,
which is approximately equal to the absorbed dose rates
averaged over the bone tissues, was of the same small
magnitude, comparable to that of 137+134Cs _-radiation
in the later years and an order of magnitude lower than
_-radiation of 137+134Cs in early period after the accident.
We conclude that doses at late time points were not
underestimated by our approach.
Dose dependence of the observed eVects
Data presented in Tables 2 and 3, as well as in Figs. 5
and 7, hint at a strong dependence of the observed biological
eVects on radiation exposure. This is manifested
by relatively low levels of aberrations and, where present,
embryonic losses at the less contaminated sites 1 and 2,
the medium levels at sites 3 and 4, and the relatively high
level at the most contaminated site 5. Furthermore, there
is a statistically signiWcant linear correlation between
mean values of the chromosome aberration frequency
observed 10 years post-accident and the mean wholebody
dose rates prevailing in the studied animals at the
same time (Fig. 4). A similar association of mean
micronuclei frequencies with exposure at low dose rates
(4.22–39.4 _Gy/day) has been shown in bank vole populations
studied 2 years after the accident at four Swedish
regions contaminated by Chernobyl fallout at deposition
levels of 1.8, 22, 90, and 145 kBq/m2 [17], i.e., levels similar
to those at our monitoring sites 1–3 (Table 1). Mean
levels of chromosome aberrations were also increased in
a dose-dependent manner in laboratory mice, exposed in
our monitoring areas during 4 months to whole-body
dose rates from external and internal irradiation of 3–
145 _Gy/day [8], i.e., at exposure levels close to those in
free-living animals.
Dose dependence of the biological damage is a basic
requirement for conclusions concerning the causal role
of radiation exposure for the observed eVects. This
includes the possibility of direct as well as indirect modes
of action, in other words, the induction of the eVects seen
in a given generation by radiation exposure of the same
or of the preceding generations of animals. In such a
way, a linear correlation observed between mean values
of chromosome aberration frequencies and the dose
rates in late generations of animals 10 years after the
accident can as well demonstrate a linear correlation of
the aberration frequencies with the dose rates in all previous
years, because the dose rates at the various sites
remained in constant proportions to each other, while
they decreased in an exponential fashion over the monitoring
period (Fig. 4). On the other hand, the scatter
plots of the individual frequency of chromosome aberrations
versus the individual level of the whole-body
absorbed dose rate or whole-body dose may hint at a
dose dependence of the biological eVect even on the
exposure of the individual animal (Fig. 2 and data not
shown). An analogous correlation has been observed
between the individual frequencies of micronuclei and
the individual radiation exposures in the same population
of animals [6]. These dose-eVect relationships should
be studied in future.
Long-term persistence of biological eVects and possible
mechanism of their transgenerational transmission
Our data (Tables 2, 3; Figs. 5, 7) suggest that biological
eVects in bank voles living in regions contaminated by
the Chernobyl accident persisted on a stable level (chromosome
aberrations) or even increased (embryonic
loses) over 10 years, corresponding to at least 22 animal
generations, although the dose rates have substantially
decreased during this period. This observation is in fundamental
contrast to the assumption that the biological
injury would gradually disappear in direct connection
with the exponential reduction of the whole-body dose
rate. Thus, the biological eVects cannot be explained
alone by the exposure experienced by the individual animals.
Rather, our observations call for a diVerent interpretation,
namely that the biological eVects reXect a
transgenerational transmission of radiation damage that
occurred in early generations, especially in the generations
living at the time of the accident and in the early
post-accidental generations, and was further accumulated
in the course of the chronic exposure of the following
generations. The fact that oVspring from captured
females, irradiated in utero at contaminated areas, but
born in the laboratory and fed with uncontaminated
food, did not show a signiWcant reduction in aberration
frequencies in comparison with animals born and grown
up in the contaminated environment (Fig. 6) corroborates
the hypothesis that some mechanism of transgenerational
transmission of radiation damage has been
The occurrence of radiation damage in cells and tissues
that have not been directly exposed to radiation is
not without precedence. For example, genomic instability
is a long-term radiation eVect that has been intensively
investigated. The term “genomic instability” refers
to the de novo production of delayed responses of the
genome in the progeny of the exposed cells in a cell culture
or in an irradiated animal [18–22]. Genomic instability,
including chromosomal instability in hemopoietic
stem cells in vitro and in vivo, has been shown to be activated
by high- and low-LET ionizing radiations at relatively
low doses [21–25] and to exhibit a dose-dependent
increase at higher doses [24]. Dose dependence was also
observed for chromosomal instability in mice fetuses
irradiated as zygotes [26]. Other authors did, however,
not Wnd a dose relationship for genomic instability over
a wide range of radiation doses [27–29]. The expression
of genomic instability was described to persist for up to
or even more than 70–80 population doublings at higher
doses [29, 30]. In the progeny of long-term repopulating
haemopoietic stem cells of mice, chromosomal instability
was observed over 24 months after cell irradiation and
transplantation [22]. There is evidence that the genetic
background or genotype can inXuence radiation-induced
instability [25]. The mechanisms of genomic instability
are largely unknown. At least part of them may rely on
epigenetic transmission of information. For instance,
genomic instability was suggested to reXect altered signal
transduction pathways and changes in the DNA microenvironment
[31, 32]. Since DNA methylation is linked
to chromosome condensation [33], chromosome and
chromatid fragments as well as Robertsonian translocations
may be due explained by the radiation-induced
increased fragility of the chromatin during chromosome
condensation [34]. These types of chromosome aberrations
have been suggested as hallmarks of induced genomic
instability, although other types of chromosome
aberration can also occur [24, 26, 27, 30].
However, in contrast to studies where long-term
eVects are investigated in the same animals, which had
been exposed to radiation at an earlier stage of individual
development, the interpretation of the eVects seen in
our study must deal with damage persisting over many
generations of animals. In the literature, the term “transgenerational
transmission of radiation damage” has
mainly been used in a descriptive sense, including genetic
and epigenetic as well as hitherto uncharacterized mechanisms
of transmission damage to the progeny [21, 35–
42]. Delayed radiation-induced responses attributable to
genomic instability were revealed in the Wrst few animal
generations after irradiation of male germ cells [43] or
zygotes [37]. Heritable tumors and anomalies were
observed in descendants of exposed mice [35], and physiological
or developmental disorders were seen in the
progenies of irradiated parents, which resulted in embryonic
and early postnatal death, fertility disturbances,
congenital abnormalities or malformations [44]. Genomic
instability in the Wrst-generation oVspring (F1) of
irradiated mice or rats covers diverse endpoints such as
chromosome aberrations, micronuclei, point mutations
and reversions in somatic cells (reviewed in [39]) as well
as early (pre-implantation) and late (post-implantation)
embryonic resorption [37], reduced fertilization rate [45],
increased number of sterile females [37] and reduction in
the proliferative ability of both F1 and F2 embryonic
cells [36] and F1 liver cells [38]. Transgenerational germline
instability was also demonstrated in the oVspring of
irradiated mice (reviewed in [39]), as well as in children
born from irradiated parents [46] and in barn swallows
breeding not far (25–50 km) from the Chernobyl nuclear
power plant [47]. In addition, transgenerational transmission
of genomic and developmental eVects was
reported after parental exposure to certain chemicals
[e.g., 35, 41].
In conclusion, we suggest that in the studied bank vole
populations the radiation exposure of the parental generations
has lead to an accumulated pool of germline mutations
and/or of epigenetic changes, which resulted in the
observed, persistently elevated levels of chromosome aberrations
in somatic cells and in increased embryonic losses
in later generations. With regard to the continuous buildup
of transgenerationally-transmitted damage in the
course of chronic radiation exposure of the parental generations,
we will shortly speak of the transgenerational
accumulation of transmitted biological damage.
The present report is the Wrst in which the time course
of biological damage in mammals chronically exposed to
ionizing radiation over a series of generations has been
studied. So far, the long-term development of transgenerationally
transmitted radiation damage has been studied
in a restricted number of consecutive generations of animals,
generally in the Wrst 1–2 generations after irradiation
of parents [21, 31, 35–39, 43, 46]. Increased
frequencies of chromosome aberrations were also
observed in bone marrow cells after 25–30 generations of
red vole [3] and after 75–80 generations of common vole
[48] chronically irradiated in regions of heavy radioactive
contamination in the Urals as well as in bordering areas.
The latter observation was explained by hereditary chromosome
instability [48], but there is no data on the quantitative
time course of these eVects. The question arises
whether the levels of chromosomal aberrations and
embryonic lethality that in our study were reached in
1996 after 22 animal generations, will be maintained permanently,
or whether a phase of reduction will follow in
subsequent animal generations. Our study indicates that
a further reduction of chronic radiation exposure has to
be anticipated, and that there exists an eVective selection
mechanism, the dose-dependent increased embryonic
lethality in animal populations in radiocontaminated
areas. Further research will be required to clarify this
genetically and ecologically interesting point.
Acknowledgments This work was performed within the framework
of the State Program of the Republic of Belarus for minimizing and
overcoming consequences of the Chernobyl Accident (1986–2002).
The authors wish to thank the administrations of the Chernobyl
Exclusion Zone and the Berezinsky Biosphere Reserve for access to
the zone and the reserve. We are indebted to former and present staV
of the Antimutagenesis Laboratory, the Institute of Genetics and
Cytology, NAS of Belarus, for enthusiastic Weld and laboratory
assistance. We are grateful to Dr. M. Malko, the Institute of Physical
and Chemical Radiation Problems, NAS of Belarus, for the recommendations
in dose rate assessment. The authors especially
acknowledge Prof. D. Harder, University of Göttingen, for critically
reviewing the manuscript and helpful discussions that signiWcantly
improved this work.
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