Introduction
The year 2011 has been declared the International Year of Chemistry. This is also
the 100th anniversary of awarding the Nobel Prize in Chemistry to Maria Curie-Skłodowska.
Maria Skłodowska, by marriage Curie, born and brought up in Poland, is one of the
best known and most distinguished scientists. She is a patroness of many institutions,
schools, scientific establishments, and universities not only at home but also abroad.
Among them, Maria Curie-Skłodowska University occupies an important position.
To bring to mind the circumstances and reasons for its foundation, one should go 67 years
into the past. It was 1944 and the Second World War was still in progress. However,
in July that of year, Lublin became free from German occupation. Liberated Lublin
attracted the intellectuals who had survived the turmoil of the war. Among the scientists
who arrived in Lublin then was Henryk Raabe, the zoologist, a lecturer at Jagiellonian
University and a professor at Lviv University, who on behalf of the scientific circle
proposed the government should open a new university in Lublin. The first state university
in Lublin, Maria Curie-Skłodowska University, was set up and started its activities
on 23 October 1944 [1].
It can be presumed that the choice of the Lublin University patroness was made for
two reasons. The world scientific prestige of the distinguished Pole was to highlight
the status of the newly opened university. However, the reason might be the close
family relationships of Maria Skłodowska with Lublin and the Lublin region. Her grandfather,
Jozef Skłodowski, was the headmaster of the Province School in Lublin.
The first rector of the university was Henryk Raabe, who contributed greatly to its
organization. The grand inauguration of the first academic year was held on 14 January
1945. Then, the new academic year was inaugurated on 23 October [1].
Since that time, Maria Curie-Skłodowska University in Lublin has been developing its
own structure for meeting the academic standards and requirements resulting from changing
scientific and social actuality. Presently, it includes ten faculties: arts; biology
and earth sciences; economics; philosophy and sociology; humanities; mathematics,
physics and computer science; pedagogy and psychology; law and administration; political
science; as well as chemistry [2].
Situated in the center of the campus is the monument of the patroness, which watches
over Maria Curie-Skłodowska University. It was unveiled in the centenary of the great
chemist's birth. Thus, special attention is paid here to chemistry, which was inherent
in her university from the very beginning although in various evolving institutional
forms. Presently, in the Faculty of Chemistry about 280 people are employed, including
150 university teachers (43 hold the position of professor or lecturer) and the achievements
of student education and research staff have resulted in about 3,500 graduates with
the Master title in chemistry and environment protection, 246 scientists with the
Doctor title in chemical science, and 78 members of habilitation studies who were
granted the habilitation doctor title [3, 4].
The current organizational structure of the faculty is as follows:
Department of Adsorption
Department of Planar Chromatography
Department of Physicochemistry of Solid Surfaces
Department of Interfacial Phenomena
Department of Analytical Chemistry and Instrumental Analysis
Department of Inorganic Chemistry
Department of General and Coordination Chemistry
Department of Organic Chemistry
Department of Polymer Chemistry
Department of Environmental Chemistry
Department of Chemical Technology
Department of Theoretical Chemistry
Department of Chemical Education
Department of Crystallography
Department of Chromatographic Methods
Department for the Modeling of Physicochemical Processes
Department of Optical Fiber Technology
Department of Radiochemistry and Colloid Chemistry
In the Department of Radiochemistry and Colloid Chemistry investigations are carried
out which are a continuation of the research done by Maria Curie, who after 1918 concentrated
intensively on radiochemistry, that is, the chemistry of the radioactive elements:
polonium, actinium, radium, and thorium isotopes.
The research of the department is colloid-chemistry-oriented; particularly issues
connected with generally understood surface physicochemistry, the ionic theory of
double layers, and the stability of dispersed systems.
The other group of issues is widely understood environment protection, particularly
radiochemical monitoring. These investigations started in 1986 after the nuclear power
station disaster in Chernobyl. Presently, they include monitoring and study of the
mechanism of radionuclide migration and accumulation in various areas with particular
regard to adsorption of radioactive isotopes in soil, plants, bottom and alluvial
sediments, as well as permanent monitoring of air and research into the presence of
indoor radon.
Radiotracer and radiochemical researches
The first radioisotope laboratory in the Faculty of Chemistry was created in 1969.
At the beginning it was intended to carry out various experiments in the field of
physical chemistry using radioactive isotopes as tracers. These studies allowed the
establishment of the reaction mechanisms of fine solid particle surfaces (mainly minerals)
in connection with their flotation behavior [5].
Intensive study of flotation systems using radioisotope tracers and the development
of didactic activity caused a transformation of the Isotope Laboratory into the Department
of Radiochemistry and Application of Radioisotopes in 1974. The name of our department
was changed again in 1983 when the additional scientific issues were introduced. The
name Department of Radiochemistry and Colloid Chemistry has been used up till now.
At first, our department was equipped with a few Geiger–Müller and scintillation counters,
which enabled beta and gamma radiation measurements of the radioactive tracers used.
These days, the scientific interest of our department is devoted to the physicochemistry
of mineral enrichment by flotation, clarifying waste suspensions by the flocculation
method, and dewatering of sediments using the agglomeration technique.
In particular, through the application of various radionuclides and isotope dilution
techniques, studies on sorption mechanisms of flotation collectors have been performed.
The selectivity of the chosen collectors was tested for various sulfide minerals such
as galena, sphalerite, pyrite, marcasite, and chalcocite [6–10] and oxidized surface
minerals such as calcite and zinc carbonates [11].
Adsorption of macromolecules onto the mineral surface with the aim to establish the
stability of fine particle suspensions of minerals such as calcite, titanium dioxide,
and hematite in the presence of polymer aqueous solutions was also a subject of our
studies [12–17].
Simultaneously, applied research on dewatering of waste mineral suspensions (such
as coal slimes from coal mines or remains after floatation enrichment of minerals)
by means of the flocculation method in connection with the spherical agglomeration
process was carried out [18–21]
Intensive development of radiochemical methods of analysis and determination of radionuclides
in the environment proceeded after the Chernobyl disaster, which took place in 1986.
It is marked as the greatest emergency on a seven-measure scale worked out by the
International Atomic Energy Agency [22]. Erroneous decisions of operators together
with unfavorable parameters of this type of work for such reactors led to damage of
the reactor core. This resulted in the destruction of the reactor and release of large
quantities of radioactive materials into the atmosphere, as well as the radioactive
contamination of large regions, particularly in Europe.
The total release of radioactive substances was about 1.4 × 1019 Bq (as of 26 April
1986), which included 1.8 × 1018 Bq of 131I, 8.5 × 1016 Bq of 137Cs and other cesium
isotopes, 1 × 1016 Bq of 90Sr, and 3 × 1015 Bq of plutonium radioisotopes. Also noble
gases were released in large amounts [23]. Most of the strontium and plutonium isotopes
were deposited within a 100-km zone around the reactor [23]. The most important isotopes
are 137Cs, 90Sr, 238,239+240Pu, and 241Am [23], which have a rather long half-life.
The most airborne isotopes, such as 131I and 137Cs, contaminated a large region of
Europe, from which the deposition was highly heterogeneous. A characteristic trait
of the Chernobyl contamination was the appearance of so-called hot spots, meaning
fragments of the reactor core of several micrometers in diameter, containing highly
radioactive isotopes, such as europium and promethium, and also plutonium, americium,
and curium [24].
The Lublin region (eastern Poland) was at that time subjected to high radionuclide
concentrations, which enabled the values to be monitored in a relatively easy way.
Our department systematically enlarged its measuring base by installing new spectrometric
equipment for radionuclide determination of the environmental concentrations. As the
Chernobyl fallout was formed of short-lived isotopes to a large extent, the observed
radioactivity rather quickly came back to its “normal level.” Artificial radioactivity
present up till now is a result of the presence of 137Cs, 90Sr, and plutonium. This
is because these radionuclides have long half-lives.
In the course of the 24 years which have passed since the Chernobyl event, many multifaceted
studies have been carried out, and a summary of the effects of the catastrophe has
been presented in many reports by the IAEA and the Nuclear Energy Agency [23, 25].
In Poland, studies were also carried out on different issues related to the appearance
and spreading of Chernobyl contamination. Such studies were also conducted in the
Department of Radiochemistry and Colloid Chemistry. A brief summary of our results
which reflected the current view of Chernobyl contamination in the Lublin region was
presented in [26].
Presently, our department has two gamma spectrometers equipped with high-purity germanium
detectors and the program Genie 2000 (Canberra), four alpha spectrometers with planar
silicon detectors with a passivated surface (PIPS) with the same software, a liquid
scintillation counter for labeled samples, and a Quantulus ultra-low-level liquid
scintillation spectrometer. There is also equipment for radon measurements (Pico-Rad
detectors) as well as determination of the concentration of radon and thoron daughters
(automatic radon thoron daughters meter from Silena).
At the beginning our study on the presence of radionuclides in the environment focused
on defining a radiological state of the environment—determination of the contamination
level of soil, river, and lake sediments, as well as ground-level air with anthropogenic
isotopes such as 137Cs, 90Sr, and plutonium isotopes [27–37]. The occurrence of plutonium
isotopes in ground-level air [38–40] and indoor radon [41, 42] as well as radon in
bottled mineral waters [43] was studied.
As a result of various phenomena occurring in the environment, the concentration of
anthropogenic radionuclides is still diminishing. Therefore, as a continuation of
the study on the radiological state of the environment, the study of radionuclide
transport in soil profiles was undertaken [44–48].
Determination of lower concentrations of anthropogenic radionuclides demands, besides
sufficiently sensitive equipment, also proper analytical procedures. Such procedures
are necessary for determination of alpha-emitting isotopes (238Pu, 239+240Pu) and
beta-emitting isotopes (90Sr/90Y and 241Pu). In the Department of Radiochemistry and
Colloid Chemistry, proper procedures were developed for radiochemical separation of
the above-mentioned isotopes from environmental samples, where they are present at
a very low level of activity. In this way, it was possible, for example, to measure
the concentration of 90Sr, 239+240Pu, and beta-emitting 241Pu in subsequent layers
of the soil profiles. This enabled us to calculate a vertical migration rate of the
isotopes [46, 48–52].
The development of new and fast radioanalytical methods for the determination of radionuclides,
especially those which are potentially hazardous in the case of uncontrolled radionuclide
emission into the environment, is still a subject of our study. One of our recent
interests is research on the method of determination of beta-radiating 241Pu. Plutonium
is one of the most widely spread man-made radioactive elements in the environment.
Its presence is a result of atmospheric nuclear tests being carried out intensively
about 50 years ago. One of the possible methods of plutonium activity determination
is using liquid scintillation spectrometry. This technique enables both alpha radiation
measurements with very high efficiency and beta radiation measurements with efficiency
dependent on energy. In all cases quenching is an important factor. With use of liquid
scintillation spectrometry it is possible to determine alpha-radiating 238Pu, 239Pu,
and 240Pu as well as beta-emitting 241Pu.
Plutonium determination in the environmental sample requires radiochemical treatment
of the sample. The best method for this purpose is solvent extraction. Plutonium can
be extracted from a matrix consisting of nonradioactive substances and, which is especially
important, from radioactive uranium and thorium (present usually at relatively high
concentrations). Selection of a suitable extracting agent with a high extraction yield
and possible low quenching effect is very important. These questions are being investigated
[53].
Owing to the radioanalytical methods described earlier, it was possible to determine
plutonium isotopes in particular fractions extracted from soil and waste samples by
the sequential extraction method. This method, named fractionation, allowed the determination
of the bioavailability of radionuclides of interest despite their very low concentration
not enabling speciation determination (as an ionic form or an oxidation state). Alpha-radiating
plutonium isotopes, their determination, and their bioavailability in soils were subjects
of our studies [54, 55].
Plutonium isotopes are present in the environment in a very low concentration and
are not a hazard for human health. They are rather strongly bound with the surface
soil layer and can be transferred to some degree into plants or can be resuspended
in ground-level air. A chemical form of the isotope in soil is a main factor of its
bioavailability. Because of the low concentration, it is difficult to determine speciation
of plutonium. However, the fractionation method gives us information about the extent
of plutonium binding with a specified fraction of defined characteristics. In this
way, it is possible to estimate plutonium mobility in soil [56, 57].
Applying the sequential extraction method for the samples of surface soil from the
Lublin region, we extracted five different fractions: easily available, exchangeable
and carbonate bound, iron and manganese oxide bound, organic matter bound, and insoluble
(not available). In all extracted fractions, plutonium isotopes were determined and
their percentages calculated.
Several years ago a new research field was set up. It concerns the subject of radioecology
with respect to birds which are in relations with other mammals. Birds can sometimes
be treated as bioindicators of the environment. It is also relatively easy to collect
proper experimental materials from dead birds. These birds are the victims of collisions
with power lines, aerial masts, or high buildings in cities, and are brought to licensed
veterinary clinics, where attempts at treatment are made. When and where treatment
is not possible, or failed, injured birds are put to sleep to reduce their suffering.
However, the period of time that these birds spend in veterinary clinics is no longer
than 1 week. In a few cases, skeletons used in research were found during field surveys,
thus at the place of death of a given bird.
Examination of the contamination of the bodies of birds and correlation of the results
with their nestling places, migration routes, and feeding habits allowed us to assess
the role of birds in contaminant transfer and to prove their role as environment bioindicators
[58–60].
Modern spectrometers for radiation measurement which make use of the liquid scintillation
technique are widely applied to determine beta-emitting radionuclides, especially
those of low-energy radiation. A Quantulus (Wallac-PerkinElmer) spectrometer, owing
to sophisticated electronics (amplitude comparator and anticoincidence systems) being
an active system of background reduction, and passive shielding made of a large mass
of lead, enabling very low activity of beta radiation to be measured, are used to
determine beta-emitting radioisotopes occurring in the environment [61].
Preparation of environmental samples for alpha and beta radioisotope determination
usually requires the application of multistage radiochemical separation procedures
for isolation of the radionuclide to be determined. This is also the case when liquid
scintillation spectrometry is used as a tool for radioactivity measurement. The final
step of the radiochemical procedure usually involves solvent extraction of the radionuclide
of interest. A sample obtained in this way is directly introduced into a scintillation
cocktail. However, insufficient separation from contaminants present as matrix components
causes a lowering of scintillation efficiency (so-called quenching). For this reason,
the optimization of measurement conditions is an important factor in the determination
of specific activity. Quenching is the main disadvantageous feature of liquid scintillation
counting. The level of quenching is influenced by the chemical composition of the
scintillation cocktail (whose precise composition is usually not known), the chemical
composition of the sample, and the sample-to-cocktail volume ratio. There are many
different scintillation cocktails available on the market, and they differ in their
chemical composition and dedicated application to a given kind of sample [61]. Producers
suggest for which kind of sample their cocktails are suitable. If one tries to utilize
the cocktails for conditions other than those suggested, it is necessary to check
what changes in measurement parameters this can introduce. These can cause various
responses to quenching. Therefore, a study on optimization of measurement conditions
of selected radionuclides using a number of commercial scintillation cocktails and
various quenching levels was carried out [62, 63].
The aim of this paper is to present a history of radiochemistry research in the Department
of Radiochemistry and Colloid Chemistry at Maria Curie-Skłodowska University. The
results of our studies in the fields of radiochemistry, radioecology, and radioanalytical
methods and summarized data concerning the anthropogenic radioisotope contamination
of the area of eastern Poland are now presented.
Eastern Poland radioisotope contamination study
Systematic studies of the contaminated environment of eastern Poland began several
years after the Chernobyl disaster after measurement apparatus of appropriate quality
(gamma spectrometers with germanium detectors and alpha spectrometers with silicon
detectors) had been obtained. With use of this apparatus, research was carried out
on the contamination of various components of the environment.
The samples analyzed were collected according to IAEA [64] guidelines. The subjects
of the studies were noncultivated soil samples, including forest soil and bog peat
(in some cases cultivated soil was also analyzed), sediments from rivers and lakes,
specified plants (grass, cultivated plants, pharmaceutical herbal plants), and bones
and eggshells of birds.
The following anthropogenic isotopes were determined: 137Cs, 90Sr, 238Pu, 239+240Pu,
and 241Pu. Determination of each of these isotopes required appropriate preparation
of samples, including radiochemical treatment necessary for qualitative and quantitative
identification of isotopes emitting alpha and beta rays.
The measurement equipment and the procedures for preparing samples for determination
of selected radionuclides are presented next.
Equipment and method for determination of gamma-radiating 137Cs
The material collected for gamma spectrometric measurement was dried and then mechanically
crushed to a size of several millimeters in width (bones, plants) or 2 mm (sediments
and soils). Small samples prepared in this way (about 20 g) were placed on a flat
container of 50-mm diameter and about 5-mm height. Samples of appropriately large
volume (about 500 cm3) were placed in Marinelli beakers.
The samples were subjected to spectrometric measurements performed with a gamma spectrometer
(Canberra, Silena) equipped with a high-purity germanium detector of 87-cm3 active
volume, 17.5% relative efficiency, and 1.8-keV full width at half maximum (FWHM) resolution.
A quantitative analysis was performed using Canberra Genie 2000 with the goal of determining
the activity of natural and artificial radioactive isotopes.
Equipment and method for determination of alpha-radiating 238Pu and 239+240Pu
Determination of plutonium in environmental samples requires separation of the isotopes
in a pure form and fixing them on a steel plate, which enables alpha spectrometric
measurements. This was achieved by a radiochemical multistage procedure. In the first
stage of this procedure, a dried and ashed sample was leached with 6 M HCl. From the
solution obtained, trace elements were coprecipitated with iron(III) hydroxide by
ammonia addition. The precipitate was dissolved in 6 M HCl and then separated from
iron by coprecipitation with calcium oxalate. Next, the precipitate was dried, burned
in an oven, dissolved in concentrated HCl, and then a second coprecipitation with
iron(III) hydroxide was performed. In the next stage, the precipitate was dissolved
in 8 M HNO3 and, after the oxidation state of plutonium had been changed to Pu4+,
the solution was introduced onto an anion-exchange column (filled with Dowex 1 × 8)
to separate plutonium from other trace elements. In the final stage, plutonium was
eluted from the column using HCl with 0.1 M NH4I. To obtain a sample suitable for
alpha spectrometric measurement, plutonium was electrodeposited onto a stainless steel
plate. Measurements were performed using a 7401 Canberra Alpha spectrometer with a
1520 mixer-router and an S-100 multichannel analyzer. The PIPS detector of 17-keV
FWHM resolution was used. For quantitative analysis, Canberra Genie 2000 was applied.
The chemical yield was monitored by the addition of a standard 242Pu solution to the
sample.
Plutonium fractionation study
Based on literature procedures for fraction isolation by sequential extraction [56,
65], the following procedure was applied in our study. The method involved the separation
of the following fractions: (1) readily available fraction, extracted by 1 M magnesium
chloride; (2) carbonate-bound fraction, extracted by acetic buffer (1 M sodium acetate/acetic
acid); (3) hydrous iron and manganese oxide fraction, extracted by 0.04 M hydroxylamine
hydrochloride in 25% acetic acid; (4) oxidizable organically bound fraction, extracted
by 0.02 M nitric acid with hydrogen peroxide, then by 3.2 M ammonium acetate in 20%
nitric acid; (5) residual fraction extracted by boiling with 6 M hydrochloride acid.
Extracted fractions were submitted for further radiochemical separation with the aim
of purification of plutonium to measure its activity by alpha spectrometry. For this
purpose a sample of every extracted fraction was acidified with hydrochloric acid
and tracer 242Pu and ferric ions as a carrier were introduced. Next, ammonia was added
and ferric hydroxide was precipitated at pH 8. After dissolution of the precipitate
in 6 M hydrochloric acid, a small amount of calcium ions and oxalic acid was added,
then calcium oxalate was precipitated by changing the pH with ammonia. The next step
involved destruction of oxalates by ashing, dissolution in 8 M nitric acid, and coprecipitation
of plutonium with ferric hydroxide. Further separation of plutonium isotopes was performed
by the anion-exchange method and a sample for alpha spectrometric measurement was
prepared by electrodeposition, as described earlier.
Equipment and method for determination of beta-radiating 90Sr and 241Pu
Determination of beta-radioactive 90Sr in environmental samples demands isolating
it from other natural beta-emitting isotopes present in the sample materials. The
procedure for separating this isotope was worked out by Solecki [66, 67] with the
goal to determine its concentration by measurement using a liquid scintillation method.
It consists in sample ashing, leaching with concentrated nitric acid, extracting yttrium
ions from nitric acid solution (after adding the yttrium carrier) with tributyl phosphate,
precipitating yttrium hydroxide using ammonia after dilution of the organic phase
with ethanol, and purification of Y(OH)3 by dissolving the precipitate with nitric
acid and adding a few grams of Dowex 1 × 8 resin into a bulk solution. Next, yttrium
oxalate was precipitated and after it had been dissolved in nitric acid, 210Pb was
removed by coprecipitation with PbI2. From the solution, yttrium oxalate was reprecipitated
with the aim of yield determination by weight. Next, dry precipitate was suspended
in water and added to the scintillation cocktail (Insta-Gel Plus, Packard) in a low-potassium
standard glass vial. Measurements were made with the use of the Quantulus (Wallac-PerkinElmer)
ultra-low-level spectrometer for 300 min. The procedure applied allowed a minimum
detectable amount of 0.02 Bq/kg to be obtained [66].
Beta-emitting 241Pu was determined from the sample prepared for alpha spectrometric
measurement. For this purpose, the stainless steel plate with plutonium electrolytically
built up was treated with hot 4 M nitric acid to remove plutonium into solution. Next,
plutonium was extracted using 0.2 M trioctylphosphine oxide solution in cyclohexane
and the organic phase was transferred into a scintillation vial containing a Permablend
III (Packard)/toluene scintillation cocktail. Activity measurement was performed as
described earlier using the liquid scintillation spectrometry technique.
Results and discussion
In the short period of several years after the Chernobyl catastrophe, it was possible
to determine in the samples analyzed the activity of the gamma-emitting isotope 134Cs,
which has a relatively short half-life (2.06 years). The entire quantity of this isotope
was a result of Chernobyl residue. Therefore, knowing the ratio of cesium isotopes
from the Chernobyl emission to be 134Cs/137Cs = 0.528 [23], we could calculate the
participation of 137Cs coming from this fallout in the entire concentration in the
sample. The remaining quantity of 137Cs present in the environment was the result
of atomic bomb tests carried out to the greatest intensity in the 1960s. The possibility
of distinguishing the source of 137Cs (global or Chernobyl fallout) was lost at the
moment when the 134Cs activity fell below the level of minimum detection activity.
Analogously, testing the relation of the isotopic ratio 238Pu/239+240Pu in the samples,
one can identify the participation of plutonium of Chernobyl origin. The participation
of 238Pu in global fallout was about 4%; however, in the Chernobyl fallout, it was
about 50%. In the same way, one can calculate a value for the Chernobyl fraction of
241Pu in the total concentration of 241Pu in the sample [51].
The summarized results of our studies concerning determination of isotopes in various
components of the environment are gathered in Tables 1, 2, 3, 4, 5 and 6, most often
presented in the form of minimum and maximum values as well as arithmetic mean values
and the year in which the measurements were made. Table 1 concerns 90Sr determination
[49, 60, 66, 68, 69], Tables 2 and 3 concern determination of 137Cs [31, 47, 48, 60,
66, 69, 71–74], Table 4 concerns determination of alpha-emitting plutonium isotopes
[47, 48, 76], and Table 5 concerns beta-emitting 241Pu isotope [51, 77]. Table 6 presents
the vertical migration rate of plutonium isotopes [51].
Table 1
Concentration of 90Sr in the surface soil, soil profiles, and selected vegetable samples
[49, 60, 66, 68, 69]
Sample
Minimum
Maximum
Mean ± σ
Year
Surface soil
Soil of Bug River valley at the bank (Bq/kg)
5.3
69.8
29 ± 21
1999
Soil of Bug River valley at the bank (Bq/m2/10 cm)
364
12,300
320 ± 350
1999
Soil of Bug River valley (Bq/kg)
6.8
85.3
24 ± 23
1999
Soil of Bug River valley (Bq/m2/10 cm)
530
10,330
370 ± 30
1999
Bug River sediment (Bq/kg)
3.7
20.5
10 ± 5
1999
Soil profiles (Lake District, north Lublin region)
Mean concentration in arable soil layers (Bq/kg)
Profile 1
1.7
8.7
5.3 ± 2.6
2004
Profile 2
0.15
12
6.9 ± 3.5
2004
Profile 3
1.9
4.8
2.9 ± 1.1
2004
Profile 4
0.45
9.6
3.0 ± 2.8
2004
Profile 5
0.1
1.5
0.75 ± 0.50
2004
Profile 6
0.27
7.1
2.9 ± 2.3
2004
Profile 7
3.6
23
11 ± 6.3
2004
Mean concentration in noncultivated soil layers (Bq/kg)
Profile 1
0.72
8.7
3.8 ± 3.2
2005
Profile 2
2.6
14
10 ± 4.2
2005
Profile 3
0.26
4.3
1.6 ±1.6
2005
Profile 4
0.64
8.6
3.9 ± 2.8
2005
Profile 5
0.13
1.9
0.87 ± 0.69
2005
Profile 6
0.53
5.3
2.4 ± 2.0
2005
Profile 7
0.34
7.6
2.0 ± 2.6
2005
Vegetables
Potato Solanum tuberosum (Bq/kg dw)
0.09
2.4
1.0 ± 0.9
2004
Sugar beet Beta vulgaris (Bq/kg dw)
0.49
2.87
1.7 ± 1.0
2004
Carrot Daucus carota (Bq/kg dw)
0.21
0.57
0.39 ± 0.18
2004
Grass (not specified) (Bq/kg dw)
0.05
0.57
0.31 ± 0.26
2004
Birds of prey—bones (15 species, 57 individuals)
4.6
15
8.8 ± 3.4
2005
Pharmaceutical herbal plants, 17 species (mBq/kg dw)
4
1810
140 ± 120a
2002
dw dry weight
aExcluding one large result (1.8 Bq/kg)
Table 2
Concentration of 137Cs in the surface soil, profile samples, and sediments [31, 47,
48, 66, 71]
Sample
Minimum
Maximum
Mean ± σ
Year
Surface soil
Soil of Wieprz River valley at the bank (Bq/kg)
237
4,460
1,620 ± 910
1998
Soil of Wieprz River valley (Bq/kg)
588
3,730
1,750 ± 840
1998
Soil profiles (Lublin region), total activity
Brown soil (cambisol) (Bq/m2)
4,500
1998
Lessive soil (luvisol) (Bq/m2)
12,000
1998
Podzol soil (cambic podzol) (Bq/m2)
5,200
1998
Soil profiles (Lake District, north Lublin region)
Mean concentration in arable soil 0–20 cm (Bq/kg)
Profile 1
11.1
16.9
14.2 ± 2.4
2004
Profile 2
18.3
19.7
19.0 ± 0.60
2004
Profile 3
3.2
9.9
7.4 ± 2.9
2004
Profile 4
9.5
10.2
9.9 ± 0.3
2004
Profile 5
2.6
9.5
7.4 ± 3.2
2004
Profile 6
13.3
15.2
14.0 ± 1.0
2004
Profile 7
9.4
12.5
11.0 ± 1.3
2004
Mean concentration in noncultivated soil 0–20 cm (Bq/kg)
Profile 1
3.6
33.3
19.8 ± 15.7
2005
Profile 2
4.7
12.7
9.0 ± 3.3
2005
Profile 3
14.6
20.7
18.2 ± 2.9
2005
Profile 4
6.2
15.4
11.5 ± 3.8
2005
Profile 5
7.4
12.3
10.9 ± 2.7
2005
Profile 6
6.0
21.3
13.2 ± 6.3
2005
Profile 7
4.8
10
7.8 ± 2.4
2005
Lake sediments (Lake District, north Lublin region)
Piaseczno Lake (Bq/kg dw)
0.7
240
77 ± 68
1995
Masluchowskie Lake (Bq/kg dw)
3.5
290
100 ± 94
1995
Zemborzycki artificial lake (Bq/kg dw)
1.0
53
24 ± 18
1995
Wieprz River (Bq/kg dw)
6.9
1995
Table 3
Concentration of 137Cs (Bq/kg) in selected lichen, bird bones and eggshells, plant,
and peat samples [31, 60, 69, 72–74]
Sample
Minimum
Maximum
Mean ± σ
Year
Lichens (Parmeliaceae), 15 samples
2.6
108
46 ± 29
1998
Pharmaceutical herbal plants, 17 species (mBq/kg dw)
53
9850
346 ± 263a
2002
Birds of prey—bones (15 species, 57 individuals)
1.1
27
5.4 ± 6.8
2005
Birds—eggshell
Grey heron Ardea cinerea (4 samples)
4.9
16.1
10.4 ± 4.6
2003
Mute swan Cygnus olor (13 samples)
3.3
112
27.4 ± 31.3
2003
Montagu harrier Circus pygargus (6 samples)
15
40
29.4 ± 10.7
2003
Marsh harrier Circus aeruginosus (4 samples)
44
71
54.7 ± 11.5
2003
Saw sedge Cladium mariscus (Chelm, eastern Poland)
Sampling point A
56.9 ± 0.7
2005
Sampling point B
48.5 ± 0.5
2005
Peat (Chelm, eastern Poland)
Sampling point A
33.7 ± 0.9
2005
Sampling point B
53.2 ± 1.3
2005
aExcluding three very large results (3.4, 5.5, and 9.8 Bq/kg)
Table 4
Concentration of 239+240Pu (mBq/kg) in various samples and its Chernobyl fraction
[47, 48, 76]
Sample
Minimum
Maximum
Mean ± σ
Chernobyl (%)
Year
Surface soil of Lublin region (0–2 cm), 9 samples
80
340
210 ± 80
15
1993
Lake sediment
10
650
2,600 ± 100
1995
Forest soil profiles (0–45 cm) (Bq/m2)
60
335
40.4–61.2
0.7–2.9
1996
Lublin river sediments, 14 samples
70
315
50 ± 36
17
1996
Soil of Wieprz River valley at the bank
20
65
150 ± 15
2–15
1998
Soil of Wieprz River valley
10
420
160 ± 15
3–23
1998
Wieprz River sediment
10
200
40 ± 10
2–27
1998
Soil of Bug River valley at the bank
170 ± 20
2–60
1999
Bug River sediment
60 ± 5
35–70
1999
Soil of Wieprz River valley (Bq/m2)
4–21
2–15
1998
Saw sedge Cladium mariscus (Chelm, eastern Poland)
Sampling point A
46.7 ± 10.8
2005
Sampling point B
34.8 ± 23.1
2005
Peat (Chelm, eastern Poland)
Sampling point A
450 ± 78
2005
Sampling point B
361 ± 27
2005
Table 5
Concentration of 241Pu in the soil samples and its Chernobyl fraction (calculated
on the date of the Chernobyl disaster) [51, 77]
Sample
Mean ± σ
Chernobyl (%)
Year
Surface soil (0–10 cm)
Soil of Lublin region, 8 samples (Bq/kg)
0.95 ± 0.54
2000
Forest soil of Lublin region (Bq/kg)
5.4 ± 0.7
2000
Cultivated soil profiles (Lake District, north Lublin region)
Summarized fallout in the profile 0–40 cm (Bq/m2)a
Profile 1
1,040 ± 40
96.9
2007
Profile 2
610 ± 10
99.7
2007
Profile 3
530 ± 10
84.0
2007
Profile 4
460 ± 10
89.9
2007
Profile 5
390 ± 10
76.1
2007
Profile 6
320 ± 10
67.9
2007
aValues calculated in March 2007
Table 6
Mean migration rate of plutonium isotopes in soils calculated separately for the global
and Chernobyl fractions [51]
Sample
Plutonium fallout fractions
Year
Global (cm/year)
Chernobyl (cm/year)
239+240Pu
Surface soil from east Poland (0–10 cm), 40 samples
0.55 ± 0.16
1.58 ± 0.80
1996–1998
Cultivated soil (Lake District, north Lublin region)
6 soil profiles 0–20 cm
0.64 ± 0.17
1.92 ± 0.69
2006
14 soil profiles 0–15 cm
0.57 ± 0.27
2008
241Pu
Cultivated soil (Lake District, north Lublin region)
6 soil profiles 0–20 cm
0.67 ± 0.18
1.54 ± 0.65
2007
The average global fallout of 90Sr was 3,200 Bq/m2, but up to the year 1986 (when
additional amounts of this isotope were introduced into the atmosphere as a result
of the Chernobyl disaster) it was lowered to 1,500 Bq/m2 [70] as a result of the radioactive
decay. Table 1 presents the average specific activities of 90Sr in the surface soil
and sediments of Bug River valley, in soil profiles of the Lake District in the Lublin
region (noncultivated and arable soil), selected vegetables, pharmaceutical plants,
and bones of birds of prey. As can be seen, the highest 90Sr concentrations were found
in the soil of Bug River, but they are still lower than the average value of fallout.
It is interesting to note that the mean concentration of 90Sr in soil profile layers
of the Lake District is rather low and there is no marked difference between arable
and noncultivated soil. Relatively high concentrations were observed in bones of birds.
This is connected with nutrition preferences of these birds. They usually feed on
small mammals living on the ground, and these can contain more 90Sr in their bodies.
The results in Table 2 show that the radioactive fallout of 137Cs on the eastern territories
of Poland was relatively not homogeneous (large differences between the minimum and
maximum values). Global fallout is considered rather as homogeneous; therefore, the
observed nonhomogeneity comes from large participation of Chernobyl fallout in the
entire fallout of 137Cs, reaching even 100%, like in the case of the lake sediments.
Then, the mean concentration of 137Cs in the soil profile layers of the Lake District
was low in comparison with other data for the surface soil. In the case of soil profiles,
it was observed that 137Cs concentration diminishes with the depth of the profile,
as this nuclide migrates slowly down into the soil. The data presented in Table 2
are averaged values up to 20-cm depth, which corresponds to the soil layer with the
largest radiocesium concentration.
In Table 3 radiocesium concentrations in selected plants, lichens, peat, and bird
bones as well as eggshells are presented. The concentration of 137Cs in one of the
studied species of lichens from the family Parmeliaceae is about 50 Bq/kg. Similar
quantities were observed also in the samples of peat and saw sedge, taken from the
same territories of calcareous peat bogs. Lichens are considered biomonitors of contamination
in the air, since most of their vital elements are taken from the atmosphere. Therefore,
the observed contamination with 137Cs points to their resuspension character.
The studied concentration of 137Cs in bones and eggshells of the selected species
of birds shows an interesting difference: the concentration in eggshells is 2–10 times
higher than in bones. This is confirmation of the mechanism of removal of contaminants
from the bird body into the eggshell.
The 137Cs content in the eggshells of Montagu harriers, Circus pygargus, showed specific
activity 40–45 times lower in comparison with the concentrations found in saw sedge,
Cladium mariscus, the plant in which the mentioned raptors nest. The concentrations
in eggshells were also 27–43 times lower than those determined in the peat of marshes,
the areas on which this plant grows [74]. The above observations indicate that transfer
of 137Cs from the environment to the tissues of the raptors from food comprising small
mammals and insects related to the soil environment occurs to a very small extent.
The determined mean specific activity of 137Cs for all the birds studied was very
low in comparison with that for the bones of other Polish raptors studied up to the
present [75].
As can be noticed in Table 4, the average value of 239+240Pu in the soil is about
200 mBq/kg; the concentration of plutonium in lake sediments exceeds this value by
10 times; however, in river sediments it is about 4 times lower. Taking into account
the same mechanism for the transfer of radionuclides from soil to water, we can suppose
that the lack of water flow in the case of lakes does not allow for removal of radionuclides
coming into the water. In this way, plutonium entering the water after a certain time
is accumulated in the bottom sediment in quantities depending on the distribution
coefficient of the radionuclide between the water phase and the sediment, including
the subsequent chemical reactions such as complexation, coprecipitation, and radiocolloid
formation. Similar behavior was also observed in the case of 137Cs, an anthropogenic
element coming from global and Chernobyl fallout [47, 48]. Both of these elements,
being the result of human activity, are closely bound to the upper layer of soil.
Therefore, they do not easily enter the water phase, even after their long-term contact
with soil (as a result of occasional flooding of river valleys).
It is interesting to compare the quantity of Chernobyl fractions of 239+240Pu in the
samples coming from the Wieprz River and Bug River valleys, despite the relatively
large error associated with calculating this fraction. The average participation of
the Chernobyl 239+240Pu in soils and bottom sediment of the Wieprz was about 8%, and
in the case of the Bug it reached about 35%. This difference can be tied to the fact
that the Bug flows through the territories that were exposed to Chernobyl fallout
to a greater extent.
The relatively large contamination of peat with plutonium isotopes is connected with
its large surface area and high concentration of humic substance, a compound which
easily binds plutonium.
In Table 5 the concentration of beta-radioactive 241Pu in soil is presented and the
Chernobyl fraction of this isotope is shown as well. As expected, the mean concentration
of 241Pu is a few times larger than that of 239+240Pu, which is related to the isotopic
ratio of the fallout. It is interesting to see that the Chernobyl fraction of this
isotope is almost 100%. This means that, in contrast to 239+240Pu, beta-radiating
241Pu is mainly of Chernobyl origin.
Further study on the migration rate of plutonium isotopes allowed us to calculate
the velocity of vertical plutonium transport in soil profiles. The results are summarized
in Table 6. As shown, the mean migration rate is relatively low. A difference between
the Chernobyl and global origin of plutonium is also observed. As a rule, the Chernobyl
plutonium migrates 3 times more quickly than the global fallout plutonium. This is
probably connected to different chemical forms of plutonium in these two types of
fallout.
As a result of the fractionation study of plutonium isotopes in soil, it was found
that in arable soil about 45–60% of plutonium is bound with easily available and carbonate
(also available if the pH of the soil is a little lowered) fractions. Only a small
percentage of plutonium was bound permanently as an insoluble fraction. In contrast,
in noncultivated soil, most plutonium was bound with an organic fraction, more than
10% is permanently bound, and only a small percentage is available.
Conclusions
Among the main classes of environmental contaminants, radionuclides tend to be generally
less known and less frequently studied than heavy metals or organic compounds. This
is because the problem of radioactive contaminants concerned a small number of regions
to which, on account of military matters, there was limited access. However, this
changed after the disastrous explosion of the Chernobyl nuclear reactor.
As shown, global fallout contamination is still important in eastern Poland, as well
Chernobyl fallout. The anthropogenic radionuclides can still be found and determined
in the environment. The global character of both fallouts, their accumulation in natural
as well as transformed ecosystems, and their transportation through biotic and abiotic
pathways made them a subject of interest for research while taking advantage of various
models. Therefore, our department became involved in such a study and the study will
continue and be developed further.