Environmental Radioactivity in Different Climate Types: Measurement, Terrestrial Transport Process and Radiation Exposure
Vom Fachbereich für Physik und Elektrotechnik der Universität Bremen
Zur Erlangung des akademischen Grades eines Doktor der Naturwissenschaften (Dr. rer. nat.) genehmigte Dissertation
von
Ahmed Ali Husein Qwasmeh
Eingerich am: 22.05.2008 Tag des Promotionskolloquiums: 01.07.08
Gutachter: Prof. Dr. Justus Notholt Prof. Dr. Gerald Kirchner
Prüfer: Prof. Dr. Jörn Bleck-Neuhaus Prof. Dr. Stefan Bornholdt
I Table of Contents
List of Figures V
List of Tables VIII
Abstract 1
1 Environmental Radioactivity and its Measurements 4
1.1 Introduction 4 1.1.1 Natural and Man-Made Radiation 4 1.1.2 Radioactivity 5 1.1.2.1 Types of Radiation 6 1.1.2.2 Radiation Levels and Their Effects 7 1.2 Measurement of Radiation 8 1.2.1 Gamma Detection 8 1.2.1.1 Types of Gamma Detectors 10 1.2.1.2 General Characteristics of Gamma Detectors 10 1.2.1.2.1 Detector Efficiency 11 1.2.1.2.2 Detector Resolution 12 1.2.1.3 Semiconductor Detectors 13 1.2.2 Beta Measurement 17 1.2.2.1 Gas-Filled Detectors 17 1.2.2.2 Proportional Counters 18 1.2.3 Sample Geometry 23 1.2.4 Evaluation of Gamma Spectra 24 1.2.4.1 Energy Calibration 24 1.2.4.2 Efficiency Calibration 25 1.2.4.3 Counting Statistics 26 1.2.4.4 Soil Sample Spectra 28 1.3 Cesium-137 and Strontium-90 in the Environment 29 1.3.1 Cesium 29 1.3.2 Strontium 31 1.3.3 90Sr and 137Cs Sources 32 1.3.3.1 Global Fallout 32 1.3.3.2 Chernobyl Fallout 35 1.3.4 Chernobyl Impact on Jordan and its Neighboring Countries 40 1.3.4.1 Syria 41 1.3.4.2 Egypt 45 1.3.4.3 Lebanon 46 1.3.4.4 Jordan 47 1.4 Effects of Soil Characteristics on the Depth Distribution of 137Cs 49 1.4.1 Organic Matter Content 50 1.4.2 Particle Size Distribution 51 1.4.3 Cation Exchange Capacity (CEC) and K, Mg and Ca Concentrations 52 II 1.4.4 Soil pH 53
2 Radioactivity Concentrations in Jordanian Soil and Plants Samples 54
2.1 Introduction 54 2.2 Motivation and Goals 57 2.3 Sampling, Samples Locations and Identification and Sampling Preparation 58 2.3.1 Sampling Procedure 61 2.3.2 Sample Preparation 67 2.3.2.1 Preparation of Samples for Gamma Measurements 67 2.3.2.1.1 Preparation of the First Set of Samples 67 2.3.2.1.2 Preparation of the second Set of Samples 68 2.3.2.2 Radiochemical Separation to Determine 90Sr Concentrations (Beta Measurements) 69 2.3.2.2.1 Introduction 69 2.3.2.2.2 Sample Preparation 69 2.3.2.2.3 Radiochemical Separation 70 2.4 Measurements and Analysis 71 2.4.1 Gamma Analysis 71 2.4.1.1 Measuring the Activities 71 2.4.1.2 Determining and Building the Efficiencies 73 2.4.1.3 Results and Discussion of Gamma Analysis 74 2.4.1.4 Soil-to-Plant Transfer Factors 87 2.4.2 Beta Analysis 88 2.4.2.1 Measuring the Activities 88 2.4.2.2 Results of Beta Analysis 89 137 2.5 Comparison of Cs Concentrations in Soil in Jordan with Some European and Middle East Countries 91 2.6 External Dose 93
3 Depth Distribution and Migration of 137Cs in Jordanian Soils 98
3.1 Introduction 98 3.2 Soil Analysis and the Effect of its Characteristics on 137Cs Migration in Soil 98 3.3 Determining the Origin of 137Cs in the Jordanian Soils 101 3.3.1 137Cs-90Sr Ratio 101 3.3.2 Convection Dispersion Migration Model of 137Cs in Soil 104 3.3.2.1 Introduction 104 3.3.2.2 Theory 105 3.3.2.3 Results and Discussion 108 3.3.2.4 Statistical Evaluations of Fit1 and Fit2 122 3.3.2.5 Comparison with migration parameters from other studies 123 3.3.3 Correlation between the Annual Rainfall in the Sites and 137Cs Inventories (Climate Effects) 128 3.3.4 Correlation between the Sites Altitudes and 137Cs Inventories 129
III 4 Conclusions and Outlook 130
5 References 132
IV List of Figures Figure 1-1: Decay chain of 238U [Gentry 2003]...... 5 Figure 1-2: Spectrum of a radioactive source collected by germanium detector (left) and NaI(Tl) detector (right) [CANBERRA (a)]...... 11 Figure 1-3: FWHM for a peak whose shape is Gaussian...... 13 Figure 1-4: Band structure of electron energies in insulators and semiconductors...... 14 Figure 1-5: Cross-sectional view of a Ge-semiconductor detector [CANBERRA (a)].. 15 Figure 1-6: Schematic drawing of a multichannel analyzer (MCA)...... 16 Figure 1-7: Expanded view of a photo peak [CANBERRA (b)]...... 16 Figure 1-8: Gamma spectrum obtained with a 137Cs calibration source...... 16 Figure 1-9: The basic components of ionization chamber...... 17 Figure 1-10: Gas Detector Output vs. Anode Voltage...... 19 Figure 1-11: Avalanche formation by a charged particle traversing the detector gas...... 20 Figure 1-12: The solid angle (Ω) subtended by the frontal area (A) of the detector at source (S) position (D) [Knoll 1999]...... 22 Figure 1-13: 2π gas flow proportional counter [Knoll 1999]...... 22 Figure 1-14: 4π gas flow proportional counters [Knoll 1999]...... 22 Figure 1-15: Marinelli-beaker...... 24 Figure 1-16: Efficiency calibration curve for a high purity semiconductor detector...... 26 Figure 1-17: Peak and background areas for background subtraction [Gedcke]...... 28 Figure 1-18: Gamma spectrum obtained from a soil sample...... 29 Figure 1-19: Decay scheme of 137Cs [Firestone 1996]...... 30 Figure 1-20: Decay scheme of 134Cs [Firestone 1996]...... 31 Figure 1-21: Tests of nuclear weapons in the atmosphere and underground [UNSCEAR 2000; Annex C]...... 33 Figure 1-22: Annual deposition of radionuclides produced in atmospheric nuclear testingin the northern hemisphere [UNSCEAR 2000; Annex C]...... 34 Figure 1-23: The site of Chernobyl power plant and the surrounding regions [UNSCEAR 2000; Annex J]...... 36 Figure 1-24: The contamination plumes from Chernobyl and the corresponding arrival dates in the European contries [UNSCEAR 1988; Annex D]...... 37 Figure 1-25: Surface ground deposition of 137Cs in the immediate vicinity of the Chernobyl reactor in the closest zones (30 km and 60 km)of Chernobyl nuclear power plant [IAC 1991]...... 38 Figure 1-26: Surface ground deposition of 90Sr Released from Chernobyl reactor [IAC 1991]...... 39 V Figure 1-27: Soil contamination with 137Cs in the Federal Republic of Germany in 1986 according to the Department of Federal Health [Bundesgesundheitsamt 2000]...... 40 Figure 1-28: Map of Jordan and its neighbouring countries...... 41 Figure 1-29: The estimated trajectories of radioactive plume, ------, and clean air mass. -.- .-.-, air mass trajectories were constructed by Department of Meteorology in Syria using satellite photographs [Othman 1990]...... 42 Figure 1-30: The coastal Syrian mountains with the studied sites (dots) [Al-Rayyes 1999]...... 43 Figure 1-31: Mapping of 137Cs inventory in Syria[Al-Masri 2006(a)]...... 44 Figure 1-32: A comparison between total 137Cs inventory and mathematically derived nuclear bomb tests 137Cs [Al-Masri 2006(a)]...... 44 Figure 1-33: Nile Delta and the north coast [Shawky 1997]...... 45 Figure 1-34: Burullus Lake location in Egypt [El-Reefy 2006]...... 45 Figure 1-35: The map of Lebanon with locations of sampling sites [El Samad 2007]. ... 47 Figure 1-36: Jordan’s map with sample locations [Al Hamarneh 2003]...... 49 Figure 2-1: 137Cs profile in a sediment core from kinneret lake [Kirchner 1997]...... 55 Figure 2-2: Jordan's map with samples locations...... 59 Figure 2-3: Population density of Jordan...... 60 Figure 2-4: Soil sampling plans [Jacobsen]...... 61 Figure 2-5: Hand auger used in soil sampling...... 62 Figure 2-6: Plastic containers used to collect the samples...... 63 Figure 2-7: The sampling area in Kufrsum (AQ1)...... 63 Figure 2-8: The sampling area in Foua'ra (AQ2)...... 64 Figure 2-9: The sampling area in Baliela (AQ3)...... 64 Figure 2-10: The sampling area in Qafqafa (AQ4)...... 64 Figure 2-11: The sampling area in Dair Elleyyat (AQ5)...... 64 Figure 2-12: The sampling area in Abien (AQ6)...... 65 Figure 2-13: The sampling area in Aien El Basha (AQ7)...... 65 Figure 2-14: The sampling area in Wadi El Naqah (AQ8)...... 65 Figure 2-15: The sampling area in Irhab (AQ9)...... 65 Figure 2-16: The sampling area in El Ramtha (AQ10)...... 66 Figure 2-17: The sampling area in As Subeihi (AQ11)...... 66 Figure 2-18: Soil-Wax pellet...... 68 Figure 2-19: Sealed sample...... 68
VI Figure 2-20: 20 mm plastic petri-dish soil sample...... 69 Figure 2-21: Gamma spectrometry used for gamma detection...... 73 Figure 2-22: 137Cs depth profile in AQ1, AQ2, AQ3 and AQ8...... 78 Figure 2-23: 137Cs depth profile in AQ4, AQ5 and AQ6...... 78 Figure 2-24: 137Cs depth profile in AQ7, AQ9, AQ10 and AQ11...... 79 Figure 2-25: 137Cs depth profile in AQ3 and AQ3new...... 84 Figure 2-26: 137Cs depth profile in AQ4 and AQ4new...... 84 Figure 2-27: 137Cs depth profile in AQ5 and AQ5new...... 85 Figure 2-28: 137Cs depth profile in AQ6 and AQ6new...... 85 Figure 2-29: 137Cs depth profile in AQ9 and AQ9new...... 86 Figure 2-30: 137Cs depth profile in AQ10 and AQ10new...... 86 Figure 2-31: 137Cs inventories for the first and the second sets of samples...... 87 Figure 2-32: Gas-filled proportional detector of kind Low-Level-Handprobenwechsler LB 750 L, Berthold...... 89 Figure 2-33: 90Sr inventories for the first set of samples...... 90 Figure 2-34: 90Sr depth profiles for AQ4, AQ5 and AQ6...... 90 Figure 2-35: 90Sr depth profiles for AQ4 and AQ4new...... 91 Figure 3-1: The calculated inventories of 137Cs from Chernobyl and nuclear bomb tests...... 102 Figure 3-2: 137Cs depth profile in AQ3new using Fit1...... 116 Figure 3-3: 137Cs depth profile in AQ3new using Fit2...... 116 Figure 3-4: 137Cs depth profile in AQ4new using Fit1...... 117 Figure 3-5: 137Cs depth profile in AQ4new using Fit2...... 117 Figure 3-6: 137Cs depth profile in AQ5new using Fit1...... 118 Figure 3-7: 137Cs depth profile in AQ5new using Fit2...... 118 Figure 3-8: 137Cs depth profile in AQ6new using Fit1...... 119 Figure 3-9: 137Cs depth profile in AQ6new using Fit2...... 119 Figure 3-10: 137Cs depth profile in AQ9new using Fit1...... 120 Figure 3-11: 137Cs depth profile in AQ9new using Fit2...... 120 Figure 3-12: 137Cs depth profile in AQ10new using Fit1...... 121 Figure 3-13: 137Cs depth profile in AQ10new using Fit2...... 121 137 Figure 3-14: CsNB inventories vs. sites average annual rainfall...... 128 Figure 3-15: 137Cs inventories vs. sites altitudes...... 129
VII List of Tables Table 1-1: Dose rates and their effects [Hall 1984]...... 9 Table 2-1: Rainfalls in northwestern section of Jordan in May 1986 collected from the Jordanian metrological department...... 55 Table 2-2: Soil samples identification...... 60 Table 2-3: Concentrations of 137Cs, 134Cs and 40K (d.m. ≡ dry mass)...... 76 Table 2-4: A comparison with Al Hamarneh’s study (Al Hamarneh, 2003)...... 81 Table 2-5: 137Cs concentrations in plant samples...... 82 Table 2-6: Concentrations of 137Cs in the second set of samples (d.m. ≡ dry mass). .... 83 Table 2-7: Soil-to-plant transfer factors for 137Cs...... 88 Table 2-8: The Average Concentration of 137Cs in pasture Soil (0-10 cm) in Germany [BMU 2004]...... 92 Table 2-9: The deposition of 137Cs in Jordan and some Middle East and European countries...... 93 Table 2-10: Dose rate conversion factors (μGy/y per Bq/cm3) at 1m above the ground for uniform slab sources between the ground and different soil depths for 600 keV...... 94 Table 2-11: Annual effective dose equivalent at 1m above the ground...... 95 Table 2-12: Annual effective dose equivalent at 1m above the ground for Jordan and its neighboring Countries...... 96 Table 3-1: Physical and chemical proprieties of the analyzed soil samples...... 100 Table 3-2: The measured ratio of the total 137Cs to the total 90Sr and the calculated ratio of 137Cs from Chernobyl to nuclear bomb test 137Cs...... 102 Table 3-3: Chernobyl deposition of 137Cs in soils from Jordan and from some European countries...... 103 Table 3-4: The parameters of Fit1...... 113 Table 3-5: The parameters of Fit2 (using different velocities and different dispersion coefficients for Ch. and NB)...... 113
Table 3-6: CsCh-CsNB ratios resulting from different methods...... 114 Table 3-7: F test; SSE values, degrees of freedom, calculated F and P values...... 123 Table 3-8: Migration parameters of 137Cs found in this work and other works...... 126
VIII Abstract
The radionuclide 137Cs dose not exist naturally in the environment. Its main sources in the environment are the nuclear bomb tests, which took place mainly during 1954–1964, and the nuclear power plants accidents. The most severe accident was Chernobyl (26 April 1986). There is no record for 137Cs from nuclear bomb tests fallout (pre-Chernobyl fallout) in Jordan. So the main questions arising about the probable source of 137 Cs in Jordan are; What is the 137Cs contamination fraction due to nuclear bomb tests? Has Chernobyl affected Jordan? if yes, how large was its effect? How does 137Cs migrate in the Jordanian soils? Is it still available for the plants uptake? Is the presence of 137Cs in Jordanian soil a risk for public health? The current work is an effort to study the artificial radioactivity in Jordan due to 137Cs in soils to answer the above mentioned questions and to compare it with that in European countries, which have different climate types and large areas with high contamination from Chernobyl. For that task, two sets of soil samples were collected from pre-assumed undisturbed areas from the northwestern part of Jordan, where most of the population live. The first set of samples was collected in April 2004 from eleven different sites of area of about 10 m × 10 m each, comprising 67 samples in total. The second set of samples was collected in July 2005 from six of the previous sites where higher 137Cs contaminations were found and it consists of 104 samples. The soil profiles in the second set of samples were thinly sliced for a detailed study of 137Cs profiles. The second set of samples was collected from small area ( about 10 cm × 20 cm) that made it less representative as compared to the first set regarding the total inventory of 137Cs. The necessary laboratory preparations were performed before submitting the samples to gamma measurements. The 137Cs-90Sr ratio was used to find the ratio between 137Cs from Chernobyl and 137Cs 137 137 90 from the nuclear bomb tests ( CsCh− CsNB), thus a chemical separation of Sr was done using the so-called “Nitric Acid Method” before submitting it to the beta measurements. 1 Gamma measurements were done using a HPGe detector with 50% relative efficiency and 2 keV resolution at 1.33 MeV. Beta measurements were done using a gas-filled proportional detector of type Berthold Low-Level-Handprobenwechsler LB 750 L with efficiency of 21.3% cps/Bq. The specific activities of 137Cs were measured for the first and the second set of samples and the surface activities were calculated. A comparison was held between the contamination of 137Cs in the Jordanian soils and that in the neighboring countries and some countries from south, north, east, west and central Europe. The effective dose equivalent due to 137Cs in soil was calculated for the first set of samples at a height of 1 m above the soil surface in order to estimate the risk on the public health due the external irradiation. The correlations were studied between the depositions of 137Cs for the first set of samples and each of the sites average annual rainfalls and sites altitudes. The total inventories of 90Sr were measured as averages for all profiles in the first set of samples. In addition, 90Sr was measured for every layer in three profiles from the first set and one profile from the second set of samples. The mobility of 90Sr was clearly higher than that of 137Cs. In order to study the migration of 137Cs in soil, soil analysis was carried out for the second set of soil samples and two methods were applied to the measured data, namely: the 137Cs- 90Sr ratio, which was applied on the first set of samples, and a convection dispersion model, which was applied to the second set of samples. The “137Cs-90Sr ratio” method was useful to estimate a ratio between 137Cs from Chernobyl 137 137 137 and Cs from the nuclear bomb tests ( CsCh− CsNB), whereas the convection dispersion model was able to find this ratio in addition to the migration velocity and dispersion coefficient of 137Cs in soil. In this work, the convection diffusion fit was carried out in two methods, the first method (Fit1) was done assuming that the depositions from Chernobyl and from global nuclear bomb tests have the same migration velocity and the same dispersion coefficient, which is the method usually implied, while the second method (Fit2) assumes different migration
2 velocities (vCh and vNB) and different dispersion coefficients (DCh and DNB) and the possible physical justifications for this assumption were discussed. Visually, relatively more representative fits have been achieved using Fit2. Using the F- test, Fit2 was very statistically significant better than Fit1. 137 137 Comparing the results of Fit1 and Fit2, different CsCh− CsNB ratios were obtained for 137 137 four sites and very similar ratios for two sites, whereas the CsCh− CsNB ratios obtained using the “137Cs−90Sr ratio” method were significantly different from the fits results for all sites. The fit parameters have been tabulated and a comparison to the migration parameters from other studies was held. The comparison includes some important information about the soil profiles.
3 1 Environmental Radioactivity and its Measurements
1.1 Introduction
1.1.1 Natural and Man-Made Radiation
Natural sources of radiation represent the greatest part of radiation received by the world’s population [LLNL 2002]. Cosmic rays and radionuclides naturally present in our environment, such as radioactive materials in soil and water, are the sources of the natural radioactivity. The important terrestrial radionuclides are 235U, 238U, 232Th and their radioactive decay isotopes and 40K. Cosmogenic isotopes are formed due to the interaction of cosmic rays with atoms in the atmosphere, hydrosphere, or the top layers of the lithosphere. These include stable isotopes such as 3He and radioactive isotopes such as 10Be, 14C, 26Al, 36Cl, 41Ca and 129I. Cosmic radiation increases with altitude. Earth’s magnetic field diverts the radiation, which makes the level of cosmic radiation higher in the regions of the poles as compared to those in the equatorial regions. Exposure to cosmic rays depends strongly on the altitude and weakly on the latitude [UNSCEAR 2000; Annex A]. Human exposure to radiation can be classified as external exposure and internal exposure. External exposure is the type where the radiation reaches man from radioactive substances existing outside his body whereas in the internal exposure the radioactive substances exist inside the body (by inhalation or ingestion). The natural terrestrial radiation levels vary from one area to another on the earth, due to the soil and rock compositions variation. The release of the radionuclides as a result of the human activities makes what we call man-made sources of radiation. Some main sources of man-made radiation are the radionuclides used in medicine, in the nuclear power plants for energy production, nuclear weapons production, nuclear bombs testing and nuclear power plants accidents. Nuclear weapons tests in the 1950s and 1960s result in global fallout, which was precipitated mainly on the northern hemisphere of the earth. 4 1.1.2 Radioactivity
Radioactivity is the spontaneous emission of radiation (particles and energy) from a nucleus. The time needed for a radioisotope to lose half of its radioactivity is called the half- life. The radioisotopes decay exponentially with time as clear from in Eq. 1-1.
= − λ t A (t ) A o e Eq. 1-1 with ln 2 λ = Eq. 1-2 T1/ 2
λ where is the radioisotope decay constant , Ao is the initial activity, T1/ 2 is the radioisotope half-life, and A(t) is the activity at the time t, measured in disintegration per second or Becquerel (Bq) in the SI unit system. Curie (Ci) is another unit for activity, where 1 Ci is equal to 3,7 * 1010 Bq. Decay chain refers to the process when a radioisotope decays to another radioisotope, which in turns, decays further until a stable isotope is reached. An example is the decay chain of 238U (Figure 1-1), where it decays through many radioactive daughters till reaching the stable isotope 206Pb.
Figure 1-1: Decay chain of 238U [Gentry 2003].
5 1.1.2.1 Types of Radiation
I. Alpha radiation: It is a particle, which consists of two protons and two neutrons (helium nuclei). It has positive charge. Alpha particles have relatively high atomic mass of 4 and their energies range from 4 to 8 MeV. Alpha particle of 4 MeV has a range of about 2.5 cm in air and about 14 μm in tissue, whereas an alpha particle of 8 MeV has a range of about 7 cm in air and 42 μm in tissue [RSC]. Alpha-emitting substances do not represent a health danger out of the body but they are hazardous if they enter the body through breathing, eating, or drinking, where they can expose the internal tissues in the body directly. II. Beta radiation: A beta particle is an electron (β-) or a positron (β+). They have energies from a few keV to a few MeV and a mass of an electron. They are more penetrating than alpha particles. In general, an aluminum sheet a few millimeters thick is required to stop beta particles of few MeV energy [RSC]. III. Neutron radiation: Neutrons are energetic uncharged particles. They have a high ability of penetration. Neutrons are commonly produced as a result of fission processes in nuclear reactors but they can be emitted due to spontaneous fission by a few radionuclides, like 235U and 239Pu. IV. Gamma rays radiation: Gamma rays are electromagnetic radiation. Their typical energies in radioactive decay range from 0.1 to 3 MeV. Since that they have no charge, they are much more penetrating as compared to alpha and beta particles. Penetration of gamma ray in a specified material depends on its energy and on the mass attenuation coefficient of that material. V. X-rays radiation: X-rays are also electromagnetic radiation like gamma rays, but are produced outside the nucleus. X-rays have usually lower energies as compared to Gamma rays. However, in some applications, X-ray could reach high energies such as the bremsstrahlung generated by some medical linear accelerators (up to 20 MeV).
6 1.1.2.2 Radiation Levels and Their Effects
It is known that the radiations are dangerous and have adverse effects on the body tissues. They are dangerous because they can not be sensed by the body organs and their effects could appear along period of time after the irradiation [Strettan 1965]. Exposing to large doses (much larger than the background levels) increases the cancer risk. From the results of the experiments on plants and animals, it is assumed, that ionizing radiation can cause genetic mutations. High doses of radiation can cause sickness and death within a short time period of exposure. The damage caused by radiation depends on many factors; dose, dose rate, type of radiation, the part of the body exposed, age and health [Hall 1984]. Natural background radiation dose is about 2–3 mSv/y and average value of 2.4 mSv/y has been established by the United Nations Scientific Committee on the Effects of Atomic Radiation (UNSCEAR) [UNSCEAR 2000; Annex B] depending on many studies received by people all over the world. Estimating the risk of excess cancer due to low dose irradiation in humans has been the issue of many studies where it has been tried to establish critical estimations for it [e.g. ICRP 1991, UNSCEAR 1993, Muirhead 1993, Cox 1995, UNSCEAR 2000]. For radionuclides usage, the ICRP [ICRP 1991] recommendations limit the maximum dose from a single source to a member of the public to 0.3 mSv/y and an annual dose limit of 1mSv/y (in addition to the background dose) was recommended. The maximum occupational dose limit recommended by the ICRP [ICRP 1991] was 20 mSv/y averaged over five years with the further provision that it should not exceed 50 mSv in any single year. According to the ICRP, the level of fatal cancer risk associated with the 0.3 mSv dose is about 10−5 per year, whereas a level of fatal cancer risk of 10−6 per year is regarded as trivial and the corresponding annual dose of about 10–20 μSv has been regarded by the IAEA and the Council of the European Union [IAEA 1996, CEU 1996] to be the level where no consideration of individual protection is needed.
7 The cancer risk due to radiation dose is usually calculated by the linear no-threshold theory of radiation carcinogenesis, which states that the cancer risk in an irradiated population is proportional to the irradiation dose. According to this theory, any dose, no matter how small, involves the possibility of developing cancer. High-level doses of radiation are used in radiotherapy to kill cancerous cells and higher doses are used to kill harmful bacteria in food, and to sterilize bandages and other medical equipment [Hall 1984]. Eric J. Hall has listed effects and uses of the radiation doses according to their levels in a table [Hall 1984] (see Table 1-1).
1.2 Measurement of Radiation
The measurement techniques for the radiations resulted from radioactive decay are based on detection of their ionization products. As we are mainly interested in measuring 137Cs, which is a gamma emitter, a brief description of the commonly used gamma detectors and their characteristics will be given in this section. Since a high-purity germanium (HPGe) detector was used to measure the activity of 137Cs, semiconductor detectors will be described in some more details. Gas-filled detector will be also described since it has been used to detect beta emissions to determine 90Sr activity.
1.2.1 Gamma Detection
All methods to detect the charged particles or the electromagnetic radiations depend on the interaction these radiations with the matter that they traverse and to which they impart energy by the ionization or excitation of its atoms or molecules [Mann 1980]. In this section, the commonly used kinds of gamma detectors are being described as well as the main characteristics of the detectors like the efficiency and the resolution.
8 40,000-70, 000 mS (40- 70 Sievert) used in radiotherapy. 10,000 mSv (10 Sv) as a short-term and whole-body dose would cause immediate illness, such as nausea and decreased white blood cell count, and subsequent death within a few weeks. Between 2 and 10 Sv in a short-term dose would cause severe radiation sickness with increasing likelihood that this would be fatal. 1,000 mSv (1 Sievert) in a short term dose is about the threshold for causing immediate radiation sickness in a person of average physical attributes, but would be unlikely to cause death. Above 1000 mSv, severity of illness increases with dose. If doses greater than 1000 mSv occur over a long period they are less likely to have early health effects but they create a definite risk that cancer will develop many years later. Above about 100 mSv, the probability of cancer (rather than the severity of illness) increases with dose. The estimated risk of fatal cancer is 5 from every 100 persons exposed to a dose of 1000 mSv (i.e. if the normal incidence of fatal cancer were 25%, this dose would increase it to 30%). 50 mSv is, conservatively, the lowest dose at which there is any evidence of cancer being caused in adults. It is also the highest dose which is allowed by regulation in any one year of occupational exposure. Dose rates greater than 50 mSv/y arise from natural background levels in several parts of the world but do not cause any discernible harm to local populations. 20 mSv/y averaged over 5 years is the limit for radiological personnel such as employees in the nuclear industry, uranium or mineral sands miners and hospital workers (who are all closely monitored). 10 mSv/y is the maximum actual dose rate received by any Australian uranium miner. 3-5 mSv/y is the typical dose rate (above background) received by uranium miners in Australia and Canada. 3 mSv/y (approx) is the typical background radiation from natural sources in North America, including an average of almost 2 mSv/y from radon in air. 2 mSv/y (approx) is the typical background radiation from natural sources, including an average of 0.7 mSv/y from radon in air. This is close to the minimum dose received by all humans anywhere on Earth. 0.3-0.6 mSv/y is a typical range of dose rates from artificial sources of radiation, mostly medical. 0.05 mSv/y, a very small fraction of natural background radiation, is the design target for maximum radiation at the perimeter fence of a nuclear electricity generating station. In practice the actual dose is lower. Table 1-1: Dose rates and their effects [Hall 1984].
9 1.2.1.1 Types of Gamma Detectors
A gamma detector was used to measure the activity of 137Cs. The commonly used types of detectors for gamma radiation can be categorized as:
I. Gas-filled Detectors II. Scintillation Detectors III. Semiconductor Detectors
Choosing a certain detector type for an application depends on certain factors such as the X-ray or gamma energy range of interest, the efficiency and the resolution requirements of the application, the suitability of the detector for timing applications, and the price. Different types of detectors have different operating characteristics. The operation of these detectors is based on the points that can be summarized as follows [Debertin 1988]:
I. The photon converts its energy into kinetic energy of electron (and eventually positron) by photoelectric absorption, Compton scattering or pair production. II. These electrons produce electron-ion pairs, electron–hole pairs or excited molecular states. III. Collection and measurement of the charge carriers or the light emitted in the deexcitation of the molecular states.
1.2.1.2 General Characteristics of Gamma Detectors
The spectrum resulting from a source of gamma emitters is made up of groups of photons, each group being mono-energetic. The detector converts such a line spectrum into a combination of lines and continuous components (Figure 1-2). The ability of the detector to produce lines or peaks for mono-energetic photons is characterized by the peak width and the peak efficiency [Debertin 1988] (Figure 1-2).
10 Figure 1-2: Spectrum of a radioactive source collected by germanium detector (left) and NaI(Tl) detector (right) [CANBERRA (a)].
1.2.1.2.1 Detector Efficiency
The detector efficiency for a specified energy is the ratio of the number of counts that occur in the peak to the total number of photons emitted by the gamma source. Various kinds of efficiency definitions are used for gamma ray detectors:
I. Absolute Efficiency: the ratio of the number of counts produced by the detector in the peak to the number of photons emitted by the source. II. Intrinsic Efficiency: the ratio of the number of pulses ( counts) produced by the detector in the peak to the number of the photons striking the detector. III. Relative Efficiency: efficiency of one detector relative to another detector in a certain geometry and for a certain photon energy.
The density of the detector material, its atomic number Z and its volume are important factors for detector efficiency determination [Debertin 1988]. The probability that the photon will interact with detector and retain all of its energy in the detector depends on these factors. Gas detector are usually filled with methane or an argon-methane mixture (low density material). This means that they have a low efficiency. NaI(Tl) scintillation detectors have higher material density, higher Z and larger thickness (wide range of sizes), which means they have higher efficiencies. These detectors are useful to measure photons of energies up to
11 several MeV. The Si and Ge semiconductors have material densities and Z higher than those of the gas detectors and lower than those of the NaI(Tl) scintillation detectors. The material density and Z are higher for Ge than for Si.
1.2.1.2.2 Detector Resolution
The full width at half maximum (FWHM) of a single energy peak determines the resolution of the detector. The detector resolution (R) for a peak is conventionally defined as the FWHM divided by the peak centroid (H0) (see Figure 1-3). The FWHM is usually expressed in keV (Ge Detectors), or as a percentage of the energy at that point (NaI(Tl) Detectors). Higher resolution (smaller FWHM) means that the system has the ability to separate the peaks within the spectrum more clearly. The peak width depends on the energy needed to produce the charge-carrier pair. This value is about 30 eV for gas detectors and about 3 eV for semiconductor detectors. Consequently the number of charges-carriers created per photon detected is higher, which makes it possible to obtain better energy resolution with low noise. Thus, the semiconductor detectors have peaks with much less width (i.e. much higher resolution). The NaI(Tl) detectors can not be compared directly to the other two types of detectors since their operation depends on the collection of light photons rather than of charge [Debertin 1988], and the average energy needed to produce a light photon is about 100 eV. Figure 1-2 shows two spectra collected from the same source. The spectrum on the left half of the figure has been collected using a germanium detector and the one on the right half using a sodium iodide detector. It can be obviously seen in Figure 1-2, that a germanium detector has a higher ability of resolving the peaks (higher resolution), whereas the peaks presented by the sodium iodide detector are overlapping to certain degree and small peaks are not visible.
12 dN/dH
H 0 Puls Hight (H)
Figure 1-3: FWHM for a peak whose shape is Gaussian.
1.2.1.3 Semiconductor Detectors
The semiconductors are the elements of the 4th group of the periodic table. Thus the semiconductor can act as an insulator or as a conductor. Silicon and Germanium are the most widely used semiconductors. In the metallic crystals the conduction band and the valance band usually overlap at the room temperature. Therefore the electrons can migrate from the valance band to the conduction band easily even with a small amount of energy. Thus, the metals have been categorized to have high conductivity. In the case of insulators and semiconductors, the electron must cross the band-gap to reach the conduction band. Therefore, the conductivity of the insulators and the semiconductors is many orders of magnitude lower than for the metals. The band-gap is usually 5 eV or more for the insulators, where as it is considerably less for the semiconductors (Figure 1-4) [Knoll 1999]. Figure 1-5 shows a cross-sectional view of a Germanium semiconductor detector. The incident photons, that hit the semiconductor crystal, interact within the depletion region producing the charge carriers (holes and electrons). The respective electrodes collect the charge carriers.
13 Figure 1-4: Band structure of electron energies in insulators and semiconductors.
The interaction of the incident photons can undergo three processes: photo effect, Compton scattering and pair production. The incident photons undergo the photo effect process transfer all of their energies to electrons. Therefore the pulses resulting from this process are specific for the emitting nuclei. Unfortunately, the photons, which undergo Compton scattering, transfer only a part of their energies to electrons. Therefore the pulse that results from this process is unspecific for the emitting nucleus. The resultant charge is converted to a voltage pulse amplified by a charge sensitive preamplifier and then by the main amplifier. The amplitude of every pulse is proportional to the original energy of the corresponding incident photon. The amplified pulses will be converted to digital information (signals) by the analog-to- digital converter (ADC). These digital signals are then transferred to a multichannel analyzer (MCA) (see Figure 1-6). The MCA is a device with a digital memory that consists of several thousands of channels. Every channel is specified to store digital data, which correspond to a specified pulse voltage, which in turns corresponds to a specified energy value. A peak of specified energy is registered in some channels (see Figure 1-7). The MCA has usually a monitor as an output, where the registered spectrum can been seen such as the spectrum shown in Figure 1-8 for a 137Cs calibration source. The peak of 137Cs in the center of the spectrum is usually called photo-peak since it is generated by photons undergoing photo
14 effect, whereas the region at the left of the peak is generated by photons undergoing Compton scattering and backscattering. Compton edge is maximum kinetic energy that an electron can receive from a photon (for 137Cs is about 477 keV), which take place when the gamma ray scattering angle is 180° Compton plateau results due to the photons scattering through an angle less than 180° and thus receive less energy than the Compton edge. The backscatter peak results from the photons scattered into the detector crystal by shielding, holders, etc.
Figure 1-5: Cross-sectional view of a Ge-semiconductor detector [CANBERRA (a)].
Semiconductor detectors must be cooled in order to reduce the thermal charge carrier generation (and associated noise) to an acceptable level [Lumb 2006]. Liquid nitrogen , which has a temperature of about 77 °K, is commonly used to cool germanium detectors with a large cryogenic container attached to the detector [Schery 2001], however, since the 1990s electromechanical coolers were available, but they were all mains powered, heavy and expensive [Keyser].
15 Figure 1-6: Schematic drawing of a multichannel analyzer (MCA).
Figure 1-7: Expanded view of a photo peak [CANBERRA (b)].
Figure 1-8: Gamma spectrum obtained with a 137Cs calibration source.
16 1.2.2 Beta Measurement
For the purpose of beta measurements (measuring the activity of 90Sr), a gas-filled proportional counter was used.
1.2.2.1 Gas-Filled Detectors
A gas-filled detector is basically a chamber filled with a pure gas and has insulated electrodes, an anode and cathode. An electric field can be applied across the gas by means of these electrodes. An incident radiation passing through the gas ionizes the neutral molecules along its path, which produces free electrons and positive ions. The electrons are attracted to the anode wire and collected to produce an electric pulse. The basic components of an ionization chamber are illustrated in Figure 1-9.
I Ammeter
V
Figure 1-9: The basic components of ionization chamber.
At low applied voltage, recombination may occur between electrons and ions when the electric field is insufficient to prevent this. Recombination is also possible for a high density of ions. In the case of the recombination the charge collected by the electrodes is less than the original produced ion pairs. By increasing the applied voltage, recombination is suppressed and the number of the collected ion pairs increases till a certain value of voltage where nearly all ion pairs are collected. The number of the collected ion pairs stays constant 17 after that over a range of the applied voltage. This operation region is known as the ionization chamber region. The number of the collected ion pairs in this region is an accurate measure for the ion pairs formation rate. Increasing the voltage accelerates more electrons toward the anode at energies high enough to ionize other atoms, thus creating a larger number of ion pairs (ion pairs multiplication). This operation region is known as proportional region and the detector working in this region is known as proportional counter. Over this region the ion pairs multiplication is mostly linear and the number of the collected charge is proportional to the original number of ion pairs created by the incident radiation [Knoll 1999]. Increasing the applied voltage further introduces nonlinear effects. The most important of these is related to the fact that the ions move toward the cathode slower than electrons. Therefore each pulse within the counter creates a cloud of positive ions. If the concentration of these ions is high enough, they represent a space charge, which alters the shape of the electric field. In this region the ion pairs multiplication is not any more linearly proportional to the applied voltage. This region is known as the region of limited proportionality. By making the applied voltage sufficiently high, the space charge, created by the positive ions, will be enough to terminate producing the ion pairs. This means the same number of ions will be created regardless the number of the original ion pairs created by the incident photons. Thus, the output pulses have the same amplitude and don not reflect the properties of the incident radiation [Knoll 1999]. This operation region is known as the Geiger-Mueller region and the detector works at this region as the Geiger-Mueller detector. The different operation regions, for alpha and beta particles, are illustrated in Figure 1-10. The actual voltages can vary widely from detector to another and depending up on the detector geometry and the type and pressure of the used gas.
1.2.2.2 Proportional Counters
Proportional counters are gas-filled detectors that operate in the proportional region. What happens in this region, is the amplification (multiplication) of the original number of the ion pairs created by the incident radiation (Figure 1-11). Therefore the resulting pulses are considerably larger than those from ionization chambers operated under the same conditions.
18 Therefore, these counters are commonly used for detecting beta and alpha particles and for measuring low energy x-ray radiation.
Figure 1-10: Gas Detector Output vs. Anode Voltage. (http://felix.physics.sunysb.edu/~allen/252/PHY251_Geiger.html)
Ion pair creation and gas multiplication is illustrated in Figure 1-11. In this simple Figure, a charged particle is traversing the gas producing four primary ion pairs and consequently four avalanches (usually many more ion pairs are produced by incident radiation). These four avalanches here contribute to a single pulse. Since the pulse size depends on the energy of the incident particles, one can distinguish between the pulses produced by alpha particles and the pulses produced by betas or gamma rays. Since alpha particles have much higher mass as compared to beta particles and higher electrical charge, an alpha particle is more ionizing than a beta particle has the same energy. It is illustrated in Figure 1-10 that alpha particles produce larger pulses than those produced by beta particles. Actually the size of the pulse depends also on the operating voltage.
19 Incident RADIATION A valanche
Figure 1-11: Avalanche formation by a charged particle traversing the detector gas. (http://www.orau.org/ptp/collection/proportional%20counters/introprops.htm)
There should be no electronegative components in the detector gas, since in the presence of electronegative gas, such as oxygen, the electrons may attach to the neutral molecules of that gas forming negative ions [Mann 1988]. Consequently, a negative ion goes to the anode rather than an electron. This ion usually fails to make further ionizations like an electron. Thus the produced pulse is very small compared to one produced by an electron, and it is probable that this pulse may not be able to exceed the threshold setting to be counted. Usually a noble gas is used as fill gas in a proportional counter gas for two reasons: 1) Noble gases are not electronegative and 2) Noble gases do not react chemically with the detector components. Multiplication in the proportional counter is based on the collisions between electrons and neutral gas molecules forming the secondary ionization. These collisions may also produce simple excitation of the gas molecule without creation of a secondary electron. De- excitations take place the emissions of visible or ultraviolet photons, which, in turns, could create additional ionization elsewhere in the fill gas or could produce electrons due to interactions at the wall of the counter. This can lead to a loss of proportionality and/or spurious pulses. In order to solve this problem, a small amount of polyatomic gas, such as methane, has to be added to the fill gas. These polyatomic gases, or often called the quench gases, absorbs the de-excitation photons in order to stop further ionizations [Knoll 1999]. Argon is the most widely used noble gas because of its low costs, and a mixture of 90% argon and 10% methane (P-10 gas) is the most common gas used in the gas proportional detectors [Knoll 1999]. For some applications, where the higher efficiency is required, the
20 heavier noble gases xenon and krypton are used for detecting higher energy X-rays or gamma rays. Gas flow proportional counters can be categorized into two geometries according to the solid angle (Ω) subtended by the detector at source position (see Figure 1-12) [Knoll 1999]:
I. 2π gas flow proportional counters: Figure 1-13 (Figure p165) shows geometry of 2π gas flow proportional counter with a hemispherical volume and loop anode wire. The effective solid angle is very close to 2π because any photon emerging from the surface of the source finds its way into the active volume of the counter. Therefore the detector can have an efficiency that is close to maximum possible efficiency for sources in which the radiation emerges from one surface only. A 2π gas flow proportional counter was used in measuring 90Sr activity in our samples. II. 4π gas flow proportional counters (Figure p 167): Figure 1-14 (Figure p 167) shows geometry of 4π gas flow proportional counter that is used to detect radiations that emerge from both surfaces of the sample. Such detectors provide a higher counting efficiency than the 2π counters.
21 Figure 1-12: The solid angle (Ω) subtended by the frontal area (A) of the detector at source (S) position (D) [Knoll 1999].
Figure 1-13: 2π gas flow proportional counter [Knoll 1999].
Figure 1-14: 4π gas flow proportional counters [Knoll 1999].
22 1.2.3 Sample Geometry
The maximum effective solid angle is obtained for geometries in which the source surrounded the detector (like the Marinelli-beaker in the case of the Germanium spectrometry) or when the detector surrounds the source (like the sources measured in the 4π proportional counters). This kind of geometries is preferable or sometimes necessary in case of the low-level activity measurements. For this purpose the volume of large samples can be reduced by ashing or evaporating. The optimum shape of the source material (or the beaker) depends on the detector itself and the available amount of the material. The beaker should fit into the detector housing. The calculations should be done on the dimensions of the sample in order to minimize the self- attenuation and the average distance between the radioactive material and the detector. Marinelli-beaker (Figure 1-15) presents the optimum geometry for large materials quantities in gamma spectroscopy [Debertin 1988]. The sample geometry depends on sample properties, such as sample density, filling height and chemical composition of the sample, and on the sample holder properties, such as the diameter of the sample holder, bottom and sidewall thickness and density and composition of the holder material. For low-level activity sources, a high amount of the source should be used to achieve a good count rate. Increasing the amount of a material increases the count rate but on the other hand the additional material will be farer away from the detector, which makes the modification in the count rate not significant after a certain material size. The influence of the sample geometry for a Petri-dish vial has been studied [Bossus 1998], where it was found that variations in the sample properties have a much more significant influence on the full peak counting efficiency, than the variations in the sample holder properties (the vial).
23 Figure 1-15: Marinelli-beaker.
1.2.4 Evaluation of Gamma Spectra
In order to evaluate the activity of the radionuclides in the spectrum, energy and efficiency calibrations have to be performed for the spectrometer.
1.2.4.1 Energy Calibration
The energy calibration specifies a relationship between channel numbers and the corresponding gamma energies in the spectrum. Channel numbers are proportional to pulse height if we assume that we have linear amplifiers. A spectrum of a calibration source with several known gamma energies has to be recorded. Using Eq. 1-3 the energy calibration can then be performed using a least squares fit.
E = a + b ⋅ N Eq. 1-3 where E is the gamma energy corresponds to the channel number N with proportionality factor b and an offset a. Energy calibration for gamma detectors is often done using 152Eu since it has many gamma lines including the low energy range. In our measurements three energy lines, at least, and their corresponding channel numbers were chosen from the spectrum of the analyzed sample in order to perform the energy calibration.
24 1.2.4.2 Efficiency Calibration
Efficiency calibration is important due to geometrical reasons, self-absorption and coincidence effect. The efficiency depends also on the detector properties. Due to the first two reasons, only some of the emitted photons will reach the detector. The coincidence effect takes place when we have an isotope that emits multiple cascade gamma rays in its decay. For example, if a cascade decay took place from an initial state by emitting a gamma ray (γ1) to an intermediate state and then by (γ2) to the ground state, the two gamma rays (γ1 and γ2) will be emitted in coincidence if the half-life of the intermediate state was short enough. If the time for interaction of these two gamma rays was short compared with the response time of the detector or the resolving time of the following electronics, the will be registered as one gamma line of energy equals to their energies summation and “sum coincidence peak” will be observed [Knoll 1999]. This also will lead to a loss in the individual full-energy peaks that could build by γ1 or γ2. Moreover, depending on the detector properties only a part of these incident photons interacts with the detector, and few of them undergo the photo effect. The efficiency also strongly depends on the photon energy, i.e. for the same detector and the same sample geometry; the efficiencies are different for different energies. The efficiency calibration is highly time-consuming procedure, as it should be done for a large number of samples with different materials and different geometries. Well known amounts of different radioisotopes have to be used to perform the efficiency calibration. Eq. 1-4 is used to calculate the absolute peak efficiency.
N ε(E) = Eq. 1-4 t ⋅ f ⋅ A where N is the net count of a full energy peak corresponding to the gamma photons with energy E and gamma yield f, A is the activity of the source and t the counting time. A computer program has to be used to fit the points of the different measured energies to get an efficiency calibration curve. An efficiency calibration curve for one of the semiconductor detector in our lab is shown in Figure 1-16. In this calibration measurement, a solution of 11 radionuclides has been used.
25 These lines have been marked on the curve using numbers (1 to 11) and they correspond to the radionuclides 109Cd, 57Co, 139Ce, 203Hg, 113Sn, 85Sr, 137Cs, 88Y, 60Co (lines 9 an 10), 88Y, respectively. This efficiency calibration has been done for a soil sample with density of 1.5 and 1-liter marinelli beaker. The fit for these points was done using a Silena-Gamma-Plus 1.020 software using two polynomial functions, the first covering the energies 0 - 210 keV and the second covers the energies from 210 keV to more than 2000 keV. The polynomial coefficients are shown in the figure (Figure 1-16).
Figure 1-16: Efficiency calibration curve for a high purity semiconductor detector.
1.2.4.3 Counting Statistics
Gamma spectrometry counting is the counting of individual events. Therefore the counting statistics normally meet the conditions of a Poisson distribution [Jenkins 1981, Bevington 1992]. The Poisson distribution is adequate for the statistical calculations for a peak with zero background. In the general case (with background), subtraction of the background is required to determine the net peak counts.
26 For the purposes of calculating the standard deviations, the Gaussian distribution is an adequate approximation to the Poisson distribution for mean counting values more than or equal to 9. To find the net peak counts, the background lying under the peak has to be subtracted.
Figure 1-17 shows a simple method for this subtraction, where ηP is width of the chosen region of interest (ROI) marked on the peak and NT is the total integrated area that includes the background counts (B) in addition to the net peak counts (P).
= + NT P B Eq. 1-5
As shown in (Figure 1-17) two additional regions of interest are considered in order to estimate the background (B). These regions of interest are at a distance d to the left and to the right of the peak ROI and each has a width of ηB/2. The estimated value of the background under the peak becomes
η B ≈ P (N + N ) Eq. 1-6 η B1 B2 B where NB1 and NB2 are the counts integrated under the left and the right backgrounds respectively. with a standard deviation
η σ = P B Eq. 1-7 B η B
Using Eq. 1-5, the estimated net peak counts is
= − P NT B Eq. 1-8 with a standard deviation
σ = σ 2 + σ 2 Eq. 1-9 P NT B which simplifies to
27 η σ = P + (1 + P ) B Eq. 1-10 P η B
Figure 1-17: Peak and background areas for background subtraction [Gedcke].
1.2.4.4 Soil Sample Spectra
Soil and rocks cover the upper layer of the earth. The source of the natural radionuclides in soil and rocks comes from the earth's crust where they have been present since the earths creation. The decay of the primordial radionuclides such as 235U, 238U, 232Th and 40K represents the dominant part of the natural radioactivity in soil. Natural radioactivity in soil represents a significant part of the background radiation exposure for the population. In addition to the natural radionuclides many artificial radionuclides can be found in soil especially 137Cs, which resulted mainly from nuclear bomb tests (global fallout) and/or nuclear reactors accidents like Chernobyl accident. 137Cs has a long half life (about 30.17 y) and generally migrates slowly in soil. Figure 1-18 shows a spectrum of one of the soil samples brought from Jordan. This soil sample represents a 4-5 cm slice, where the 137Cs line can be clearly seen.
28 Figure 1-18: Gamma spectrum obtained from a soil sample.
1.3 Cesium-137 and Strontium-90 in the Environment
1.3.1 Cesium
Cesium is a silvery gold element with atomic number of 55 and a relative atomic mass of 132.9. Its melting point is 28.44 °C (301.59 °K), and it has a boiling point of 671°C (944 °K). It has 37 known isotopes: Cs-112 to Cs-148 [Pfennig 1995]. Cesium belongs to the alkali metals and is least inert and thus most reactive element in this group. Also it is highly explosive when it comes in contact with water. Since Cesium belongs to the first group in periodic table as potassium, its chemical and metabolic-physiological reactions are similar to those of potassium [DAVIS 1963], which is essential for many organisms and is enriched intracellularly. Potassium cannot be replaced by Cesium in its metabolic functions and organisms usually take up cesium in different proportion as potassium [Kornberg 1961]. This could be attributed to the difference in the ionic radii, where cesium has larger radius. Cesium has only one stable isotope (133Cs), which occurs naturally in soil and rocks mainly in the mineral pollucite (hydrated silicate of aluminum and cesium) with concentrations up to 30%, whereas its main artificial sources are the nuclear bomb tests and nuclear power plants accidents.
29 The 137Cs isotope is a fission product. It is formed with relatively large amounts in nuclear bomb explosions (fission yield of 5.57%) [UNSCEAR 2000; Annex C]. The significance of 137Cs is due to its relatively long physical halflife (30.17 y), which means that it remains in the environment for a long time, whereas it has a relatively short biological half life in man, which is about 110 days [ICRP 1989]. The decay scheme of 137Cs is presented in Figure 1-19; it decays into the stable element 137Ba. This decay takes place directly, by emitting beta particles, with a branching ratio of about 5.6% and indirectly via the metastable 137mBa with a branching ratio of about 94.4%. The decay to the metastable 137mBa includes a release of energy of 513 keV as a beta particle. The physical halflife of the 137mBa is 2.55 min, which then decays into the stable 137Ba releasing gamma ray of energy 661.66 keV.
Figure 1-19: Decay scheme of 137Cs [Firestone 1996].
From the radiological point of view, the isotope 134Cs is less important than 137Cs, since it has much shorter physical halflife (2.062 y). 134Cs is formed mostly by the neutron activation of 133Cs and the yield of 134Cs in fission is negligible. In a power reactor, there is enough time to produce 133Cs due the decay of other isotopes and subsequently to produce 134Cs by activation. 134Cs resulted from the nuclear bomb tests, which mainly took place at 1961-1965, decayed to levels below the detection limits by the time of Chernobyl accident. Thus, 134Cs can be found practically only in Chernobyl fallout [Cigna 1971]. Therefore, the presence of 134Cs in soil samples has been used as indicator for Chernobyl fallout. 30 The activity ratio 137Cs/134Cs from Chernobyl (at 1986) was about 2:1 [UNSCEAR 2000; Annex J]. This ratio was useful to distinguish between 137Cs from Chernobyl fallout and 137Cs from nuclear bomb tests. The decay scheme of 134Cs is presented in Figure 1-20, where it decays into the stable elements 134Ba and 134Xe. The scheme shows the gamma lines of 134Cs decay and the branching ratio of every line. The lines 604.67 keV and 795.56 keV have very high branching ratios (97.56 and 85.44, respectively). Thus they can be seen clearly in environmental samples that contain 134Cs.
Figure 1-20: Decay scheme of 134Cs [Firestone 1996].
1.3.2 Strontium
Strontium is a silvery white or yellowish metallic element with atomic number 38 and a relative atomic mass of 87.6. Its melting point is 777 °C (1050 °K), and it has a boiling point of 1382 °C (1655 °K). It has 31 known isotopes: Sr-73 to Sr-102 [Pfennig 1995]. Strontium occurs naturally in some minerals such as celestine and strontianite. The 90Sr isotope has no natural source and it was introduced into the environment as a result of the above ground nuclear weapons tests and as a result of the nuclear power plant accidents like Chernobyl. It is a pure beta emitter with average energy 195.8 keV and a half-life of 28.6 y. It decays to the beta emitter 90Y, which has a half-life of 64.1 h [Kocher 1977].
31 It is a fission product with a fission yield of 3.50% in the nuclear bomb explosions, [UNSCEAR 2000; Annex C]. Since Strontium has much higher boiling temperature than Cesium, the 90Sr isotope is considered as a non-volatile element from Chernobyl. Therefore, only small amounts of 90Sr have contaminated the neighboring countries in comparison to 137Cs. For example, the 137Cs- 90Sr ratio was 90 in grass samples from southern Bavaria [Bunzl 1990], 159 in rainwater samples collected from Munich in May 1986 [BMU 1988], 77 in air filters from Mainz [Denschlag 1987] and between 50 and 250 in air filters in Krakow [Broda 1986, Florkowski 1987]. This value was lower (about 14) in rainwater samples collected from Bremen in May 1986 [BMU 1988]. Moreover, this ratio varied strongly in the places close to Chernobyl NPP (2.1 to 55)
[IAEA 1991]. [J. W. Mietelski, 1] In contrast, the average value of this ratio is about 1.5 in the northern hemisphere from nuclear weapons testing fallout [UNSCEAR 2000; Annex C]. Since strontium belongs to the second group in periodic table as calcium, its chemical and metabolic-physiological reactions are similar to those of calcium and the mineral substance of bones preferentially takes it up, where it has a biological mean life time of about 50 y in this tissue.
1.3.3 90Sr and 137Cs Sources
Both 137Cs and 90Sr do not occur naturally in the environment. They are exclusively anthropogenic in origin. Atomic bomb testing and Chernobyl fallouts are the main sources of them in the environment.
1.3.3.1 Global Fallout
Through atomic bomb testing since 1945, radioactive fission products have artificially entered and spread worldwide throughout the atmosphere. Large yield test programs took place during 1954-1958 and 1961-1962, whereas individual tests have occurred since 1964 [UNSCEAR 1982; Annex E].
32 Nuclear bomb testing in the atmosphere was the most significant man-made source of radiation exposure of the world population [UNSCEAR 2000; Annex C]. The number of the atmospheric nuclear tests was reported in [UNSCEAR 2000; Annex C], to be 543 tests, and the total yield was 440 Mt. The highest total explosive yields were in the years 1954, 1958, 1961 and 1962 (Figure 1- 21) [UNSCEAR 2000; Annex C]. United States [DOE 1994], the former Soviet Union [MRFAE 1996], the United Kingdom [Johnston 1994], and France [Doury 1996] have published within the last few years information on atmospheric nuclear tests. These information includes the date, the name, the location, the type, the purpose, and the total explosive yield of each test. Each test produces about 200 fission products of which many are not identifiable due to their very short physical halflife.
Figure 1-21: Tests of nuclear weapons in the atmosphere and underground [UNSCEAR 2000; Annex C].
The radioactive fission products from the nuclear bomb testing in the atmosphere were transferred mainly into the stratosphere and a minor part into the troposphere where they precipitated by rainfall on the earth surface later on.
33 When the tests were taking place, the world average deposition densities of radionuclides produced in atmospheric testing of several short-lived radionuclides, especially 144Ce, 106Ru, and 95Zr, were highest, but since 1965, 137Cs and 90Sr dominated in the residual cumulative deposit because of there relatively higher half-lives [UNSCEAR 2000; Annex C]. Strontium-90 has been measured in surface air routinely at a number of locations around the world. In the years 1957 to 1962, the United States Naval Research Laboratory has established a global surface-air monitoring network for these measurements [Lockhart 1964]. This has continued in the years 1963 to 1983 by the Environmental Measurements Laboratory of the United States Department of Energy [Feely 1985]. After 1983, the activity levels were undetectable with the methods used, thus the concentrations of 90Sr in the air were derived from averaging the results of several sites in the mid-latitudes of both hemispheres.
160 Cs137 140 90Sr
120
100
80
60
Deposition (PBq) 40
20
0 1945 1948 1951 1954 1957 1960 1963 1966 1969 1972 1975 1978 1981 1984 1987 1990 1993 1996 1999
Figure 1-22: Annual deposition of radionuclides produced in atmospheric nuclear testingin the northern hemisphere [UNSCEAR 2000; Annex C].
Since 137Cs and 90Sr have similar half-lives (30.07 y and 28.78 y, respectively), and because the deposition occurs according to the ratio of fission yields and (inversely) half- lives, the ratio of 137Cs/90Sr is 1.5 [UNSCEAR 2000; Annex C]. Thus, 137Cs deposition was estimated for the period 1958 to 1985 using this ratio and the measured 90Sr deposition. The
34 depositions of 137Cs and 90Sr are listed in [UNSCEAR 2000; Annex C] and shown for the northern hemisphere (Figure 1-22).
1.3.3.2 Chernobyl Fallout
The accident of Chernobyl occurred at 26 April 1986 at the Chernobyl nuclear power plant in Ukraine about 20 km south of the border with Belarus (Figure 1-23). This accident was the most severe in the history of the nuclear industry [UNSCEAR 2000; Annex J]. A large amount of radioactive material released from the reactor due to that accident. This amount was estimated with 1-2 EBq and the release took place over a period of 10 days and
the largest was in the first day with 25% of the total release [UNSCEAR 1988; Annex D]. (EBq=Exa Bq = 1018) About 10-20% of the volatile radionuclides iodine, cesium and tellurium and 3-6% of other less volatile radionuclides such as strontium, plutonium, cerium etc., were estimated to be released [UNSCEAR 1988; Annex D]. The radionuclides 131I and 137Cs are the most important radionuclides from the radiological point of view because they are responsible for most of the radiation exposure received by the general population [UNSCEAR 2000; Annex J]. The radioactive plume released from the reactor exceeded 1200 m altitude on 27th of April, with maximum radiation at 600 m [Izrael 1987(a)]. The contaminated zones were generally classified into two zones: the near zone (<100 km) and the far zone (from 100 km to approximately 2,000 km). The volatile radionuclides cesium and iodine were detectable at higher altitudes (6-9 km) [Jaworowski 1988]. These elements were more widely dispersed into the far zone, whereas the refractory elements (elements that vaporize at high temperatures) such as zirconium, cerium, neptunium and strontium were mostly deposited within the former USSR (in the near zone) [Izrael 1987(b)].
35 Figure 1-23: The site of Chernobyl power plant and the surrounding regions [UNSCEAR 2000; Annex J].
The release over 10 days and the changes in the wind directions at different altitudes resulted in a very complex dispersion pattern for the contamination plumes over Europe. A highly simplified pattern of these plumes, with their reported initial arrival dates in the European countries, is shown in Figure 1-24. Ground contamination was found to a certain extent in every country of the northern hemisphere [UNSCEAR 1988; Annex D]. The radionuclide 137Cs was chosen as a reference for the ground contamination from Chernobyl due to its substantial contribution to the lifetime effective dose, its long half-life time, and because it is easy to measure [UNSCEAR 2000; Annex J].
36 Figure 1-24: The contamination plumes from Chernobyl and the corresponding arrival dates in the European contries [UNSCEAR 1988; Annex D].
The ground contamination with 137Cs was inhomogeneous because of the variations of the rainfalls at the time the plume passed above considering that the wet deposition was more effective than the dry deposition. The highest soil contamination with 137Cs was in Belarus, the Russian Federation and Ukraine. Figure 1-25 shows the highest contaminated areas with 137Cs and the closest zones (30 km and 60 km zones), which were the highest contaminated. Figure 1-26 shows the surface depositions of 90Sr, where it was mostly deposited in the near zone.
37 Figure 1-25: Surface ground deposition of 137Cs in the immediate vicinity of the Chernobyl reactor in the closest zones (30 km and 60 km)of Chernobyl nuclear power plant [IAC 1991].
The highest deposition of 137Cs in Europe outside the former USSR was recorded in Sweden north of Stockholm (85 kBq/m2), the region of Tessin in Switzerland (43 kBq/m2), southern Bavaria in Germany (up to 45 kBq/m2), Salzburg (up to 60 kBq/m2) and Carinthia (33 kBq/m2) in Austria [UNSCEAR 1988; Annex D]. Figure 1-27 shows the ground surface activity of 137Cs in Germany in 1986. It is clear in the map that the deposition of 137Cs was much higher in southern Germany than elsewhere in the country. Relatively small values of 137Cs deposition have been recorded in Japan (16-300 Bq/m2), Canada (20-40 Bq/m2) and USA (20-90 Bq/m2) [UNSCEAR 1988; Annex D].
38 Figure 1-26: Surface ground deposition of 90Sr Released from Chernobyl reactor [IAC 1991].
39 Figure 1-27: Soil contamination with 137Cs in the Federal Republic of Germany in 1986 according to the Department of Federal Health [Bundesgesundheitsamt 2000].
1.3.4 Chernobyl Impact on Jordan and its Neighboring Countries
Whilst deposition and radioecological behavior of the Chernobyl fallout is quite well documented in Central and Eastern Europe, information about the area of Jordan and its neighboring countries (Figure 1-28), though affected as well, were scarce. Some research has been done and published about artificial radioactivity in Jordan [Al Hamarneh 2003], Syria [Othman 1990, Al-Rayyes 1998, Al-Masri 2006(a), Al-Masri 2006(b)], Egypt [Shawky 1997,
40 El-Reefy 2006] and Lebanon [El Samad 2007]. In the following, a summary will be given about these works to show how these countries were affected by Chernobyl.
Figure 1-28: Map of Jordan and its neighbouring countries.
1.3.4.1 Syria
An early paper was published about the impact of Chernobyl on Syria by Othman [Othman 1990]. This paper shows the arrival of the first air masses carrying radioactivity from the Chernobyl region Figure 1-29, which entered Syria early in the morning of 7 May 1986. This figure shows two trajectories, one of them represents the radioactive air mass and the other represents the clean air mass arrived at Syria on the evening of 10 May 1986. The exposure rate at 1 m above the ground surface was higher than normal exposure rate (7-12 μSvh-1) in some Syrian cities during the period 7-10 May 1986, where it was 64 μSvh-1 in Damascus and 94 μSv h-1 in A1eppo in the northern part of the country. The radionuclides 131I, 137Cs, 106Ru and 144Ce were detected in air samples, where the average concentration of 131I was about 4 Bq/m3 and the concentration of 137Cs ranged from 0.48 to 0.12 Bq/m3 between Damascus and Aleppo. The highest concentration derived from surface soil samples were 1500 Bq/m2 for 131I and 200 Bq/m2 for 137Cs during mid May 1986.
41 Figure 1-29: The estimated trajectories of radioactive plume, ------, and clean air mass. -.-.-.-, air mass trajectories were constructed by Department of Meteorology in Syria using satellite photographs [Othman 1990].
A study was carried out, in 1995, on 137Cs, 134Cs and 90Sr contamination in the coastal Syrian mountains by Al-Rayyes and Mamish [Al-Rayyes 1999]. Soil samples were collected from 15 sites and mostly from areas under trees (Figure 1-30).
42 The samples were collected using a stainless steel sheet and each sample was divided into different depth layers (0-2, 2-5, 5-10, 10- 20, 20-35). The dry weight activity concentrations in the upper 5 cm layer of the soil ranged from 500 Bq/kg to 8000 Bq/kg for 137Cs, 15 to 230 Bq/kg for 134Cs and 34 to 235 Bq/kg for 90Sr. The ratio 137Cs/134Cs ranged between 35 and 45 in August 1995, which was comparable to the expected value (about 37). The expected value was calculated depending on the initial deposition ratio in 1986 reported by Hotzl et al. [Hotzl 1987] and was 1.75. This suggests that the major contribution of 137Cs in the studied samples could be attributed to the
Chernobyl fallout. Figure 1-30: The coastal Syrian mountains with the studied sites (dots) [Al-Rayyes 1999].
Another work was recently done by Al-Masri [Al-Masri 2006(a)] where a geographical map of 137Cs inventories was executed for Syria. In this study, soil samples were collected from 36 sites distributed all over Syria, during the period of 2000-2003, to study vertical distribution and inventories of 137Cs. The total inventory (bomb tests and Chernobyl) of 137Cs varied between 320 Bq/m2 and 9647 Bq/m2, where the highest concentrations were found in the coastal, middle and northeast regions of Syria, suggesting that Chernobyl contribution is predominant. The concentrations of 137Cs were the lowest in the southeast region (Syrian Badia) with relatively uniform distribution, which may be attributed to the global nuclear bomb test fallout. Using a surface mapping system (Surfer Software, V 7), a geographical map of total 137Cs inventories has been executed (See Figure 1-31)
43 The total inventory of 137Cs from Chernobyl and from nuclear bomb tests fallout were estimated using a mathematical software developed by Walling D. E. and He Q. [Walling 1997] (see Figure 1-32). Two models were used in Walling’s software for describing 137Cs distribution in undisturbed soils. One of them was an exponential model and the other was a convection dispersion model.
Figure 1-31: Mapping of 137Cs inventory in Syria[Al-Masri 2006(a)].
Figure 1-32: A comparison between total 137Cs inventory and mathematically derived nuclear bomb tests 137Cs [Al-Masri 2006(a)]. 44 1.3.4.2 Egypt
Shawky and El-Tahawy [Shawky 1997] published a study about 137Cs and 90Sr in the Nile delta and the adjacent regions. Sixty samples covering that area, with 25 cm depth each, were collected in 1988 (see Figure 1-33). The surface layer of the Delta is mostly cultivated and composed principally of deposits from the sedimentation processes by the Nile river. The inventories of 137Cs ranged between 18.5 and 2175 Bq/m2 and between 234 and 3129 Bq/m2 for 90Sr. The authors compared these values with the accumulated base-line in soils (at depth 30 cm) of U.K at 1982 [Cawse1985], which ranged between 780 and 7770 Bq/m2 for 137Cs. Based on this comparison, they suggested that the contribution of 137Cs from Chernobyl, if there is any, is limited. Recently, a new work has been done about the Burullus Lake (Figure 1-34) [El-Reefy 2006], which is located on the coastal part of the north-central and northwest of the Nile delta. It is a shallow, saline lagoon containing numerous (~50) islands and islets. Samples have been collected from 7 sites on its northern coast and 7 sites on three islands. Each sample was composed from 5 cores and was taken from a flat area ≥ 20 m2. Each core was 30 cm depth and was divided into 3 layers, each of 10 cm.
Figure 1-33: Nile Delta and the north coast Figure 1-34: Burullus Lake location in Egypt [Shawky 1997]. [El-Reefy 2006].
45 The concentrations of 137Cs in the soil samples were measured for the upper layers (< 10 cm) and the mean value was 1.2 Bq/kg in the coast and 15.1 Bq/kg in the islands. The higher concentration of 137Cs in the islands has been attributed to the accumulation of radionuclides derived from sea-to-land transfer.
1.3.4.3 Lebanon
Recently, El Samad et al. [El Samad 2007] published a study about the Chernobyl impact on Lebanon. In this study, more than 90 soil samples were collected in the period 1998-2000 from 90 uncultivated sites uniformly distributed all over the country (see Figure 1-35). The samples were collected using a stainless steel template from 0-3 cm for the first layer, 3-8 cm for the second layer and whenever possible a third layer 8-15 cm was collected. The concentrations of 137Cs in soil ranged between 2805 Bq/m2 and 6545 Bq/m2. The surface contamination in the superficial layer (0-3 cm) ranged between 825 and 6545 Bq/m2 with an average of 3266 Bq/m2. The concentrations of 137Cs in soils in North Lebanon and in Mount-Lebanon were higher than those from South Lebanon and were within the average range of 137Cs reported in Europe due to the Chernobyl accident. The depth distribution of 137Cs in soil showed an exponential decrease. External annual effective doses due to 137Cs in soil were estimated and ranged from 19.3 to 91.6 μSv/y.
46 Figure 1-35: The map of Lebanon with locations of sampling sites [El Samad 2007].
1.3.4.4 Jordan
In our knowledge, only one paper has discussed the issue of Chernobyl impact on Jordan and the 137Cs contamination there. This was the study executed in 2000 by Al Hamarneh I. et al. [Al Hamarneh 2003].
47 Thirty-two surface and core soil samples and one moss sample were collected in November 2000 from undisturbed areas in 21 sites all over Jordan (see Figure 1-36). Most of the collected samples were taken from the top 2 cm layer of the soil only, and some down to 5 cm, 7 cm, 17 cm, 22 cm, 27 cm and 32 cm. The concentration of 137Cs in topsoil layers (0–2 cm) ranged from 7.5 to 576 Bq/kg dry weight. Two abnormally high values (352 and 576 Bq/kg dry weight) were found in the top layers (0–2 cm) in two different samples taken from one site. 134Cs was found only in these two samples with activity of range 1.5 and 2.6 Bq/kg dry mass. In general the northwest area of Jordan was higher contaminated of 137Cs as compared to east and south of Jordan. Activities of 90Sr were measured for 5 surface samples (0-2 cm), one sample 2-7 cm and one moss sample. They range between 2.8 and 11.4 Bq/kg with an average of 6.2 ± 1.2 Bq/kg, which was believed by the authors to be in the range of 90Sr in central Europe as a consequence of Chernobyl accident. Activity ratios of 134Cs/137Cs, 90Sr/137Cs had mean values of 0.0049, 0.29, respectively. The moss sample was taken because it can function as bio-accumulator for fission products like 134Cs and 137Cs. The concentrations of 134Cs and 137Cs in the moss sample were 5 and 808 Bq/kg, respectively. The estimations of the effective dose equivalent due to 137Cs in soil ranged between 3.8 and 214.2 μSv/y with an average of 60.4 μSv/y.
48 Figure 1-36: Jordan’s map with sample locations [Al Hamarneh 2003].
1.4 Effects of Soil Characteristics on the Depth Distribution of 137Cs
The chemical and physical properties of soil, like the total contents of organic matter, soil pH number, the soil composition, cation exchange capacity (CEC) and the concentrations of the exchangeable cations like potassium (K), magnesium (Mg) and calcium, are found to effect the migration of 137Cs in soil. On the other hand, the migration velocity was not found to be significantly correlated with any of the soil parameters as in the case of a study of the global fallout in south Patagonia [Schuller 2004].
49 In the following, the effects of these soil properties will be described as they were demonstrated in different studies.
1.4.1 Organic Matter Content
Organic matter (or humic substances) are composed mainly of carbon (C), hydrogen (H) and oxygen (O) with minor quantities of N, S, P and other elements. The most common humic substances in soil are humic acid and fulvic acid. Some studies give evidences, which support the assumption that 137Cs mobility in soil is higher in the presence of higher organic matter contents, such as the studies described below:
• It has been suggested that the organic matter contents affect the migration of the radionuclides in the environment [Staunton 2002]. A possible explanation for this is that the soils of high organic content do not contain enough clay minerals, which are known for very strong 137Cs adsorption. • It was indicated by Szerbin et al. [Szerbin 1999] that cesium is immobilized rapidly in soils containing less organic matter content. • It has been found that the sorption capacity of cesium in soil increases after the removal (thermal or chemical removal) of soil organic matter [BONDAR 2003]. • Low or moderate organic matter contents in soil (<40 %) have sufficient clay mineral content to fix cesium strongly while the soils with very high organic matter contents (e.g. peat soils) have a low ability to fix cesium due to the low clay content, which means that the cesium remains available for plant uptake [Koblinger-Bokori 1996]. • In [Chibowski 2002], the migration rates of 137Cs were estimated in two types of soil; low peat-muck soil and black earth soil. The surface layer (0–5 cm) in the low peat-muck soil sample contained only 13% of soil minerals and the deepest layer (30–40 cm) was mostly organic matter (about 99%). Whereas, the surface layer (0–5 cm) in the black earth soil sample contained 22% organic matter and the deepest layer (25–30 cm) contained only 2% organic matter. The migration rates of 137Cs in low peat-muck soil sample were found to be significantly higher than in a black earth soil sample. 50 These results were also confirmed by microcalorimetric measurements that 137 showed low adsorption of Cs on the organic soils
Some of the studies give evidences that support the assumption that 137Cs mobility in soil is lower in the presence of higher organic matter contents such as the studies mentioned below:
• High fraction of organic matter in soil reduces the 137Cs mobility [Fawaris 1995] . • It was found that the cesium uptake by plants is decreased due to the clay components in the organic horizons in forests [De Brouwer 1994]. • It was also suggested that the adsorption properties of clay minerals in soil are modified due to the organic matter [Staunton 2002]. • Slow migration of cesium in forest soils, with organic matter content higher than 85%, was reported by Cheshire and Shand [Cheshire 1991], which could be an evidence that the organic horizons of forest soils have a high ability to bind cesium.
1.4.2 Particle Size Distribution
It has been found in many studies that the particle size distribution has an important effect on the 137Cs migration rate and more precisely, the clay contents have higher ability for 137Cs retention than silt and much higher than sand.
• In a study done on soil samples collected from Croatia soon after the Chernobyl accident (during July 1986) by Barisic et al. [Barisic 1999], the soil composition was shown to have a clear effect on the 137Cs migration rate where the contents of clay played an important role in cesium retention (higher retention for higher clay content). • In a study done by Ivanov [Ivanov 1997] on soil samples collected from the 30- km restriction zone of the Chernobyl Nuclear Power Plant (ChNPP) between 1987 and 1993, the migration of both 137Cs and 90Sr was observed to be slow.
51 Strontium moved faster than cesium in both the sandy and peaty soils, while the differences were least in the peaty boggy soils, in which less retention of 137Cs was expected. • The retention of 137Cs at the surface of different soils was found to depend strongly on the contents of the clay minerals in soil (higher retention for higher clay content) [Fahad 1989, Arapis 2004, Hölgye 2000, Poreba 2003, Sigurgeirsson 2005].
1.4.3 Cation Exchange Capacity (CEC) and K, Mg and Ca Concentrations
The cation exchange capacity is a measure of the soil ability to exchang cations between soil and soil solution and in turns it measures the ability of soil to hinder the cation migration in it . The cation exchange happens because of the negative charge of soil minerals surfaces. The CEC is expressed in units of charge per weight of soil. Two units are used to express the CEC: meq/kg (milliequivalents of element per kg of dry soil) or cmolc/kg (centimoles of charge per kilogram of dry soil) where 10 meq/kg ≡ 1 cmolc/kg. Clay crystalline can be classified into three major groups: Kaolin, mica and montmorillonite, which have CEC of about 10 meq/g, 19-25 meq/g, 119-150 meq/g, respectively, whereas sand and silt composed mainly by quarts and feldspars, which have CEC of only a few meq/g [Nam 2003]. The main exchange cations in soil are Ca+2, Mg+2, K+ and Na+. The negatively charged surfaces have different selectivity for the cations in soil. In other words there is competition between the cations for cation exchange at soil surfaces. This selectivity is higher for the cations of higher valences. For the cations with the same valences; it is higher for the larger cation radius. The selectivity for Cs+, Sr+2, Ca+2, Mg+2 and K+ follows the following series: Sr+2 > Ca+2 > Mg+2 > Cs+ > K+. A few studies results are presented below regarding the effect of the CEC and Ca, Mg and K contents on 137Cs mobility in soil:
• It was reported by Sigurgeirsson et al. [Sigurgeirsson 2005], that the retention of 137Cs in volcanic soils in Iceland was very high, where most of 137Cs was
52 retained in the upper 5 cm (82.7% on average). It was also reported that the retained amount of 137Cs below the 5 cm depth is not related to the CEC, and the other soil factors do not explain the variations in this amount either. • In a laboratory study it was demonstrated that the (Ca + Mg)/K ratio may play a key role in accelerating the cesium fixation. In general, the higher the (Ca+Mg)/K ratio, the lower are the migration parameters [Koblinger-Bokori 1996]. • The content of potassium in soils could be a possible explanation for low migration rate of cesium in soil [Staunton 2002], since the cesium uptake by the plants is higher in the soils with low potassium content. This uptake by plant roots to above-ground plant tissues causes a re-deposition of cesium on the soil surface. • The mobility of cesium and its absorption by roots were increased largely in soil solution with low concentrations of potassium [Sanchez 1999].
1.4.4 Soil pH
• In a study done about the migration of 137Cs and 60Co in the Australian Arid Zone, it was found that the pH is the main factor affecting the adsorption of 60Co but has little influence on the sorption of 137Cs [Payne 2001]. • In soils with pH numbers between 4 and 7, cesium is likely to be immobilized rapidly [Koblinger-Bokori 1996].
53 2 Radioactivity Concentrations in Jordanian Soil and Plants Samples
2.1 Introduction
The importance of studying 137Cs in soil is attributed to its relatively long physical half life (30.17 y), which means that it remains in the environment for a long time, and in turns represents a source for external and internal dose. Moreover, its chemical and metabolic- physiological reactions are similar to those of potassium, which makes its biological half life longer. This importance arises also from the fact of its slow migration downward in soil and its partial absorption by plant roots, which leads to uptake by the vegetation and into the human food chain [UNSCEAR 1988; Annex D]. After the Chernobyl accident in 1986 many investigations, mainly in Europe, have been done on the subject of 137Cs in soil. From the radiological point of view, 131I and 137Cs are the most important radionuclides to be considered, because they are responsible for most of the radiation exposure received by the general population. The releases of 131I and 137Cs as a consequence of nuclear bomb tests and nuclear accidents are estimated with 1,760 and 85 PBq, respectively [UNSCEAR 1988; Annex D]. The 137Cs concentration in surface soil decreases under the influence of various processes like decay, mechanical removing with rainwater, vertical migration and diffusion into deeper layers of the soil. The area of northwestern section of Jordan is expected to be contaminated from the Chernobyl accident according to the map (Figure 1-29) [Othman 1990], which illustrates the first air masses carrying radioactivity from the Chernobyl region that entered Syria early in the morning of 7 May 1986. It is obvious from the map that these trajectories could pass over the northwestern section of Jordan. Another indicator for a Chernobyl influence on Jordan is the metrological data, which was collected from stations distributed over the northwestern section of Jordan. These metrological stations have recorded some small discrete amounts of rainfall (Table 2-1) during the period of 8th to 15th of May 1986. This metrological data has been collected as part of this work from the Jordanian metrological department during the first field trip to Jordan. 54 A third indicator for a Chernobyl influence on Jordan is the 137Cs profile for a sediment core taken in 1994 from the deep, central part of lake Kinneret (32° 49´ N, 35° 36´ E, see Figure 2-2) [Kirchner 1997]. In this profile (Figure 2-1) two peaks can be distinguished, at 4- 5 cm depth and at about 17 cm depth, in which the deeper one was attributed to the nuclear weapon tests fallout and the shalower to Chernobyl fallout.
Station Date Rain fall (mm) Irbid 09/05/1986 0.1 11/05/1986 1.6 12/05/1986 0.4 14/05/1986 2.4 Sweileh 8/05/1986 4.0 11/05/1986 0.6 El Ramtha 11/05/1986 2.0 12/05/1986 3.0 14/05/1986 2.8 Ras Monief 11/05/1986 2.8 14/05/1986 1.0 Amman Airbort 11/05/1986 0.1 Al Mafraq 11/05/1986 0.2 12/05/1986 0.2 Jerash No data No data Table 2-1: Rainfalls in northwestern section of Jordan in May 1986 collected from the Jordanian metrological department.
4 3.5
) 3 2.5 2 1.5
137Cs (dpm/g 1 0.5 0 0 5 10 15 20 25 Depth (cm)
Figure 2-1: 137Cs profile in a sediment core from kinneret lake [Kirchner 1997]. 55 Whilst some research has been done and published about natural radioactivity in Jordan, only one paper was published about artificial radioactivity in Jordanian soils [Al Hamarneh 2003], which reveals high concentrations of 137Cs and 90Sr in some regions in the northwestern section of Jordan. The origin of this contamination, however, was not addressed in that paper. In addition, this study includes some weak and questionable points, which can be summarized below:
I. Thirty two soil samples were collected from 11 sites, in which:
a. One sample per site was collected from the upper 2 cm among 5 of the sites. b. Four samples were collected from the upper 2 cm from one of the sites. c. Three samples from one of the site, in which two of them were collected from the upper 2 cm and the other was divided into two layers: 0–2 cm and 2–4 cm. d. Two samples from one of the sites, in which the first was collected from the upper 2 cm and the other was a core divided into two layers: 0–5 cm and 5–10 cm. e. Two samples from one of the sites, in which the first was from the upper 2 cm and the other was a core divided as: 0– 2, 2–7, 7–12 and 12–17 cm. f. One core sample only from one of the sites was divided as: 0– 2, 2–7, 7–17 and 17–22 cm. g. Three samples from one of the sites, in which tow samples are from the upper 2 cm and the other was a core divided as: 0–2, 2–7, 7–17, 17–27 and 27–32 cm. h. The surface soil samples (0–2 and 0–4 cm) were collected using a stainless steel template of 25 cm × 20 cm area and the core samples were collected using a coring tool from an area of 100 cm2.
The small number of samples per site and the small area from which the sample were taken could lead to a biased result. In other words, the results for each site were more probable to be not representative for that site. This can be seen for example for the site where the 4 surface soil samples were taken. In this site, extremely different activities were measured in the different samples (91.5 ± 1.6, 180.3 ± 2.9, 352.3 ± 5.4 and 576.4 ± 8.8 Bq/kg).
56 Since the soil cores were sliced into relatively thick layers, it is impossible to get a representative shape for 137Cs profile or to identify the position/s of the profile peak/s. Therefore, these data are not suitable to study the migration of 137Cs.
II. The concentrations of 137Cs in the site, where the 4 surface soil samples were taken, were relatively very high as compared to those from the other sites in Jordan, which arises a question about that big difference.
III. The annual effective dose equivalent for 137Cs inventories has been estimated using a conversion factor of 1.4×10-8 Sv per Bq/m2. This conversion factor was used by [Othman 1990] for surface soil samples collected in May 1986. A reference for this conversion factor was not mentioned by [Othman 1990]. However, if we assume that this conversion factor was suitable for the surface soil samples of [Othman 1990], this should not be the case for cesium inventories of [Al Hamarneh 2003], which were collected in November 2000. On the other hand, Al Hamarneh et al. have done a survey of all the country and we based on there results of choosing the area of interest (northwest area of Jordan) for our work.
2.2 Motivation and Goals
The goal of this work is to study the artificial radioactivity in Jordan due to 137Cs and 90Sr in soils. For this purpose soil and plant samples from Jordan were collected. 137Cs was not recorded in Jordan, neither before nor after Chernobyl, with only one exception, which was the study of [Al Hamarneh 2003], where some weak points were pointed out and some questions were left unanswered. This work was an effort to achieve the following tasks and to answer the following questions:
I. How large is the contamination of 137Cs in the Jordanian soil? II. Is Jordan contaminated with 137Cs from Chernobyl accident? III. If Jordan was affected by Chernobyl, how large are the contaminations due to the Chernobyl fallout and the nuclear bomb tests fallout?
57 IV. Fitting the data of 137Cs in soil using a suitable Model to find out how 137Cs migrates in the Jordanian soils and if it is still available for the plants uptake. V. How large is the external effective dose equivalent due to the presence of 137Cs in soil? and does it represent any risk on the public health? VI. To compare the results of this study with those of [Al Hamarneh 2003] study, neighboring countries and countries with different climate types like Germany and some European countries.
2.3 Sampling, Samples Locations and Identification and Sampling Preparation
Two sets of soil samples were collected and brought from Jordan. The first set of samples was collected in April 2004 from eleven different sites of the northwestern part of Jordan. The second set of samples was collected in July 2005 from six of the previous sites where higher 137Cs contamination was found. Plant samples were also collected from the surfaces of eight of those sites; namely: AQ1, AQ3, AQ4, AQ5, AQ6, AQ7, AQ8 and AQ11 (Figure 2-2), where vegetations were found.
The northwestern part of Jordan was chosen to be our region of interest, for two reasons:
a. The population of Jordan concentrates mostly in this section of the country. Figure 2- 3 shows the population distribution of Jordan. b. Al Hamarneh et. al. [Al Hamarneh 2003] have measured relatively high level of 137Cs-contamination in some areas of northwestern section of Jordan compared with other regions in Jordan and its neighboring countries.
58 Figure 2-2: Jordan's map with samples locations.
Sampling locations are shown in Figure 2-2. The squares on the map represent the sites from where samples of the second set were brought. Site names, codes, and GPS coordinates are listed in Table 2-2. The samples in the second set have had the same codes as the first set of samples with an addition word “new”.
59 Figure 2-3: Population density of Jordan. (http://www.britannica.com/ebi/art-91996)
Sample Code Site GPS coordinates N E Alt.(m) AQ1 Kufrsum 32o 40´ 35o 49´ 506 AQ2 Foua’ara 32o 36´ 35o 45´ 373 AQ3 & AQ3new Baliela 32o 25´ 35o 55´ 740 AQ4 & AQ4new Qafqafa 32o 20´ 35o 58´ 927 AQ5 & AQ5new Dair Allyyat 32o 17´ 35o 52´ 882 AQ6 & AQ6new Abien 32o 21´ 35o 46´ 1036 AQ7 Ain El Basha 32o 04´ 35o 49´ 647 AQ8 Wadi El Naqah 32o 04´ 35o 45´ 985 AQ9 & AQ9new Bala’ama 32o 16´ 36o 05 708 AQ10 & AQ10new El Ramtha 32o 35´ 35o 58´ 542 AQ11 As Subeihi 32o 08´ 35o 42 537 Table 2-2: Soil samples identification.
60 2.3.1 Sampling Procedure
To get representative results a representative soil sample is necessary. A representative soil sample should give an average estimate of the whole sampled area. Areas alongside roads, low areas, salty or wet areas, areas with slopes and other variable areas should be avoided. Simple random, stratified random or systematic sampling pattern (Figure 2-4) can be used for uniform fields [Tan 1996].
Figure 2-4: Soil sampling plans [Jacobsen].
Using the simple random way of sampling means that the positions of the soil cores have to be selected randomly and independently, this helps in estimating a mean concentration for the sampled area. Stratified sampling can be used to reduce the variability of the sample [Mason 1992]. In stratified sampling, the area is divided into regions called Strata. These regions are expected to have uniform character (i.e. a smaller variance within the strata than that between strata). The sampling points within the strata can then be selected in a systematic or random way. In the systematic sampling, a point is being chosen as a first point, and the subsequent sampling points are determined by a specified system. In this type of sampling, the time of locating and traveling between the points may be reduced and a significant amount of costs may be saved [Lal 2001].
61 Systematic sampling is recommend by some experts only if the study focuses on estimating the population mean, whereas the random sampling is recommended if the study focuses on determining the precision of the estimate [Avery 1994]. In order to reduce the laboratory workload and the corresponding monitoring costs, a composite sampling could be used [Katz 1997]. The composite sample is a combination of the subsamples, therefore the data contained in a composite subsample is an average of all the subsamples making up the composite subsample. Thus, this method provides an excellent estimate of the mean concentration for the sampling area, without providing any information about the variation within that area [EPA 1992]. In this study the composite sampling method was used for the first set of samples. Each sample consists of a mixture of the individual cores. The accuracy and precision of the analytical result depends on the number of those cores, i.e. the probability for obtaining an inaccurate estimate of the average concentration of a radionuclide will decrease for a greater number of cores, but usually the time and the effort required for collecting the cores determines the number of cores taken.
A core tool (hand auger) (Figure 2-5) was used to collect the first set of samples. This auger is a hollow steel pipe of 18 mm inner diameter. A suitable place to take the samples was chosen. This means that the surface was undisturbed (i.e. not cultivated recently) and with an area of about 10m×10m and has a low slope. Each sample was taken in a simple random way, where the following steps were done in every site for a sample collection:
a. The core tool was inserted to the desired depth, turned and then brought out with the soil column. Figure 2-5: Hand auger b. The soil column was divided into five subsamples in all used in soil sampling. profiles except in AQ11, where it was divided into six subsamples.
62 c. The subsamples of soil cores, which had the same depth, from a minimum of 15 cores (except for AQ11 and AQ7, 10 cores for each) were mixed thoroughly in clean 0.5 L plastic containers (Figure 2-6). d. The plastic containers were properly labeled with site name, sample number, sampling depth, and sampling date. Other descriptive characteristics such as field characteristics, time of sampling and geographical coordinates of the field were recorded. A handhold GPS receiver was used to find the geographical coordinates of the sites. e. Two photographs were taken of each site (Figures 2-7 to 2-17). These photographs show either no slope, or low slope of the sampling areas.
Figure 2-6: Plastic containers used to collect the samples.
Figure 2-7: The sampling area in Kufrsum (AQ1).
63
Figure 2-8: The sampling area in Foua'ra (AQ2).
Figure 2-9: The sampling area in Baliela (AQ3).
Figure 2-10: The sampling area in Qafqafa (AQ4).
Figure 2-11: The sampling area in Dair Elleyyat (AQ5) 64
Figure 2-12: The sampling area in Abien (AQ6).
Figure 2-13: The sampling area in Aien El Basha (AQ7).
Figure 2-14: The sampling area in Wadi El Naqah (AQ8).
Figure 2-15: The sampling area in Irhab (AQ9).
65
Figure 2-16: The sampling area in El Ramtha (AQ10).
Figure 2-17: The sampling area in As Subeihi (AQ11).
The second set of soil samples was taken using two stainless steel plates with areas of 10 x 10 cm and 10 x 20 cm. The soil profiles in the second set were sliced into thinner layers, which is important for a detailed study of the profiles and to apply a migration model on our resulted data. The main aim of collecting the second set of samples was to study 137Cs vertical migration in soil, thus every sample was taken only from one position in the sampling area and no sampling compositing was necessary. Every profile was sliced into 1 cm thickness for the layers between 0 and 12 cm, 2 cm thickness for the layers between 12 and 20 cm and into 4 or 5 cm thickness for the layers between 20 and 30 cm.
66 2.3.2 Sample Preparation
2.3.2.1 Preparation of Samples for Gamma Measurements
Before submitting the samples for analysis of gamma emitting radionuclides, the samples had to be prepared in the laboratory to get the suitable geometry for analyzing.
2.3.2.1.1 Preparation of the First Set of Samples
The following steps were followed to prepare the first set of samples for gamma measurements:
1. Every subsample was dried in an oven at temperature of 105 °C until it reached a constant weight. 2. After removal of all stones and vegetation, the samples were milled and homogenized using a mixer of about 5 rps for about 15 min a sample. 3. An amount of 105 g of every sample was mixed with 11 g of wax and compressed to get a cylindrical disc of 7 cm diameter, 2 cm thickness and density of 1.5 g/cm3 (Figure 2-18). Achieving a soil disc with the mentioned dimensions was necessary since the efficiencies available were limited. Gamma-Plus software was available to analyze the first set of samples. The available efficiencies in efficiencies library of this software contains efficiencies for a cylindrical disc of soil of a diameter of 7 cm, four different disc heights (5 mm, 10 mm, 15 mm and 20 mm) and two densities (1.0 g/cm3 and 1.5 g/cm3). Therefore, the geometries of the first set of samples were built to be suitable for the available efficiencies. This was not required for the second set of samples where the new software Genie-2000 was used. Using this software, it is possible to build efficiencies of different diameters, heights and densities. Building different efficiencies using Genie-2000 will be explained in section 2.4.1.2. 4. Every sample was sealed with a sheet of metallized plastic foil, which was radon- tight, and labeled with necessary information about the sample (Figure 2-19). Sealing was necessary to reach the secular equilibrium between 226Ra and its daughter 222Rn 67 since 222Rn is a gaseous and can escape leading to non-equilibrium 226Ra progeny. After reaching that equilibrium the activity of 226Ra can be considered to be equal to the activity of 214Bi or 214Pb. Studying the activity of the natural radionuclides is not a goal of this study but it could be an objective goal for a future study.
The same procedure was used to prepare the plant samples but using 10 g of grass and 9.2 g wax to get a cylindrical disc of 7 cm diameter, 0.5 cm thickness and density of 1.0 g/cm3.
Figure 2-18: Soil-Wax pellet. Figure 2-19: Sealed sample.
2.3.2.1.2 Preparation of the second Set of Samples
The following steps were followed to prepare the 2nd set of samples for gamma measurements:
1. Samples were sieved using a 2 mm sieve. 2. Samples were dried in an oven at temperature of 105 oC until their weights became constant. 3. The dried samples were filled in 20 mm plastic Petri dishes of 76 cm3 volume (Figure 2-20).
68 Figure 2-20: 20 mm plastic petri-dish soil sample.
2.3.2.2 Radiochemical Separation to Determine 90Sr Concentrations (Beta Measurements)
2.3.2.2.1 Introduction
The separation of 90Sr was done using the so-called Nitric Acid Method. This method is well known in separating 90Sr in soil but it is time consuming. The chemical separation was done in our chemical laboratory using the procedure explained in the measuring guidance [BMU 2000] published by the German Federal Ministry for Environment, and described below.
2.3.2.2.2 Sample Preparation
Before starting the chemical extraction of 90Sr and 90Y from a soil sample, they have to be prepared for such a kind of analysis. This preparation includes drying and sieving and then ashing. Samples were dried in an air circulation drying cabinet at 50 C until they reached constant weights. Then they were broken up by hand and then sifted through a sieve with a mesh aperture of 2 mm. Samples were then ashed in an oven at about 500 C.
69 2.3.2.2.3 Radiochemical Separation
The radiochemical separation for strontium and yttrium from a soil sample can be summarized in the following steps:
1. Adding strontium-carrier to find the chemical yield for strontium:
To find the chemical yield for strontium in this chemical extraction, 10 ml of stable strontium-carrier solution was poured into a 250 ml-beaker, which contained 10 g of the ashed soil. The mixture was then boiled for 30 min with HCl and then filtered.
2. Calcium-Strontium Separation:
In order to perform the calcium-strontium separation, the filtrate from step (1) was +2 diluted with distilled water. Then oxalic acid (H2C2O4) was added to precipitate Sr
in the form of Sr-oxalate (SrC2O4). It is important to keep the pH value at 4.5 in this step to have low level of Ca+2 +2 -2 precipitation, which competes with Sr to composite with the oxalate ions (C2O4 ) and in turns affects strontium precipitation. o The precipitation was then ashed at 700 C to have strontium oxide (SrO2) and then
diluted with nitric acid HNO3. Smoking-nitric acid was then added slowly (to be sure that strontium is being precipitated only but not calcium) to the solution in order to precipitate strontium in
the form of strontium nitrate Sr(NO3)2.
3. Barium separation:
The precipitate from the last step was dissolved in distilled water and a barium carrier solution was then added to precipitate the barium in the form of barium chromate (BaCrO4).
4. Iron and Yttrium Separation:
70 The precipitate from the last step was dissolved in nitric acid (HNO3) and Iron-
carrier-solution. Then ammonia solution (NH3) was then added, to start precipitation
of iron and yttrium in the form iron hydroxide (Fe(OH)3) and yttrium hydroxide
(Y(OH)3). After finishing of precipitation, strontium again starts to build yttrium in the solution.
5. Extracting Strontium and the Build-up Yttrium:
Stable yttrium-carrier was then added to the supernatant from the last step to precipitate the yttrium and to estimate the yield of yttrium in the chemical extraction. Then supernatant was then covered and stored for 14 days to reach the activity equilibrium between 90Sr and its daughter 90Y.
Strontium carbonate SrCO3 was also precipitated with cold-saturated ammonium carbonate solution and then the yield of Sr+2 was determined. The precipitate was then dissolved with a small amount of HCl and cold-saturated oxalic acid was added to precipitate Yttrium in the form of Yttrium oxalate. The yield of Y+3 was determined and afterwards yttrium concentration was measured.
2.4 Measurements and Analysis
After preparation the samples was submitted to gamma and beta measurements and analysis.
2.4.1 Gamma Analysis
2.4.1.1 Measuring the Activities
After the samples preparation the samples were submitted for analysis of gamma emitting radionuclides using a HPGe detector (Figure 2-21) of 50 % relative efficiency and resolution of 2.0 keV at 1.33 MeV.
71 The system was set up to cover about 2 MeV–photon energy ranges over 4 k channels (4096 channels). Measurement time was at least 70000s (about 19 hours) and the dead time was less than 0.05 %. Energy calibration was performed for every spectrum. The calibration was done using the analyzing software, in which three or more gamma lines with known energies had been chosen from every spectrum and then a fit was executed with their corresponding channel numbers. The fits were always very good, where the 137Cs peak (661.7 keV) in the spectra was clearly specified with very small deviation (less than 0.2 keV). The background measurements were taken regularly for the gamma detectors. In the spectrum analysis, the background was always subtracted from the analyzed spectrum. The suitable efficiencies were chosen as well or established as described in the following section (section 2.3.1.2).
72 Figure 2-21: Gamma spectrometry used for gamma detection.
2.4.1.2 Determining and Building the Efficiencies
The analyses were carried out to determine the radioisotope 137Cs and the natural radioisotope 40K using the software of Silena–GAMMA+ version 1.02.1 for the first set and Genie-2000 V2.1. for the second set of samples. While using the GAMMA+ program, a limited number of efficiencies were available. The efficiencies for a cylindrical disc of soil were available for a diameter of 7 cm, four different disc heights (5 mm, 10 mm, 15 mm and 20 mm) and two densities (1.0 g/cm3 and 1.5 g/cm3). Therefore, the geometries of the first set of samples were chosen to be suitable for the available efficiencies. 73 Genie-2000 had been available in our laboratory since 2005, therefore it was used to analyze the second set of samples. This software supports efficiencies estimation for only one of the detectors in our laboratory. In this software, the LabSOCS (In Situ Object Counting System /Laboratory Sourceless Calibration Software) mathematical efficiency calibration software is being used for estimating the necessary efficiencies. This software is capable to produce efficiency calibration for a Germanium gamma sample in the laboratory without any need for radioactive sources. This is done by combining the detector characterization produced with the MCNP Monte Carlo modelling code and NIST (National Institute of Standards and Technology)-traceable sources. Those traceable sources were used by CANBERRA to characterize each Ge detector to be used with ISOCS/LabSOCS. A mathematical efficiency calibrations was done for specified detector and specified sample using the detector characterization, few physical sample parameters, and mathemat- ical geometry templates that have been created for a specified sample geometry. In a basic geometry template, the sample parameters can be specified such as the density, the size, the samples filling height, the material composition and the distance from the detector. This type of calibration is so useful of saving money and time. Extensive testing was done by CANBERRA for ISOCS/LabSOCS calibration results. Comparing these results to the results of the calibration performed using radioactive calibration sources; the agreement was within a few percent. Consequently, it was possible to build different efficiencies for the second set of soil samples with different densities without any need to mix them with wax to achieve specific densities, which was the case in the first set of samples.
2.4.1.3 Results and Discussion of Gamma Analysis
Below, the activities are expressed in Bq/kg (specific activity) and activities per unit area (Bq/m2). To convert from specific activity to activity per unit area (Bq/m2); the specific activity of that sample was multiplied by its dry bulk density (kg/m3) and its thickness (m).
74 For the first set of samples, the surface concentrations (the top layer) of 137Cs were in the range of 6.4 Bq/kg (414.8 Bq/m2) in AQ11 to 28.2 Bq/kg (1454.4 Bq/m2) for AQ4 with a mean value of 13.7 Bq/kg. The total inventory of 137Cs was in the range of 462 Bq/m2 in AQ9 to 2456 Bq/m2 for AQ4 with an average of 1886 Bq/m2. There were no detectable concentrations of 134Cs (concentrations were below the detection limit). This is due to the fact that the activity ratio 137Cs/134Cs at 1986 (directly after Chernobyl) was about 2:1 [UNSCEAR 2000; Annex J], and should be about 642:1 at the time of measurement (2004). In every sample the concentration of 40K in the subsamples was quite similar, which is an indicator for consistent results and homogeneous sampling (Table 2-3). The uncertainties included in Table 2-3 refer to the (SEM) standard Errors of the mean, which includes counting and calibration errors. Figure 2-22, Figure 2-23 and Figure 2-24 show depth distribution profiles of 137Cs for the first set of samples in the different sites. One can obviously see that the cesium profile in AQ8 is mostly flat especially for the top three layers (0–18 cm), which could imply that this profile is disturbed (e.g. by ploughing). In all of the profiles 137Cs concentration decreases with depth. The concentration drops faster in some profiles like AQ4, AQ5 and AQ6 than the others. The migration rate of 137Cs is slow since most of its concentration was still in top layers (0–15 or 20 cm).
75 137 134 40 Sample Cs Cs k Depth Code [Bq/kg (cm) [Bq/kg d.m.] [Bq/kg d.m.] d.m.] AQ1-1 0−5 7.66 ± 0.32 < 0.26 201.2 ± 7.1 AQ1-2 10−15 6.17 ± 0.28 < 0.29 174.2 ± 5.9 AQ1-3 15−20 6.62 ± 0.31 < 0.26 183.3 ± 6.7 AQ1-4 15−20 4.37 ± 0.23 < 0.23 197.3 ± 7.0 AQ1-5 20−25 2.10 ± 0.17 < 0.39 189.7 ± 6.8
AQ2-1 0−5 13.82 ± 0.46 < 0.27 228.2 ± 7.1 AQ2-2 15−20 11.92 ± 0.38 < 0.28 201.9 ± 6.2 AQ2-3 15−20 8.20 ± 0.33 < 0.28 200.4 ± 6.9 AQ2-4 15−20 3.62 ± 0.24 < 0.19 196.9 ± 7.2 AQ2-5 20−25 2.82 ± 0.21 < 0.23 189.9 ± 6.9
AQ3-1 0−5 8.43 ± 0.31 < 0.27 323.7 ± 9.6 AQ3-2 10−15 7.07 ± 0.30 < 0.22 305.8 ± 10.2 AQ3-3 15−20 4.46 ± 0.22 < 0.20 318.0 ± 9.5 AQ3-4 15−20 2.31 ± 0.15 < 0.19 274.9 ± 8.3 AQ3-5 20−25 0.73 ± 0.11 < 0.22 258.1 ± 8.7
AQ4-1 0−5 28.24 ± 0.80 < 0.17 237.7 ± 7.1 AQ4-2 10−15 13.39 ± 0.48 < 0.17 220.7 ± 7.5 AQ4-3 15−20 3.33 ± 0.18 < 0.18 228.9 ± 7.8 AQ4-4 15−20 1.23 ± 0.11 < 0.16 235.1 ± 7.2 AQ4-5 20−25 0.50 ± 0.09 < 0.16 198.3 ± 6.3
AQ5-1 0−5 15.96 ± 0.51 < 0.21 428.2 ± 12.4 AQ5-2 10−15 10.47 ± 0.40 < 0.21 376.5 ± 12.4 AQ5-3 15−20 4.21 ± 0.25 < 0.24 378.1 ± 12.5 AQ5-4 15−20 2.79 ± 0.18 < 0.24 408.4 ± 11.8 AQ5-5 20−25 1.27 ± 0.12 < 0.20 423.9 ± 12.3
AQ6-1 0−5 18.22 ± 0.52 < 0.20 482.2 ± 13.2 AQ6-2 10−15 10.05 ± 0.36 < 0.24 446.1 ± 13.9 AQ6-3 15−20 7.75 ± 0.31 < 0.24 559.6 ± 13.1 AQ6-4 15−20 3.10 ± 0.20 < 0.21 448.9 ± 14.5 AQ6-5 20−25 2.89 ± 0.18 < 0.21 495.7 ± 14.1 Table 2-3: Concentrations of 137Cs, 134Cs and 40K (d.m. ≡ dry mass). continue ;
76 137Cs 134Cs 40k Sample Depth [Bq/kg Code (cm) [Bq/kg d.m.] [Bq/kg d.m.] d.m.] AQ7-1 0−7 10.18 ± 0.40 < 0.24 123.1 ± 4.8 AQ7-2 14−21 8.76 ± 0.38 < 0.22 119.4 ± 5.0 AQ7-3 14−21 4.15 ± 0.22 < 0.19 136.1 ± 4.8 AQ7-4 21−31 0.76 ± 0.12 < 0.22 128.9 ± 5.0 AQ7-5 31−41 1.24 ± 0.09 < 0.15 141.1 ± 4.7 AQ7-6 41−51 0.63 ± 0.07 < 0.13 130.7 ± 4.0
AQ8-1 0−6 14.65 ± 0.52 < 0.23 354.3 ± 11.6 AQ8-2 6−12 14.98 ± 0.48 < 0.20 390.2 ± 11.4 AQ8-3 12−18 12.68 ± 0.47 < 0.41 371.0 ± 12.2 AQ8-4 18−24 7.22 ± 0.29 < 0.20 356.0 ± 10.6 AQ8-5 24−30 2.12 ± 0.16 < 0.21 257.0 ± 8.8
AQ9-1 0−6 2.69 ± 0.17 < 0.19 372.9 ± 10.5 AQ9-2 6−12 1.81 ± 0.12 < 0.19 338.2 ± 10.9 AQ9-3 12−18 1.07 ± 0.12 < 0.20 339.5 ± 10.1 AQ9-4 18−24 0.99 ± 0.12 < 0.23 303.2 ± 10.2 AQ9-5 24−30 0.26 ± 0.09 < 0.20 287.3 ± 9.67
AQ10-1 0−6 4.86 ± 0.23 < 0.19 395.0 ± 11.5 AQ10-2 6−12 3.89 ± 0.21 < 0.20 351.2 ± 10.4 AQ10-3 12−18 2.14 ± 0.19 < 0.24 342.0 ± 11.4 AQ10-4 18−24 1.31 ± 0.13 < 0.22 370.5 ± 12.0 AQ10-5 24−30 0.61 ± 0.10 < 0.18 374.7 ± 10.8
AQ11-1 0−6 6.40 ± 0.23 < 0.16 86.8 ± 3.3 AQ11-2 6−12 5.38 ± 0.24 < 0.19 73.9 ± 3.1 AQ11-3 12−18 4.45 ± 0.22 < 0.18 76.0 ± 3.3 AQ11-4 18−24 2.46 ± 0.18 < 0.21 67.2 ± 3.2 AQ11-5 24−30 0.95 ± 0.11 < 0.17 51.5 ± 2.9 Table 2-3 (continued): Concentrations of 137Cs, 134Cs and 40K (d.m. ≡ dry mass).
77 30 AQ1 AQ2 25 AQ3 AQ8
20
15 137Cs (Bq/kg) 10
5
0 0 5 10 15 20 25 30 Depth (cm)
Figure 2-22: 137Cs depth profile in AQ1, AQ2, AQ3 and AQ8.
30 AQ4 AQ5 25 AQ6
20
137Cs 15
10
5
0 0 5 10 15 20 25 30 Depth (cm)
Figure 2-23: 137Cs depth profile in AQ4, AQ5 and AQ6. 78 30 AQ7 AQ9 25 AQ10 AQ11 20
15 137Cs (Bq/kg) 10
5
0 0 5 10 15 20 25 30 Depth (cm)
Figure 2-24: 137Cs depth profile in AQ7, AQ9, AQ10 and AQ11.
Inventories of 137Cs from the southeast region of Syria (Syrian Badia) were measured in the year 2000 by Al-masri [Al-masri 2006(a)], where the concentrations were relatively uniformly distributed with values lower than 2000 Bq/m2, which were attributed to global nuclear bomb tests fallout. Comparing the results of this study (462 Bq/m2 - 2456 Bq/m2 with an average of 1886 Bq/m2) with Al-masri’s results implies that the Chernobyl impact was lower on Jordan as compared to Syria. In general, the values in this study were lower than the values from Lebanon (2805–6545 Bq/m2 in 1998-2000) [El Samad 2007], west Syria (500–8000 Bq/kg in the upper 5 cm layer in 1995) [Al-Rayyes 1999] and from Syria (320–9647 Bq/m2 in 2000–2003) [Al-masri 2006(a)], where it is believed that the Chernobyl effect was higher. Table 2-4 shows a comparison with Al Hamarneh’s study [Al Hamarneh 2003]. This comparison was done between locations in this study and their corresponding nearest locations in Al Hamarneh’s study, which can be noticed from their GPS coordinates. For the purpose of comparison, a decay correction was done for 137Cs activities in this study to the year 2000, which was the sampling date for Al Hamarneh’s samples. 79 In general, the values of 137Cs in Al Hamarneh’s study are higher as compared to the results of this study. The following points can be also noticed:
I. In Al Hamarneh’s study, the values of 137Cs differ considerably in the surface layers (0-2 cm) taken from the same location. This implies that non of these values can be representative for that location. On the contrary, the values in this study represent an average of 15 cores collected from a 10m × 10m area. II. The ratios of 137Cs activity in the surface layers (0–2 cm) in this study to those in Al Hamarneh’s study are: 1.5–3.1 (Hartha : Kufrsum), 2.9–18.7 for Qafqafa, about 7.2 for Dair Allyyat and 2.4 (Abien : Anjara). III. Some values of 137Cs for Qafqafa in Al Hamarneh’s study are extremely high as compared to those in our study and relatively high as compared to the other areas of his own study.
80 Reference Reference This work Al Hamarneh, 2003 Date Date 2000 2000
GPS Coordinates Sample 137 GPS Coordinates 137 Depth Cs Site Sample Depth Cs Site Code Alt (cm) [Bq/kg d.m.] Alt Code (cm) [Bq/kg d.m.] N E N E (m) (m) Kufrsum 32° 40´ 35° 49´ 506 AQ1-1 0−5 8.38 ± 0.35 Hartha 32° 42´ 35° 50´ 430 jor1 0−2 19.6 ± 0.61 AQ1-2 10−15 6.75 ± 0.30 jor2 0−2 26.15 ± 0.65 AQ1-3 15−20 7.24 ± 0.34 jor3.1 0−2 12.11 ± 0.49 AQ1-4 15−20 4.78 ± 0.26 jor3.2 2−7 7.44 ± 0.59 AQ1-5 20−25 2.30 ± 0.19 jor3.3 7−17 11.49 ± 0.44 jor3.4 17−27 3.19 ± 0.34
Qafqafa 32° 20´ 35° 58´ 927 AQ4-1 0−5 30.87 ± 0.88 Qafqafa 32° 22´ 35° 56´ 910 jor6 0−2 180.3 ± 2.90§ jor7 0−2 576.4 ± 8.75 jor8 0−2 352.3 ± 5.43 jor9 0−2 90.51 ± 1.61
Dair Allyyat 32° 17´ 35° 52´ 882 AQ5-1 0−5 17.45 ± 0.56 Dair Allyyat 32° 18´ 35° 52´ 930 jor10.1 0−2 126.5 ± 2.12 AQ5-2 10−15 11.44 ± 0.44 jor10.2 2−7 15.54 ± 0.48 AQ5-3 15−20 4.60 ± 0.27 jor10.3 7−12 1.13 ± 0.23 AQ5-4 15−20 3.05 ± 0.19 jor10.4 12−17 Below D. L. AQ5-5 20−25 1.39 ± 0.13 jor11 0−2 125.1 ± 2.07
Abien 32° 21´ 35° 46´ 1036 AQ6-1 0−5 19.92 ± 0.57 Anjara 32° 18´ 35° 46´ 1150 jor14 0−2 48.39 ± 0.98 Table 2-4:A comparison with Al Hamarneh’s study (Al Hamarneh, 2003).
81 Table 2-5 shows 137Cs concentrations in grass and plant samples taken from eight of the sampling sites. These samples were comprised mainly of grass and some wild leafy plants. The types of these plants have not been specified. The highest concentration of 137Cs was 5.3 Bq/kg dry mass in AQ1 while the concentration was below the detection limit in AQ11. This implies that 137Cs is still available for plants uptake and in turns to the cattle and then to humans.
Site AQ1 AQ3 AQ4 AQ5 AQ6 AQ7 AQ8 AQ11 137Cs 5.29 1.82 2.49 1.45 2.39 4.35 3.14 <0.85 (Bq/kg) ± ± ± ± ± ± ± 0.33 0.28 0.24 0.16 0.38 0.63 0.37 Table 2-5: 137Cs concentrations in plant samples.
The measured activities of 137Cs in the second set of samples have been listed in Table 2-6. In all profiles 137Cs concentration decreases with depth as expected. Most of 137Cs concentration is still in the top layers (0–15 cm) in all profiles except AQ9new where most of the concentration is in the first 20 cm. The activities of 137Cs were below the detection limits in AQ4new and AQ10new for the layers below 16 cm. Two peaks for 137Cs can be clearly seen in AQ4new at the depths of 3 cm and 8 cm, where the first peak could be attributed to Chernobyl fallout (Ch.) and the second to the nuclear bomb tests global fallout (N.B.). Such a profile with two obvious peaks is rarely found in the literature. Two peaks can be also seen in both AQ5new (at 5 cm and at 14 cm) and AQ6new (at 4 cm and at 14 cm), in which the first peak may be attributed to Chernobyl fallout and the second may be attributed to the nuclear bomb tests. Only one peak can be seen in AQ3new at 5 cm depth, in AQ10new at 3 cm depth and in AQ9 at 13 cm depth. The depth of the peak in AQ9new indicates that it is more probably attributed to the global nuclear bomb tests fallout. Figure 2-25 to Figure 2-30 show depth distribution profiles of 137Cs for the second set of samples together with profiles of the first set for comparison.
82 It is clear from the new profiles that the migration rate of 137Cs in the soil is low. It is also clear that the thinner slicing for the second set of samples has great advantage since more information can be extracted from the profiles. On the other hand the amount of 137Cs deposition is not representative for these sites since the samples were taken from small area (10 x 10 cm and 10 x 20 cm). The profiles of the first set of soil samples are insignificant to show the shape of the 137Cs profiles in soil and the positions of the peaks but the amount of 137Cs depositions are representative for these sites because the samples were taken in a simple random way from an area of about 10m×10m. This could give an interpretation for the differences in the profile shapes and the inventories of 137Cs (Figure 2-31) between the first and the second set of samples for the same sites.
Depth AQ3new AQ9new AQ10new Depth AQ4new AQ5new (cm) (cm) 137Cs 137Cs 137Cs 137Cs 137Cs [Bq/kg d.m.] [Bq/kg d.m.] [Bq/kg d.m.] [Bq/kg d.m.] [Bq/kg d.m.]
0−1 8.73 ± 0.29 2.17 ± 0.22 14.14 ± 0.47 0−1 22.91 ± 0.74 15.24 ± 0.49 1−2 9.76 ± 0.43 2.79 ± 0.13 20.68 ± 0.60 1−2 22.91 ± 0.71 17.51 ± 0.58 2−3 10.32 ± 0.36 2.13 ± 0.14 28.52 ± 0.76 2−3 25.87 ± 0.71 21.49 ± 0.69 3−4 10.50 ± 0.31 2.32 ± 0.13 24.96 ± 0.72 3−4 21.51 ± 0.65 21.80 ± 0.60 4−5 10.81 ± 0.35 1.63 ± 0.17 18.32 ± 0.55 4−5 20.47 ± 0.64 24.15 ± 0.70 5−6 10.45 ± 0.37 2.07 ± 0.25 13.17 ± 0.46 5−6 25.12 ± 0.74 23.64 ± 0.64 6−7 10.60 ± 0.36 2.61 ± 0.10 7.88 ± 0.33 6−7 28.12 ± 0.82 21.77 ± 0.62 7−8 7.97 ± 0.36 2.17 ± 0.21 6.19 ± 0.33 7−8 33.46 ± 0.95 19.32 ± 0.54 8−9 4.31 ± 0.15 2.03 ± 0.19 5.38 ± 0.32 8−9 22.89 ± 0.72 17.95 ± 0.59 9−10 3.50 ± 0.16 2.59 ± 0.15 4.86 ± 0.20 9−10 17.00 ± 0.57 14.63 ± 0.47 10−11 2.59 ± 0.15 3.27 ± 0.02 3.16 ± 0.26 10−11 8.08 ± 0.37 12.89 ± 0.48 11−12 1.71 ± 0.23 3.55 ± 0.25 3.22 ± 0.23 11−12 4.63 ± 0.29 12.24 ± 0.59 12−14 1.13 ± 0.12 4.51 ± 0.14 2.43 ± 0.27 12−14 1.81 ± 0.27 11.59 ± 0.52 14−16 0.98 ± 0.07 2.75 ± 0.12 0.48 ± 0.14 14−16 0.70 ± 0.10 9.49 ± 0.52 16−18 0.66 ± 0.10 1.96 ± 0.08 0.38 ± 0.17 16−20 <0.57 6.88 ± 0.40 18−20 0.63 ± 0.10 1.11 ± 0.10 < 0.46 20−24 <0.42 5.20 ± 0.29 20−25 0.92 ± 0.10 1.14 ± 0.08 < 0.58 25−30 <0.84 0.82 ± 0.12 < 0.51 Table 2-6: Concentrations of 137Cs in the second set of samples (d.m. ≡ dry mass). 83 12 AQ3 AQ3new 10
8
6
137Cs (Bq/kg) 4
2
0 0 2 4 6 8 1012141618202224 Depth (cm)
Figure 2-25: 137Cs depth profile in AQ3 and AQ3new.
40 AQ4 AQ4new 35
30
25
20
15 137Cs (Bq/kg)
10
5
0 0 2 4 6 8 1012141618202224 Depth(cm)
Figure 2-26: 137Cs depth profile in AQ4 and AQ4new.
84 30 AQ5 AQ5new 25
20
15
137Cs (Bq/kg) 10
5
0 0 5 10 15 20 25 Depth (cm)
Figure 2-27: 137Cs depth profile in AQ5 and AQ5new.
25 AQ6 AQ6new 20
15
10 137Cs (Bq/kg)
5
0 0 2 4 6 8 10 12 14 16 18 20 22 24 26 28 30 Depth (cm)
Figure 2-28: 137Cs depth profile in AQ6 and AQ6new.
85 5 AQ9 4.5 AQ9new 4 3.5 3 2.5 2 137Cs (Bq/kg) 1.5 1 0.5 0 0 5 10 15 20 25 30 Depth(cm)
Figure 2-29: 137Cs depth profile in AQ9 and AQ9new.
35 AQ10 AQ10new 30
25
20
15 137Cs (Bq/kg) 10
5
0 0 5 10 15 20 25 30 Depth(cm)
Figure 2-30: 137Cs depth profile in AQ10 and AQ10new.
86 First Set of Samples 5000 Second Set of Samples 4000 )
3000
2000 137Cs (Bq/m2 1000
0 AQ1 AQ2 AQ3 AQ4 AQ5 AQ6 AQ7 AQ8 AQ9 AQ10 AQ11
Figure 2-31: 137Cs inventories for the first and the second sets of samples.
2.4.1.4 Soil-to-Plant Transfer Factors
Soil to plant transfer, migration of the radionuclides to ground water, dust and direct radiation would be possible ways for the radionuclides deposited in soils to reach humans. Soil-to-plant transfer factors are commonly used to estimate the food chain transfer of the radionuclides. Their definition assumes that the concentration of a radionuclide in a plant relates linearly solely to its average concentration in the rooting zone of the soil. A transfer factor is defined as the concentration of pollutants in the plant, divided by the concentration of pollutants in the soil [Ehlken 1996]. The soil-to-plant transfer factor (Tf) for 137Cs is given, according the previous definition, in the following equation:
activity concentration of 137Cs in Bq per kg dry plant mass Tf = 137Cs activity concentration of 137Cs in Bq per kg dry soil mass within the rootingzone
Where the rooting zone was considered to be the upper two layers (i.e. the upper 10 cm in AQ1–AQ5, 12 cm in AQ8 and 14 cm in AQ7).
87 137 Values of the soil-to-plant transfer factor for Cs (Tf137Cs) were calculated, according to Eq. 2-18, using the data of the first set of soil and plants samples. They varied within the range of 0.11 in AQ5 to 0.77 in AQ1 (Table 2-7).
Site AQ1 AQ3 AQ4 AQ5 AQ6 AQ7 AQ8 AQ11 0.77 0.24 0.12 0.11 0.17 0.46 0.21 Tf137Cs ± ± ± ± ± ± ± -- 0.05 0.04 0.012 0.01 0.03 0.07 0.03 Table 2-7: Soil-to-plant transfer factors for 137Cs.
The observations of many of publications indicates that, for a number of long-lived radionuclides, soil-to-plant transfer factors show variations which may exceed three orders of magnitude [Coughtrey 1982, Frissel 1992]. This extreme variability indicates that a general relationship between the soil and plant concentrations of a radionuclide does not exist, in contrast to the basic assumption of the above equation [Ehlken 1996].
2.4.2 Beta Analysis
2.4.2.1 Measuring the Activities
After samples preparation, the samples were submitted for analysis of beta emitting radionuclides using a Berthold gas-filled proportional detector of type Low-Level- Handprobenwechsler LB 750 L for 25 mm or 50 mm dishes (Figure 2-32), flushed with P-10 gas (90 % Argon + 10 % Methan), and efficiency (eff) of 21.97%. The measurements were carried out for the time of about 9 h for the sample measured over daytime and about 15 hours for the samples measured overnight. Beta measurements were done for the first set of the soil samples, where a representative mixture from the subsamples of every profile was prepared to find the total inventory of 90Sr for that profile. Detailed profiles were also measured for AQ4, AQ5, AQ6 and AQ4new.
88 The recovery factor of 90Sr separation had a range of 70.1% to 96.4% and the detection limit ranged from 0.26 to 0.39 Bq/kg.
Figure 2-32: Gas-filled proportional detector of kind Low-Level-Handprobenwechsler LB 750 L, Berthold.
2.4.2.2 Results of Beta Analysis
The inventories of 90Sr are shown in Figure 2-33. The highest activity of 90Sr was 3.57 ± 0.22 Bq/kg in AQ6 and the lowest was 0.52 ± 0.26 Bq/kg in AQ11 with an average of 1.91 ± 0.80 Bq/kg. The concentration of 90Sr in AQ9 was below the detection limit (0.27 Bq/kg). The uncertainties for AQ1, AQ10 and A11 were relatively high (21 %, 21 % and 50 %, respectively) due to the low signal. The inventories of 90Sr will be useful to calculate the 137Cs-90Sr ratio, which could be a suitable method to determine the deposition of 137Cs from Chernobyl and from the nuclear bomb tests fallouts, as it will be described in Chapter 3. The depth profiles for AQ4, AQ5 and AQ6 are shown in Figure 2-34 and the depth profiles of 90Sr and 137Cs in AQ4new are showed together in Figure 2-35. It is obvious from these figures that the migration of 90Sr is faster than migration of 137Cs (i.e. the mobility of 90Sr in soil was higher), which is consistent with the results of many earlier
89 studies [e.g. Riise 1990, Kirchner 1992, Deborah 1992, Mahara 1995, Korobova 1998, Golovatyj 2002].
4
3.5
3
2.5
2
90Sr (Bq/kg) 1.5
1
0.5
0 AQ1 AQ2 AQ3 AQ4 AQ5 AQ6 AQ7 AQ8 AQ9 AQ10 AQ11
Site
Figure 2-33: 90Sr inventories for the first set of samples.
6 AQ4 AQ5 5 AQ6
4
3 90Sr (Bk/kg) 2
1
0 0 5 10 15 20 25 30 Depth(cm)
Figure 2-34: 90Sr depth profiles for AQ4, AQ5 and AQ6. 90 AQ4new 90Sr 40 137Cs 35 30 25 20 Bq/kg 15 10 5 0 0 5 10 15 20 25 Depth (cm)
Figure 2-35: 90Sr and 137Csdepth profiles for AQ4new.
137 2.5 Comparison of Cs Concentrations in Soil in Jordan with Some European and Middle East Countries
Germany is one of the most affected countries of 137Cs from Chernobyl especially in the states of Bavaria and Baden Württemberg (Table 2-8) [BMU 2004]. The average concentration of 137Cs in the topsoil layers (0-10 cm) in Germany for the year 2004 is about 22.5 Bq/kg. The average concentration of 137Cs in the topsoil layers (0-10 cm and 0-12 cm) was calculated for the first set of samples from Jordan and found to be about 10 Bq/kg, which is less than half the average of Germany. Most of the German states have an average concentration of 137Cs higher than its average value in northwestern part of Jordan, especially in Bavaria where the average is about ten times the value in Jordan.
91 137Cs Land of the Federal Republic [Bq/kg dry mass] Baden Wurttemberg 42.2 Bavaria 101.4 Berlin 11.3 Brandenburg 14.2 Bremen 7.6 Hamburg 5.8 Hessen 20.0 Lower Saxony 17.2 Mecklenburg-West Pomerania 30.7 North Rhine Westphalia 16.6 Rhineland Palatinate 16.2 Saarland 14.3 Saxony 16.1 Saxony Anhalt 9.3 Schleswig-Holstein 14.4 Thuringia 22.9 Table 2-8: The Average Concentration of 137Cs in pasture Soil (0-10 cm) in Germany [BMU 2004].
It can be also seen from Figure 1-27 that the deposition of 137Cs in 1986 in most parts of Germany is 0–4000 Bq/m2, which represent the regions least affected by Chernobyl. Making the decay correction of these values to the year 2004 we get a range of 0–2590 Bq/m2, which is comparable or somewhat higher than our results (460– 2455 Bq/m2). This can be an indication that the Chernobyl impact on Jordan was weak. A decay correction to the year 2004 (the sampling date of the first set) was performed for the inventories of 137Cs in some studies in European countries and neighboring countries of Jordan in the Middle East (see Table 2-9). As a comparison, the values from this work (462 - 2456 Bq/m2 with an average of 1886 Bq/m2) are higher than those from Nile delta in Egypt, which were attributed to the nuclear bomb tests fallout only. The values from this work were in general; lower than those for Syria and Lebanon, which indicates that they were more affected by Chernobyl. The values from this work were lower than those in the European studies, as expected, since these countries were highly contaminated by Chernobyl as mentioned in Chapter 1. 92 137Cs (Bq/m2) Location Reference corrected to 2004 Jordan 460 – 2455 This work, first set of samples 135 – 2370* Al Hamarneh 2003 * these values represent surface samples (0-2 cm) Egypt (Nile delta) 10 – 1505 Shawky 1997 Syria 300 – 9000 Al-Masri 2006(a) Lebanon 2500 – 5835 El Samad 2007 Hungary 1430 – 10400 Szerbin 1999 Greece N.B. fallout 1340 Liritzis 1987 Ch. fallout 91810 Simopoulos 1989 U.K. (Devoke) 8945 Smith 1997 Bulgaria 26145 Pourchet 1997 Czech Republic (Bohemia) 2250 – 18300 Hölgye 2000 Poland 5185 Poreba 2003 Iceland 280 – 4480 Sigurgeirsson 2005
Table 2-9: The deposition of 137Cs in Jordan and some Middle East and European countries.
2.6 External Dose
The gamma emitter 137Cs in soil represents nowadays the main artificial source for the external dose because of its relatively long half life (30.17 y) and due to its slow migration in soil. In order to estimate the annual effective dose equivalent (E), the following equation can be used [UNSCEAR 2000; Annex A]:
E = D ⋅t ⋅ f Eq. 2-1 where D is the absorbed dose rate (Gy/y), t is the average annual outdoor-time around the world (0.2 y) and f is the absorbed to effective dose conversion factor for adults (0.7 Sv/Gy). The last conversion factor has been used by UNSCEAR for the effective dose estimations and thus for our calculations. Higher values are used for infants (0.9 Sv/Gy) and for children (0.8 Sv/Gy) [Petoussi 1991, Saito 1998].
93 The external exposure dose rates (D) at 1 m above ground surface can be estimated by multiplying the activity concentration (A) by appropriate conversion factor (Cf);
= ⋅ D A C f Eq. 2-2
In several studies [e.g. Kocher 1985, Chen 1991, Saito 1985, Saito 1995], Monte Carlo codes were developed to calculate the activity concentration in soil to dose conversion factors for different gamma energies. These factors were calculated for different radionuclides distributions in soil such as plane sources at different soil depths, uniform slab sources between the ground and different soil depths, exponentially and Gaussian- shape distributed sources in soil. Since the data of the first set of soil samples were more representative for the sampling areas, they will be used for the effective dose estimations. The soil profiles in the first set of samples were sliced into relatively thick layers (≥ 5cm), thus fitting the profiles by an exponential or a Gaussian function would give approximate descriptions for them. Therefore, the activity concentrations for the different layers in every profile were treated as uniform slab sources. The conversion factors for uniform slab sources between the ground and different soil depths for 600 keV gamma energy and 1.4 g/cm3 soil density [Kocher 1985] were used (Table 2-10). The energies available in [Kocher 1985] around 661.6 keV were 600 keV and 800 keV, thus the gamma energy of 600 keV was used.
0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 Depth to to to to to to to to to to to to to to (cm) 0.5 1.0 2.0 3.0 4.0 5.0 7.5 10.0 15.0 20.0 25.0 30.0 40.0 50.0 Cf 78.9 155 275 371 450 517 647 740 858 923 960 981 1000 1001
Table 2-10: Dose rate conversion factors (μGy/y per Bq/cm3) at 1m above the ground for uniform slab sources between the ground and different soil depths for 600 keV.
From Table 2-10, the conversion factor for the layer 10-15 cm for example can be calculated as the difference between the conversion factors of 0-15 cm (858 μGy/y per Bq/cm3) and 0-10cm (740 μGy/y per Bq/cm3) which is 118 μGy/y per Bq/cm3. The same
94 way was used to calculate the conversion factors for the different layers in the first set of samples. The external exposure dose rates (D) at 1 m above the soil surface were calculated using Eq. 2-2 for every layer then the results of every profile were summed up to find its total D. Eq.2-1 was then used to calculate the annual effective dose (E) for every profile (see Table 2-11).
Site AQ1 AQ2 AQ3 AQ4 AQ5 AQ6 AQ7 AQ8 AQ9 AQ10 AQ11
E 1.08 1.81 1.10 2.63 1.86 1.99 1.38 2.07 0.30 0.61 0.79 ± ± ± ± ± ± ± ± ± ± ± ()Sv/y) 0.05 0.06 0.04 0.08 0.07 0.06 0.06 0.07 0.02 0.03 0.03 Table 2-11: Annual effective dose equivalent at 1m above the ground.
The ratio of E from the upper two layers to the total E had a range from 0.82 in AQ1 to 0.97 in AQ4 with an average of 0.90, which implies that the main contribution to the external dose comes from the upper soil layers. A conversion factor of 1.4×10-8 Sv/y per Bq/m2 was used by [Othman 1990] for soil samples collected from Syria soon after Chernobyl. This factor was used later on by [Al Hamarneh 2003] for samples collected from Jordan in 2000 and by [El Samad, 2006] for samples collected from Lebanon in 1998-2000. The conversion factor 1.4×10-8 Sv/y per Bq/m2 was used to estimate the value of E from every upper layer (0-5, 0-6, 0-7 cm) in each profile. The values of E estimated by this conversion factor for the upper layers were averagely 7 times the value of E estimated by the conversion factors from [Kocher 1985] and range from 5.8 times in AQ1 to 8.1 times in AQ9. Thus the use of the conversion factor 1.4×10-8 Sv/y per Bq/m2 for our samples leads clearly to overestimation of E, which implies that it is not suitable to be used. The interpretation for these differences is due to the fact that conversion factor used by [Othman 1990] was to convert activity concentration of 137Cs deposited on the soil surface after Chernobyl in 1986. Assuming that this factor was suitable for Othman’s study [Othman 1990], it should not be suitable for other studies because they were done long time after deposition. Otherwise, this includes ignoring for 137Cs migration in soil and, in effect, ignoring for self-absorption in soil.
95 The effective dose equivalent (E) in Jordan due 137Cs in soil ranges between 0.30 µSv/y and 2.63 µSv/y with an average of 1.42 µSv/y (0.00142 mSv/y) (Table 2-11), whereas the average value for Germany in the period of 2000–2004 due to Chernobyl 137Cs in soil was less than 0.01 mSv/y [BMU 2004]. Thereafter the effective dose equivalent of 137Cs in Germany is about 7 times the value in Jordan. The values of E due to 137Cs in soil were much lower in this work as compared to those estimated in Jordan by Al Hamarneh et al. [Al Hamarneh 2003] (3.8-214.2 µSv/y with an average of 60.4 µSv/y). The values of E due to 137Cs in soil in the neighboring countries were also higher as compared to Jordan (this work). Table 2-12 shows the annual effective dose equivalent at 1m above the ground for Jordan and its neighboring countries including the date of sampling.
Reference Land E ()Sv/y) Reference Date 0.30 – 2.63 2004 This work, first set of samples (1.42) Jordan Al Hamarneh 2003 3.8–214.2 * 2000 * these values represent (60.4) surface samples (0-2 cm) Egypt (Nile delta) 9.1 1988 Shawky 1999 2.8 1986-1987 Othman 1990 Syria 2 to 7 2000-2003 Al-Masri 2006(a) (4) Lebanon 19 – 91 1998-2000 El Samad 2007 Table 2-12: Annual effective dose equivalent at 1m above the ground for Jordan and its neighboring Countries.
The value of E due to 137Cs in Jordan is very small as compared to the E from natural radioactivity there. As an example, E from natural radioactivity in air in dwellings in Jordan was higher than 0.44 mSv/y [Ahmad 1998], which is about 310 times the average value of E due to 137Cs in soil estimated in this work. An average dose of 2.4 mSv/y (with about 35% obtained through external irradiation) was established by UNSCEAR, according to many studies, to be received by people all 96 over the world [UNSCEAR 1993, Zerquera 2001, Tahir 2005]. Thus, the external dose due to the presence of 137Cs in the Jordanian soils does not represent a significant health hazard.
97 3 Depth Distribution and Migration of 137Cs in Jordanian Soils
3.1 Introduction
The contamination of 137Cs in the atmosphere is a consequence of nuclear bomb testing and nuclear accidents and has thereafter deposited on the surface soil. Deposition of 137Cs was mainly by precipitation and some dry deposition took place. Around 90% of the total deposition of 90Sr and 137Cs due to the global fallout occurred as wet deposition. The deposition due to the Chernobyl accident showed deposition rates that are an order of magnitude higher when high rainfall during the cloud passage happened as compared to those observed for dry conditions [UNSCEAR 2000; Annex A]. The concentration of 137Cs in surface soil decreases under the influence of various processes like radioactive decay, mechanical removing with rainwater, vertical migration and diffusion into deeper layers of soil. The aim of this chapter is to determine the origin of 137Cs contamination in the Jordanian soils and to estimate the amounts of contamination from Chernobyl and from the nuclear bomb tests as well as to study the migration of 137Cs in soil.
3.2 Soil Analysis and the Effect of its Characteristics on 137Cs Migration in Soil
Soil analysis was carried out for the second set of soil samples using standard physical and chemical methods in the institute of Bodenkunde at Bremen University (see Table 3- 1). This was done for the upper 20 cm soil dividing them into two layers (0–10 cm and 10–20 cm). The soil particles of sizes less than 2 µm were classified as clay particles, 2–63 µm as silt particles and those of sizes 63–2000 µm as sand particles (German soil texture classification).
98 All samples contained higher amounts of clay minerals (43.3–69.8 %) and lower amounts of sand (1.3–18.1 %) and the rest is silt. It was clear that both of the clay and silt contents were lower in the first layer (0–10 cm) as compared to the second layer (10–20 cm) but in general the clay contents were high in both layers and the difference between the two layers was not high enough that different retention of cesium in the first and the second layers could be expected. The amount of organic matter was low in all soil samples (1.5–5.6 %), which suggests low effect of the organic matter on the mobilization (or fixation) of cesium in these soils. The CECs were higher in the upper soil layers (0–10 cm) as compared to the lower layer (10–20 cm), which could be a reason for higher retention for cesium in the upper layer. On the other hand, the (Ca+Mg)/K ratio was calculated and it is listed in Table 3-1. This ratio was always lower in the upper layer (0–10 cm), which could point to lower retention of cesium in the upper layer [Koblinger-Bokori 1996]. The pH number in all soils was about 7 or a little bit higher (7.1–7.7) and is thus outside the range 4–7, where the Cesium is expected to be immobilized rapidly.
99 Depth pH Exchangeable cations CECe Site (cm) Particle distribution (%) OM (%) (CaCl2) (cmolc/kg) (cmolc/kg) Ca + Mg Clay Silt Sand Ca Mg K Na K 0-10 55.9 38.9 5.2 2.1 7.1 48.89 3.40 0.64 0.19 81.7 52.3 AQ3new 10-20 60.4 34.7 4.9 1.7 7.2 48.23 2.60 0.27 0.76 188.26 47.5 0-10 46.3 46.2 7.5 5.6 7.4 40.34 2.47 0.61 0.90 70.18 67.7 AQ4new 10-20 52.4 39.4 8.2 4.2 7.6 40.63 1.49 0.28 0.50 150.43 59.1 0-10 60.4 38.1 1.5 3.7 7.4 56.72 1.55 0.84 0.20 69.34 51.1 AQ5new 10-20 69.8 28.9 1.3 3.0 7.5 55.96 2.17 0.32 0.08 181.65 40.5 0-10 55.9 41.0 3.1 4.3 7.1 47.90 1.80 1.76 0.69 28.24 56.6 AQ6new 10-20 64.8 33.4 1.8 2.5 7.2 44.02 1.80 0.84 0.43 54.55 45.3 0-10 43.3 39.5 17.2 2.2 7.2 40.71 2.11 0.77 0.33 55.61 48.3 AQ9new 10-20 44.5 37.4 18.1 1.6 7.3 41.23 2.32 0.32 0.21 136.09 44.5 0-10 49.9 47.2 2.9 2.4 7.6 40.50 2.70 0.97 0.30 44.54 58.7 AQ10new 10-20 55.6 42.1 2.3 1.8 7.7 42.50 2.69 0.21 0.52 215.19 54.1
Table 3-1: Physical and chemical proprieties of the analyzed soil sample.
100 3.3 Determining the Origin of 137Cs in the Jordanian Soils
The contamination of 137Cs in the Jordanian soils comes most probably from two sources: the nuclear bomb tests global fallout and Chernobyl fallout. To determine the source, the two methods described below were utilized:
I. 137Cs-90Sr Ratio
II. Convection Dispersion Migration Models of 137Cs in Soil
3.3.1 137Cs-90Sr Ratio
Data of the annual depositions of 137Cs and 90Sr was taken from the UNSCEAR 2000 report [UNSCEAR 2000; Annex C] for the northern hemisphere for the period 1945 to 1999. Decay correction was carried out to year 2004, which was the year of sampling for the first set of soil samples. The 137Cs–90Sr ratio was found to be about 1.83. Since the boiling temperature of 90Sr is higher than the one for 137Cs, the transfer of 90Sr from Chernobyl nuclear power plant into the atmosphere was only in one day where the temperature was high enough for its evaporation. Thus it was transported into the atmosphere in smaller amounts and reached some closer areas of Europe. Hence it is assumed that 90Sr in Jordanian soils originates only from the nuclear bomb tests fallout. Considering this assumption together with the inventories of 90Sr and 137Cs for the first 137 137 137 set of soil samples, the deposition of Cs from Chernobyl ( CsCh) as well as the CsCh 137 – CsNB ratio was calculated (Table 3-2 & Figure 3-2) using the equations;
Cs NB = 1.83 Eq. 3-1 SrNB
= + Cstot CsCh CsNB Eq. 3-2
= Srtot SrNB Eq. 3-3
Dividing Eq. 3-2 by CsNB and then substituting Eq. 3-1 and Eq. 3-2, we get the CsCh-
CsNB ratio:
101 Cs Cs Ch = tot −1 Eq. 3-4 ⋅ Cs NB 1.83 Srtot where the values of Cstot and Srtot (=SrNB) are known (measured). The calculated errors were high, which is expected as a result of the high experimental errors of 90Sr especially for AQ10 and AQ11.
Site AQ1 AQ2 AQ3 AQ4 AQ5 AQ6 AQ7 AQ8 AQ10 AQ11
CsCh : 3.77 2.97 2.94 3.90 3.42 2.41 3.35 4.45 2.14 7.57 ± ± ± ± ± ± ± ± ± ± Sr NB 0.89 0.44 0.59 0.52 0.55 0.26 0.71 0.63 0.56 3.88 CsCh : 1.06 0.62 0.61 1.13 0.87 0.32 1.83 1.43 0.17 3.14 ± ± ± ± ± ± ± ± ± ± CsNB 0.49 0.24 0.32 0.29 0.30 0.14 0.39 0.34 0.31 2.12
Table 3-2: The measured ratio of the total 137Cs to the total 90Sr and the calculated ratio of 137Cs from Chernobyl to nuclear bomb test 137Cs.
137Cs (Chernobyl) 137Cs (NB) 3000
2500
2000
1500 Bq/m2
1000
500
0 AQ1 AQ2 AQ3 AQ4 AQ5 AQ6 AQ7 AQ8 AQ10 AQ11
Figure 3-1: The calculated inventories of 137Cs from Chernobyl and nuclear bomb tests.
102 The deposition of 137Cs in Europe prior to Chernobyl accident was principally due to nuclear bomb tests fallout. The spatial distribution of 137Cs due to the nuclear bomb tests fallout was different for different latitudes. The average values of 137Cs in Europe just prior to Chernobyl accident were about 1.8, 2.4 and 2.2 kBq/m2 for latitudes 30-40°N, 40-50°N and 50-60°N, respectively [De Cort 1998]. After performing a decay correction to 1986, the mean deposition of derived nuclear bomb tests fallout 137Cs in this work was about 1630 Bq/m2 and it spanned a range of 475−2630 Bq/m2. This average value was similar to the average value of 137Cs in Europe just prior to Chernobyl accident for the latitudes 30-40°N. Table 3-3 shows Chernobyl depositions of 137Cs in Jordan and some European countries (reference date 1986). In general, Jordan is much less affected by Chernobyl than these European countries.
137 2 CsCh (Bq/m ) (1986) Location Reference Min Max Mean Jordan 200 3220 1395 This work, first set of samples Bulgaria 30400 Pourchet 1997 England 300 14200 3200 McAuley 1989 Estonia 120 samples 40000 20000 Realo 1995 over the land Finland 140 32000 Saxen 1987 Germany Upper Swabia 43175 Bilo 1993 North Rhine- 2385 Bilo 1993 Westphalia Greece (1242 samples 0 137000 Simopoulos 1989 collected over (± 1- the land) 10%) Italy Norh east 2000 60000 Battiston 1987 Campania 8100 Roca 1989 region (south) Netherlands 500 6000 Sloof 1992 Norway 5000 200000 Blakar 1992 Poland (Krakow area) 360000 Broda 1987
Table 3-3: Chernobyl deposition of 137Cs in soils from Jordan and from some European countries.
103 3.3.2 Convection Dispersion Migration Model of 137Cs in Soil
3.3.2.1 Introduction
It is important to understand and predict vertical migration of fallout radionuclides because of its radiological impact. Slow migration implies that the radionuclide stays in the upper layers of soil for long time, which makes it available to plant uptake contributing to the internal dose. Moreover, it contributes to the external dose by the direct irradiation. On other hand, fast migration of radionuclides in soil means that the radionuclide can enter the groundwater table quickly. Consequently several models were developed to describe 137Cs migration in soils and to explain its vertical distribution. A logarithmic-polynomial equation was applied by Barisic et al. [Barisic 1999] on soil samples collected in Croatia soon after Chernobyl accident (during July 1986) to study the vertical distribution of 137Cs, where a very good fit was found for the measured data. Exponential fit can be applied to the profiles in which there is no significant convection (i.e. no significant peak below the soil surface) [Smith 1999], this can be appropriate for samples collected directly after Chernobyl accident and where the cesium peak from Chernobyl is dominant. An exponential distribution fit differs from the experimental data, since it systematically underestimates the concentrations of radionuclides at deeper depths and the migration parameters (migration velocity v and dispersion coefficient D) get smaller for longer observation period [Konshin 1992(a)]. Some models were used to study the convection and the diffusion of cesium in soil. Velasco et. al used a model called RABES to study the migration of 137Cs and its convention in soil [Velasco 1997]. Takriti and Othman used the Fick's diffusion equation to find the diffusion coefficients of 90Sr and 137Cs in Syrian rocks [Takriti 1997]. Kirchner studied the applicability of the compartmental models to describe the transport of the radionuclides in soil [Kirchner 1998(a)]. In this study, it was shown that the compartmental model can only account for convection-dominated flow. In order to count for the diffusion process, this model has to be replaced by a model considering the backflow. Another important result in this study was that the number of compartments,
104 into which a soil is divided, should be determined by physical transport processes of water and solutes in the soil. Therefore, the compartmental model studies are scientifically questionable. The convection diffusion models are the most common in describing the migration of the radionuclides in soil under natural conditions and they can give a long-term prediction for the migration of the radionuclides in soil. Thus they were widely used for this purpose [e.g. Kirchner 1992, Konshin 1992(a), Konshin 1992(b), Ivanov 1997, Szerbin 1999, Bossew 2001, Bunzl 2001, Likar 2001, Krstic 2004, Timms 2004, Bossew 2004, Almgren 2006]. Thus, this model was chosen to be applied to our profiles in the second set of soil samples.
3.3.2.2 Theory
The convection diffusion model assumes that the vertical migration of radionuclides in soil is governed by physical-chemical processes. These are, convection and diffusion as transport mechanisms, and sorption as interaction mechanism of radionuclides in the liquid and solid phases [Bossew 2004]. The convection diffusion model can be described by the diffusive-convective transport (Eq. 3-4), continuity or conservation (Eq. 3-5) and the sorption of the radionuclide (Eq. 3- 5):
∂C (x,t) J(x,t) = −D* L + v*C (x,t) Eq. 3-5 ∂x L
∂C(x,t) ∂J (x,t) = − − λC(x,t) Eq. 3-6 ∂t ∂x
= CS (x,t) kd CL (x,t) Eq. 3-7 where, J(x,t) is the flux (Bq/cm2), x is soil depth (cm) in respect to the surface (x = 0), t is the migration time from the deposition (y), CL(x, t) is the concentration in the liquid 3 3 phase (Bq/cm ), CS(x, t) is the concentration in the sorbed phase (Bq/cm ), C(x, t) is the 3 total volumetric concentration, C(x, t) = CS(x, t) + θ CL(x, t), θ is the water content, cm 3 * 2 water/cm soil, D is the hydrodynamic dispersion coefficient (cm /y), kd is the
105 distribution coefficient, v* is interstitial water flow velocity and λ is the radioactive decay constant of the radionuclide. In Eq. 3-7, a linear sorption is assumed, which is valid under the assumption that the sorption is independent of the radionuclide concentration and is instantaneous and reversible. The combination of the set of the equations above (Eq. 3-5, Eq. 3-6 and Eq. 3-7 ) leads to the so-called convection diffusion equation (CDE):
∂C(x,t) ∂ 2C(x,t) ∂C(x,t) = D − v − λC(x,t) Eq. 3-8 ∂t ∂x 2 ∂x
* 137 * where D = D /Rd, is effective or apparent diffusion coefficient of Cs in soil, v = v /Rd is effective or apparent convective velocity and Rd = θ + Kd is the retardation factor. For solutes moving through a porous medium, the velocity is commonly defined as an average of solutes velocities in all flow paths over a representative elementary volume. Therefore, the velocity variations, caused by heterogeneities at scales smaller than that representative elementary volume, is not described by this average velocity. The hydrodynamic dispersion is used in the transport equation to describe these velocity variations. The hydrodynamic dispersion coefficient D* is a combination of the effective molecular diffusion Dwτw and mechanical dispersion Dh and is usually defined as
* = τ + =η *n D Dw w Dh , Dh v Eq. 3-9 where Dw is the diffusion coefficient in bulk water, τw is a tortuousity factor, η is dispersivity, and n is an empirical coefficient which was found to lie between 1 and 2 [Bear 1972]. It is shown by e.g. Padilla et al. [Padilla 1999] that the dispersivity is inversely proportional to the water content in the porous medium. Eq. 3-8 describes the transport of two physically different concentration modes: flux- averaged concentration, of a solute (measured e.g. in the outflow of a soil column) and the resident concentration (measured, e.g. by soil coring) [Kreft 1978, Parker 1984].
106 Despite of the fact that both of these concentrations satisfy Eq. 3-8, these two types of concentrations differ not only conceptually, but in general, also in magnitude [Bossew 2004]. The flux concentration was used by some authors such as Koblinger-Bokori et al. and Szerbin et al. [Koblinger-Bokori 1996, Szerbin 1999] to describe the depth distributions of cesium in soil. This solution is not the suitable solution for our case. The so-called “resident concentration” is the correct solution for our case which was used by some authors such as [Likar 2001, Bossew 2004]. As boundary condition, a half-infinite space-time is assumed x, t ∈ [0, ∝) and the solution has to be finite i.e. C(x →∝, t) → 0. The initial conditions assumes a pulse-like input at t = 0 i.e. C(x, t = 0) = 0 and J(x = 0,t) = J(x = 0) δ(t) where δ is the delta function. Considering these boundary and initial conditions, the solution takes the form:
= −λt + + ′ −λ (t +t′) C(x,t) C oCh G(x, t)e C oNB G(x, t t )e Eq. 3-10 where,
v( x−vt 2) 2D F π C C + SSV e −x2 4Dt v t v( x+vt 2) 2D x vt G(x,t) = × Ge − e D1− ΦD TTW Eq. 3-11 πDt HG 2 D E E 4Dt UUXW
Where is the error function, which has the form
x 2 − 2 Φ ( x ) = O e t dt Eq. 3-12 π 0 where t is the time span between nuclear bomb tests and Chernobyl accident (it is set to 22 y) and t' is the time span between Chernobyl accident and the date of sampling (about 19 y). Many simplifications have been used to solve Eq. 3-7, where its validity must be tested by checking its ability of describing the vertical depth distributions of the radionuclides and their evolution with time [Bossew 2004]. Some of the important simplifications were discussed by Bossew and Kirchner [Bossew 2004]:
107 1) The migration velocity (v) and the dispersion coefficient (D) are considered to be constant over the soil column. 2) A linear sorption isotherm is used to describe the sorption equilibrium in order to find an analytical solution for the convection dispersion equation, which is not available for a non-linear sorption isotherm. This may be an oversimplification since there are evidences that the sorption of strontium and cesium depends nonlinearly on its concentration in soil [Torstenfeld 1982, Smith 1990, Kirchner 1993, Hsu 1994, Kirchner 1996]. 3) Only two phases are considered for the radionuclide in soil: the mobile phase and the reversibly sorbed phase ignoring the fraction, which could be fixed in the soil and becomes irreversible for the exchange process. 4) The migration velocity (v) and the dispersion coefficient (D) are considered to be constant over time.
3.3.2.3 Results and Discussion
The solution in Eq. 3-9 and Eq. 3-10 was applied to the second set of samples but not to the first one since the first set was thinly sliced. The soil profiles in the first set of samples were sliced into 5 cm, 6 cm or 7 cm thick slices. Thus the measured activity of any slice in the first set of samples represents an average value for that slice. Therefore, it was not possible to determine the positions of peaks precisely. In addition, the measured profile of cesium in soil represents an approximate shape and not a precise one. The second set of samples offered more precise profile (wit higher resolution) of cesium. A Matlab program (V. 7.1.0.246) was build to execute the fit using the fminuit program (v 2.3). Fminuit is an optimization and chi-square fitting program for Matlab and Scilab. Fminuit is based on the MINUIT minimization engine. Minuit is a library of subroutines build by Fortran 77 programming language. Minuit was developed by F. James (version 94.1) at CERN Geneva, Switzerland 1994-1998. Fminuit (v 2.3) is copyright by Giuseppe Allodi (1996-2007). The fit was done by two methods (under two different assumptions):
108 I. First method of fit (Fit1): the fit was performed assuming that the depositions from Chernobyl and from global nuclear tests have the same migration velocity (v) and the same dispersion coefficient (D). II. Second method of fit (Fit2): in this fit, the depositions from Chernobyl and from global nuclear tests were assumed (allowed) to have different migration
velocities (vCh and vNB) and different dispersion coefficients (DCh and DNB).
Fit1 was applied to all samples (Figure 3-2, Figure 3-4, Figure 3-6, Figure 3-8, Figure 3-10 and Figure 3-12). Fit2 was applied to all samples as well (Figure 3-3, Figure 3-5, Figure 3-7, Figure 3-9, Figure 3-11 and Figure 3-13). However, relatively more representative (successful) fits were achieved using Fit2. This could be clearly seen visually from the fits and the values of the sum squares error (SSE), where the SSE values were lower for all fits in which Fit2 was used. In AQ3new, a peak of 137Cs can be clearly seen at about 5 cm depth, which could be attributed to Chernobyl fallout. Most of 137Cs concentration belongs to that peak, which can be directly seen from the profile. In this profile, the concentration of 137Cs due to Chernobyl fallout according to both methods of fit was higher as compared to 137Cs from nuclear bomb tests, where the CsCh:CsNB ratio was 1.6 according to Fit1 and 4.1 according to Fit2. A better fit for the concentrations in the deeper layers in AQ3new was achieved using Fit2 (see Figure 3-2 and Figure 3-3) and the SSE value was lower using Fit2 (1.75×10-5) as compared to Fit1 (1.63×10-5). In AQ4new, two peaks could be clearly distinguished (at 2.5 cm and 7.5 cm depth). Using Fit1, it was not possible to get a fit, which is able to fit the data of these two peaks properly. Fit1 resulted in one peak between these two peaks (at about 5cm depth), which represents an average of them (see Figure 3-4). In contrast, Fit2 resulted in a fit, which represents most of the points and fits these two peaks (Figure 3-5). Moreover, the SSE value of Fit1 (6.25×10-4) was about 7 times the SSE value of Fit2 (8.51×10-5). In AQ5new, there is a peak, which could be clearly seen, at about 5 cm depth and another possible peak at about 13 cm depth. Fit1 underestimates most of the concentration values and overestimates some values (Figure 3-6), whereas Fit2 passes through most of the measured concentration points or within their error ranges (Figure 3-
109 7). Thus, the SSE value of Fit1 (1.87×10-4) was about 13 times the SSE value of Fit2 (1.44×10-5). In AQ6new, there are two peaks, which could be clearly seen, at about 3 cm and 10 cm depths. Fit1 resulted in one peak as an average for these two peaks (Figure 3-8), whereas Fit2 resulted in a fit of two peaks. Fit2 passed through most of the measured points of these two peaks (Figure 3-9), but it did not fit the deeper values as good as Fit1. Thus, the SSE values for Fit1 and Fit2 were mostly the same (5.10×10-5 and 4.87×10-5, respectively). In AQ9new, there are two peaks at about 2 cm and 13 cm depths. Fit1 resulted in one peak as an average for these two peaks (Figure 3-10), whereas Fit2 resulted in a fit of two peaks, which is more representative for the measured points of these two peaks (Figure 3- 11). The SSE value for Fit1 (2.09×10-5) was about two times the SSE value for Fit2 (8.63×10-6). In AQ10new, there is a peak, which could be clearly seen, at about 3 cm depth and another possible peak at about 13 cm depth. Fit1 underestimates and overestimates most of the concentration values (Figure 3-12), whereas Fit2 passes through all the measured concentration points or within there error ranges (Figure 3-13). Thus, the SSE value of Fit1 (8.87×10-5) was about 11 times the SSE value of Fit2 (7.83×10-6). After long time of deposition (Chernobyl ~ 19 y, global nuclear tests ~ 41 y), the most portions of 137Cs were still in the upper 15 cm of soil (AQ3 ~ 97%, AQ4 ~ 100%, AQ5 ~ 95%, AQ6 ~ 95%, AQ9 ~ 80%, AQ10 ~ 100%). This is consistent with soil properties, where high amounts of clay were found in all soils (see Table 3-1). The fit parameters of Fit1 and Fit2 are listed in Table 3-4 and Table 3-5, respectively. Apparently, the soils in northwestern territory of Jordan are contaminated with 137Cs from Chernobyl, as every profile of 137Cs has a peak within the upper 5 cm soil layer, which could be attributed to Chernobyl fallout. In addition, a second peak, which could be attributed to the nuclear bomb tests fallout, can be clearly seen in the deeper soil layers in four of the profiles. This consists with the expectations for a possible contamination from Chernobyl in that area. These expectations relied on some information collected at the beginning of this work such as the metrological data for the rainfall in that area in may 1986 (Table 2-1),
110 the trajectory of the radioactive air masses entered Syria on 7 May 1986 (Figure 1-29), which could pass over the north- western territory of Jordan, the 137Cs profile in a sediment core taken in 1994 from kinneret lake (Figure 2-1), where two peaks of 137Cs were found and the shallower one was attributed to Chernobyl fallout, and the data published by [Al Hamarneh 2003]. 137 The concentrations of Cs in Table 3-4 and Table 3-5 (C0Ch) represent the initial concentrations (i.e. at 1986). These concentrations span a range of 140-2650 Bq/m2 using Fit1 and 130-2350 Bq/m2 using Fit2. These values are much lower than these in the Chernobyl-contaminated territories in Europe (see Table 3-3). This relatively low Chernobyl-contamination in Jordan can be explained by the low rainfall rates, which occurred there in may 1986 (see Table 2-1).
The following points can be noticed in Table 3-4:
• The lowest contamination from Chernobyl fallout was in AQ9new, while the highest was in AQ5new and/or AQ6new (no significant difference between AQ5new and AQ6new statistically). • The convention velocities of 137Cs were low as expected and ranged from 0.09 cm/y in AQ10new to 0.23 cm/y in AQ9new. • The dispersion coefficients ranged from 0.10 cm2/y in AQ4new to 1.49 cm2/y in AQ5new and/or AQ6new (no significant difference between AQ5new and AQ6new statistically). It can be said that the diffusion process was dominant in all sites especially in AQ5new, AQ6new and AQ9new. • The ratio between Chernobyl and nuclear tests fallouts ranged from 0.08 in AQ9new to 1.59 in AQ3new.
The following points can be noticed in Table 3-5:
• The lowest contamination from Chernobyl fallout was in AQ9new, while the highest was in AQ10new and/ AQ5new (no statistically significant difference between AQ10new and AQ5new).
111 137 • The convention velocities of Cs in the upper layers (vCh) were relatively low and ranged from 0.05 cm/y in AQ4new and AQ9new to 0.18 cm/y in AQ5new. 137 • The convection velocities of Cs in the lower layers (vNB) were significantly higher than those in the upper layers except in AQ5new (they have mostly the same value) and ranged from 0.15 cm/y in AQ4new to 0.32 cm/y in AQ3new. 137 • The dispersion coefficients for Cs in the upper layers (DCh) spanned a range of 0.09 cm2/y in AQ4 to 0.42 cm2/y in AQ3new. 137 • The dispersion coefficients for Cs in the lower layers (DNB) span a range of 0.08 cm2/y in AQ4 to 0.52 cm2/y in AQ9new (note: no significant statistical difference between AQ9new and AQ6new).
• The dispersion coefficients for nuclear tests portion (DNB) were significantly higher
than those for Chernobyl (DCh) in AQ5new, AQ6new and AQ9new, which agrees with the assumption of increase in the radionuclides dispersions with the transfer time, whereas these values were not significantly different for AQ3new AQ4new and AQ10new. • The ratio between Chernobyl and nuclear tests fallouts ranged from 0.08 in AQ9new to 4.13 in AQ3new. • The significant difference between the convention velocities in the upper layers
(vCh) and in the lower layers (vNB) could perhaps be attributed to two reasons: 1) The higher cation exchange capacities (CECs) in the upper layers (see Table 3-1), which leads to higher retention of 137Cs in these layers and 2) The higher dispersivity in the lower layers.
It has been pointed out by Kirchner [Kirchner 1998(b)] that the dispersion of radionuclides in soil may increase with the square of the transport time due to the spatial variations in hydrodynamic and sorption properties. Practically, this means that the 137Cs fractions (Ch. and NB.) could have different dispersion coefficients because the time span after the nuclear tests fallout (~ 41 y) was about two times the time span after Chernobyl fallout (~ 19 y). Moreover, soil structure, soil composition, soil density and water content are varied with depth. It can be noticed in Table 3-1 that the CECs were higher for the upper layers
112 as compared to the lower layers and the (Ca + Mg)/K ratios were lower for the lower layers as compared to the upper layers.
C0Ch C0NB v D Sample C0Ch : C0NB (Bq/cm2) (Bq/cm2) (cm/y) (cm2/y) AQ3new 0.126 ± 0.008 0.082 ± 0.013 0.119 ± 0.012 0.420 ± 0.042 1.592 ± 0.002 AQ4new 0.107 ± 0.007 0.447 ± 0.015 0.148 ± 0.018 0.097 ± 0.013 0.239 ± 0.018 AQ5new 0.262 ± 0.058 0.590 ± 0.098 0.134 ± 0.018 1.490 ± 0.198 0.445 ± 0.066 AQ6new 0.265 ± 0.051 0.335 ± 0.085 0.142 ± 0.002 1.494 ± 0.019 0.791 ± 0.044 AQ9new 0.014 ± 0.006 0.171 ± 0.009 0.230 ± 0.012 0.611 ± 0.034 0.081 ± 0.001 AQ10new 0.171 ± 0.014 0.204 ± 0.022 0.088 ± 0.008 0.250± 0.028 0.836 ± 0.007 Table 3-4: The parameters of Fit1.
C0Ch C0NB vCh DCh vNB DNB C0Ch : Sample 2 2 2 2 (Bq/cm ) (Bq/cm ) (cm/y) (cm /y) (cm/y) (cm /y) C0NB 0.157 0.038 0.133 0.423 0.320 0. 474 4.131 AQ3new ± ± ± ± ± ± ± 0.011 0.004 0.009 0.031 0.082 0.037 0.522 0.121 0.487 0.050 0.087 0.154 0.079 0.249 AQ4new ± ± ± ± ± ± ± 0.007 0.018 0.010 0.019 0.008 0.012 0.006 0.196 0.759 0.179 0.393 0.171 1.521 0.258 AQ5new ± ± ± ± ± ± ± 0.064 0.098 0.017 0.135 0.052 0.926 0.074 0.122 0.567 0.107 0.180 0.210 0.489 0.215 AQ6new ± ± ± ± ± ± ± 0.009 0.095 0.015 0.015 0.034 0.132 0.039 0.013 0.178 0.046 0.089 0.245 0.521 0.079 AQ9new ± ± ± ± ± ± ± 0.002 0.011 0.017 0.038 0.079 0.094 0.001 0.235 0.137 0.115 0.124 0.188 0.141 1.715 AQ10new ± ± ± ± ± ± ± 0.011 0.006 0.005 0.009 0.012 0.021 0.003 Table 3-5: The parameters of Fit2 (using different velocities and different dispersion coefficients for Ch. and NB).
113 Jordan has Mediterranean climate with a relatively rainy winter (about 4 months), and very dry for the rest of the year including a hot, dry summer (about 4 months). In such a climate, the upper surface of the soil becomes more dry during the dry seasons as compared to the lower layers, which creates a difference in the water content between the upper and the lower layers (i.e. different conditions for the radionuclides transfer in soil). This could explain the differences in the migration velocities and dispersion coefficients of cesium between the upper layers (Chernobyl portion) and the lower layers (nuclear bomb tests portion). 137 137 The CsCh− CsNB ratios, which were obtained, using Fit1, were significantly different from those, which were obtained using Fit2 in AQ3new, AQ5new, AQ9new and AQ10new while it was nearly same for both of AQ4new and AQ9new. 137 137 The CsCh− CsNB ratios that were obtained using the convection dispersion fit (Fit1 and Fit2) were different, for all sites, from those which were obtained using the 137Cs−90Sr ratio (see Table 3-6). These differences could be attributed to the fact that this ratio represents the average value of the northern hemisphere and not a specific value for Jordan. To determine the exact value of 137Cs−90Sr ratio from nuclear bomb tests in Jordan, records of 137Cs and 90Sr before Chernobyl are required. Such records are not available in Jordan.
Cs : Cs Ch NB Cs : Cs Cs : Cs Sample (Using 137Cs-90Sr Sample Ch NB Ch NB (Using CDE Fit1) (Using CDE Fit2) ratio) AQ3 0.607 ± 0.322 AQ3new 1.592 ± 0.002 4.131 ± 0.522 AQ4 1.133 ± 0.286 AQ4new 0.239 ± 0.018 0.249 ± 0.006 AQ5 0.87 ± 0.298 AQ5new 0.445 ± 0.066 0.258 ± 0.074 AQ6 0.315 ± 0.143 AQ6new 0.791 ± 0.044 0.215 ± 0.039 AQ9 Not determined* AQ9new 0.081 ± 0.001 0.079 ± 0.001 AQ10 0.171 ± 0.307 AQ10new 0.836 ± 0.007 1.715 ± 0.003
Table 3-6: CsCh-CsNB ratios resulting from different methods. * This value could not be determined since 90Sr was not detectable in this sample.
114 In this work, the average value of vNB was about two times the average value of vCh and the average value of DNB was also about two times the average value of DCh. It was found by Barisic et al. [Barisic 1999] that some fractions of 137Cs have reached deeper depths in some clay-rich soils since these soils can have relatively deep desiccation cracks. This could be a probable interpretation for the relatively higher values of 137Cs in the lower tails of the depth profiles of AQ3new (at 19 cm and 23 cm depths), AQ8new (at 19 cm depth) and AQ9new (at 23 cm and 28 cm depths). In AQ9new, the 137Cs profile could be disturbed in the upper 8 centimeters of soil and the peak found by the fit is not representative.
115 0.014 fit exp 0.012 Ch NB 0.01
0.008
0.006 137Cs (Bq/cm3)
0.004
0.002
0 0 5 10 15 20 25 30 Depth (cm)
Figure 3-2: 137Cs depth profile in AQ3new using Fit1.
0.014 fit exp 0.012 Ch NB 0.01
0.008
0.006 137Cs (Bq/cm3)
0.004
0.002
0 0 5 10 15 20 25 30 Depth (cm)
Figure 3-3: 137Cs depth profile in AQ3new using Fit2.
116 0.04 fit 0.035 exp Ch NB 0.03
0.025
0.02
137Cs (Bq/cm3) 0.015
0.01
0.005
0 0 5 10 15 20 25 30 Depth (cm)
Figure 3-4: 137Cs depth profile in AQ4new using Fit1.
0.04 fit 0.035 exp Ch NB 0.03
0.025
0.02
137Cs (Bq/cm3) 0.015
0.01
0.005
0 0 5 10 15 20 25 30 Depth (cm)
Figure 3-5: 137Cs depth profile in AQ4new using Fit2.
117 0.035 fit exp 0.03 Ch NB 0.025
0.02
0.015 137Cs (Bq/cm3)
0.01
0.005
0 0 5 10 15 20 25 30 Depth (cm)
Figure 3-6: 137Cs depth profile in AQ5new using Fit1.
0.035 fit exp 0.03 Ch NB 0.025
0.02
0.015 137Cs (Bq/cm3)
0.01
0.005
0 0 5 10 15 20 25 30 Depth (cm)
Figure 3-7: 137Cs depth profile in AQ5new using Fit2.
118 0.025 fit exp Ch 0.02 NB
0.015
0.01 137Cs (Bq/cm3)
0.005
0 0 5 10 15 20 25 30 Depth (cm)
Figure 3-8: 137Cs depth profile in AQ6new using Fit1.
0.025 fit exp Ch 0.02 NB
0.015
0.01 137Cs (Bq/cm3)
0.005
0 0 5 10 15 20 25 30 Depth (cm)
Figure 3-9: 137Cs depth profile in AQ6new using Fit2.
119 -3 x 10 7 fit exp 6 Ch NB 5
4
3 137Cs (Bq/cm3)
2
1
0 0 5 10 15 20 25 30 Depth (cm)
Figure 3-10: 137Cs depth profile in AQ9new using Fit1.
-3 x 10 7 fit exp 6 Ch NB 5
4
3 137Cs (Bq/cm3)
2
1
0 0 5 10 15 20 25 30 Depth (cm)
Figure 3-11: 137Cs depth profile in AQ9new using Fit2.
120 0.04 fit 0.035 exp Ch NB 0.03
0.025
0.02
137Cs (Bq/cm3) 0.015
0.01
0.005
0 0 5 10 15 20 25 30 Depth (cm)
Figure 3-12: 137Cs depth profile in AQ10new using Fit1.
0.04 fit 0.035 exp Ch NB 0.03
0.025
0.02
137Cs (Bq/cm3) 0.015
0.01
0.005
0 0 5 10 15 20 25 30 Depth (cm)
Figure 3-13: 137Cs depth profile in AQ10new using Fit2.
121 3.3.2.4 Statistical Evaluations of Fit1 and Fit2
Statistically, the F test can be used to compare between models if they were nested (i.e. one of them is a simple case of the other). Considering Fit1 as Model1 and Fit2 as Model2, which is the model with more parameters, SSE1 is usually higher than SSE2 because Model1 has more degrees of freedom (the number of the degrees of freedom (DF) is the number of the data points minus the number of the fitted parameters). The relationship between the relative difference in the sum of square errors, due to the improvement of Model1 to Model2, and the relative difference in degrees of freedom can be quantify by the F ratio:
(SSE1− SSE2) / SSE2 (SSE1− SSE2) /(DF1− DF2) F = = Eq. 3-13 (DF1− DF2) / DF2 SSE2 / DF2
If Model1 is correct, then F ratio is expected to be near 1.0. If it is greater than 1.0, either Model2 is correct or Model1 is correct and the better fit of Model2 was due to random scatter. However, this can be checked by the probability value (P value), which can be found in the distribution tables or calculated by some build-in software packages. The question to be answered by the P value is: If Model1 is correct, what is the probability of obtaining data randomly that fits Model2 much better. Therefore, if P value is low, Model2 is significantly better than Model1. Otherwise, there is no evidence supporting the use of Model2. The F and P values were calculated and presented in Table 3-7. In AQ4new, AQ5new, AQ9new and AQ10new, the use of Fit2 (Model2) is very significantly better than the use of Fit1 (Model1). The difference between Fit1 and Fit2 is not statistically significant in AQ3new and AQ6new. This was because the SSE values resulted from Fit1 are not significantly higher than those resulting from Fit2.
122 Profile SSE1 SSE2 DF1 DF2 F value P value AQ3new 1.75×10-5 1.63×10-5 13 11 0.41 0.6766 AQ4new 6.25×10-4 8.51×10-5 10 8 25.85 0.0003 AQ5new 1.87×10-4 1.44×10-5 12 10 51.64 < 0.0001 AQ6new 5.10×10-5 4.87×10-5 14 12 0.28 0.7581 AQ9new 2.09×10-5 8.63×10-6 14 12 8.53 0.005 AQ10new 8.87×10-5 7.83×10-6 11 9 53.94 < 0.0001 Table 3-7: F test; SSE values, degrees of freedom, calculated F and P values.
3.3.2.5 Comparison with migration parameters from other studies
A comparison with migration parameters from other studies can be seen in Table 3-8. This table includes some information about the soil profiles, their number, sampling date and the locations where the studies were done. In the following are these studies:
A. In [Ivanov 1997], the clay amount was low (less than 5%) in most of the places and the organic matter was low as well (less than 2% in most sites) with two exceptional sites where clay amount was 23.2% and 74-85%. The migration velocities were considerably higher in these two sites. The values of v and D showed a tendency to decrease with time over the test period (1987-1993). The migration parameters in [Ivanov 1997] represent mostly Chernobyl fallout, since the soil samples were collected from the 30-km restriction zone of the Chernobyl. It was pointed out by [Ivanov 1997] that the migration velocities were higher in
wet organic soils. In general, the values of vCh in Jordan (this work) had very similar values to those in [Ivanov 1997] despite the differences in the soil types (mostly clay in Jordan (this work)), time of sampling and amount of 137Cs deposition (much higher in [Ivanov 1997]). In Jordan (this work) the average
value of DCh was similar to the average values of DCh in [Ivanov 1997]. B. In [Szerbin 1999], soils were mostly sandy, while the clay contents were in general not small (6.51- 37.87% with a mean of 21.99%), which means that clay could play an important role in 137Cs retention. It was found out that there is no strong direct correlation between the physico-chemical characteristics of the soils
123 and the migration parameters. The average value of v in [Szerbin 1999] was about two times the average value in Jordan (this work), which could be attributed to the differences in soil types (mostly sand in [Szerbin 1999] and mostly clay in Jordan (this work)). Other factors, which could influence the migration parameters, are the time span after deposition (sampling in Jordan (this work) was 8-10 years later) and rainfall rates (higher in Hungary than in Jordan (this work)). The D values spanned the same range for both of Hungary [Szerbin 1999] and Jordan (this work), while the average value of D for Jordan (this work) was a bit higher than the average value of D for Hungary [Szerbin 1999]. C. Even that the soils in southern Costa Rica [Bossew 2001] were taken from rainy forests with sandy type and they were sampled about 9 years earlier than those
from Jordan (this work), the range and the average value of vNB were lower than
those in Jordan (this work), while the average value of DNB was higher for Costa Rica as compared to Jordan (this work). A possible interpretation could be due to the plants cover that can be found in such a rainy area, which absorbs 137Cs from soil creating upward migration. This is a possible interpretation, which could be not the only or the right reason since no information has been mentioned about that in [Bossew 2001]. Moreover, the results in [Bossew 2001] were considered as rough estimates because the soil cores were divided into only 3 layers (0-5, 5-10 and 10-15 cm in most cases). D. In Central Serbia [Krstic 2004], many of the soil samples were clay-rich especially the vertisol soils. Sampling date was only 4 years earlier than that in Jordan (this work). The correlation between the soil characteristics and v was considered to be weak. The values of v and D in [Krstic 2004] were similar to those in Jordan (this work). E. A large number of soil profiles (about 328) were mostly taken between 1987-1989 from six regions of different landscape types and geology in Austria by [Bossew 2004]. Every soil profile represents one core or one cubic sample (with an area of 18 × 18 cm2 and a depth of 10-20 cm). Physical and chemical analyses were not
performed due the large number of samples. The average value of vNB was about
half the value of vNB in Jordan (this work) despite that the sampling date in [Bossew 2004] was about 16-18 years earlier and the annual rainfall in Austria is
124 higher than in Jordan. The average value of DNB, which represents a geometric mean, was relatively low but with a high standard deviation. However, physical and chemical analyses of soils in [Bossew 2004], which are not available, would be required in order to perform a more precise comparison to the data of Jordan (this work). F. In [Almgren 2006], soil samples were collected from western Sweden in 2003 and no soil physical and chemical analyses were performed. The ranges and the average values of v and D were a bit higher than those in Jordan (this work). The sampling date in [Almgren 2006] was just 2 years earlier than the sampling date in this work (Jordan). A more precise comparison to the results of Jordan (this work) was not possible since the physical and chemical characteristics of the soils in [Almgren 2006] were not available.
In general, the average values of v in Table 3-8 don not vary so much (they span a range from 0.1 to 0.2 cm/y) and relatively high values of v (more than 0.3 cm/y) were found in some profiles but rarely. The values of D have a tendency to increase with time after deposition.
125 Reference Location Sampling v (± 1SDE) (cm/y) D (± 1SDE) (cm2/y) [Mean Comments date [Mean value] value] A. 30-km 1987, Region1 (BP): Region1 (BP): - 37 profiles from 37 different sites. Ivanov zone of 1988 1992: 0.170 1992: 0.110 - Standard deviation: 20 – 50 %. 1997 ChNPP, 1990 – - *: mostly organic matter soil (74 – 85 %). Ukraine, 1993 Region2 (UAP): Region2 (UAP): - +: Gley sand Belarus 1992: 0.047 – 0.347 [0.183] 1992: 0.095 – 0.820 [0.277] - The given data of v and D represent: and 1987: 8 sites in UPP and 8 sites in UIP. Russia Region3 (UPP): Region3 (UPP): 1988: 8 sites in UPP and 5 sites UIP. 1987: 0.050 – 0.155 [0.101] 1987: 0.155 – 0.505 [0.263] 1990: 6 sites UIP. 1988: 0.079 – 0.177 [0.123] 1988: 0.079 – 0.568 [0.242] 1991: 2 sites UIP. 1993: 0.054 – 0.158 [0.105] 1993: 0.041 – 1.419+ [0.467] 1992: 1 site in BP, 10 sites in UIP, 4 sites in UAP. 1993: 7 sites in UPP, 11 sites in UIP and 7 sites in Region4 (UIP): Region4 (UIP): RKP. 1987: 0.035 – 0.347 [0.125] 1987: 0.167 – 0.568 [0.291] - Clay ranges: 5.8 - 12.3% in UPP, 0.0 - 9.6% in UIP and 1988: 0.079 – 0.410 [0.178] 1988: 0.142 – 0.536 [0.255] <0.3 - 4.9% in RKP. 1990: 0.047 – 0.101 [0.071] 1990: 0.095 – 0.249 [0.137] - Sand ranges: 44.9 - 68.9% in UPP, 38.5 - 90.0% in UIP and 1991: 0.057 – 0.145 [0.101] 1991: 0.066 – 0.066 [0.066] 45.0 - 68.1% in RKP. 1992: 0.022 – 0.505 [0.135] 1992: 0.076 – 0.315 [0.159] - Organic matter: : 0.3 - 1.2% in UPP, 0.0 - 3.5% in UIP and 1993: 0.032 – 0.155 [0.087] 1993: 0.057 – 0.315 [0.167] <1.0 - 7.5% in RKP and 1site with 74 - 85% in RKP. - Soil in UAP is sandy in two profiles and sandy loamy the Region5 (RKP): Region5 (RKP): other two. 1993: 0.066 – 0.896* [0.316] 1993: 0.061 – 0.691 [0.250] B. Hungary 1995 and 0.056 – 0.77 [0.254] 0.051 – 1.46 [0.548] - 19 profiles from 19 counties (all counties in Hungary). Szerbin 1997 - Soil samples are mostly sandy. 1999 - Sand: 41.39 - 87.1% (Mean 69.49%), Clay: 6.51- 37.87% (21.99%) and OM: 1.05- 8.2% (3.35%).
C. Southern 1996 vNB: DNB: - 4 locations in a tropical rainforest with high-rainfall. Bossew Costa 0.09(0.07) – 0.16(0.08) 0.68(0.33) – 1.02(0.72) - 5 cores per location (13 fitted). 2001 Rica [0.14 (0.09)] [0.79 (0.49)] - v and D values represent sites averages. - Top layer (1-2 cm thick) is an organic layer. - Sandy texture and low humus content. - Results are rough estimates (5 cm thick layers). Table 3-8: Migration parameters of 137Cs found in this work and other works. continue;
126 Reference Location Sampling v (± 1SDE) (cm/y) D (± 1SDE) (cm2/y) [Mean Comments date [Mean value] value] D. Central 2001 CDE model (solution1): CDE model (solution 1): -10 profiles have been taken from 10 different locations. Krstic Serbia 0.00 – 0.31 [0.17] 0.24 – 1.45 [0.55] -solution1: normal distribution solution of CDE. 2004 - solution2: error function solution of CDE. CDE model (solution2): CDE model (solution 2): - Soil physical analysis has not been done. 0.00 – 0.26 [0.07] 0.34 – 1.47 [0.76] - A peak appeared in only 4 profiles despite they were sliced into 2 cm layers. Thus, v was 0.00 in the other 6 profiles using solution2 and around 0.10 using solution1. Soils in these 6 profiles were vertisol (clay-rich) in 3 profiles, gray brown in 2 profiles and brown-forest in the last one.
E. Austria 1987 – vNB: DNB: - 328 profiles from 6 regions of different landscape type Bossew 1989 0.086(0.041) – 0.191(0.047) 0.045(3.2) – 0.36(2.8) and geology. 2004 [0.113 (0.063)] [0.087 (4.2)] - Most of the profiles investigated in this study were taken between 1987 and 1989. The mean value represent The mean value represent - v values represent arithmetic means of the regions. 308 profiles. 328 profiles. - D values represent geometric means of the regions. - Soil physical and chemical analyses of the individual samples couldn’t be performed due to there large number. F. Western 2003 0.00(0.11) – 0.35(0.03) 0.06 (0.00) – 2.63 (1.88) - 33 profiles has been taken (one of them couldn’t be fitted). Almgren Sweden [0.21 (0.07)] [0.820 (0.42)] - Soil physical and chemical analyses were not done. 2006 E. North- 2005 v: D: - 6 profiles has been taken from 6 different sites. This work western 0.088(0.008) – 0.230(0.012) 0.097(0.013) – 1.494(0.019) - Clay amount: 43.3 – 68.7% with an average of 54.9%. Jordan [0.144 (0.012)] [0.727 (0.056)] - Sand amount: 1.2 – 18.1% with an average of 6.2%. vCh: DCh: - Organic matter: 1.6 – 5.6% with an average of 2.9% 0.046(0.017) – 0.179(0.017) 0.087(0.019) – 0.393(0.135) - v and D have been obtained using Fit1. [0.105 (0.012)] [0.216 (0.041)] - vCh, vNB, DCh, DNB have been obtained using Fit2. vNB: DNB: 0.154(0.008) 0.320(0.082) 0.079(0.012) – 1.521(0.926) [0.215 (0.045)] [0.459 (0.204)] Table 3-8 (continued): Migration parameters of 137Cs found in this work and other works.
127 3.3.3 Correlation between the Annual Rainfall in the Sites and 137Cs Inventories (Climate Effects)
The correlation between the annual rainfall and global fallout 137Cs deposition was studied by many groups. Several authors found a positive linear relationship between them [e.g. Cox 1984, Bunzl 1988, Arnalds 1989, Blagoeva 1995, Legarda 2001]. Annual rainfall data for many sites in Jordan are available online the Jordanian Meteorological Department [JMD] website. Annual rainfall data are available for most of the meteorological stations from the year of 1937 or 1960 till 2003. These data were used to calculate average annual rainfall for every sampling site. The average annual rainfall for any sampling site was considered to be the average annual rainfall of the closest station to that site, where the stations cover most of the area of the north western part of Jordan. Since the Chernobyl deposition was over a short period (some days), the correlation was tested between the annual rainfall and the fallout from the nuclear bombs tests fallout. The portion of 137Cs from the nuclear bomb tests fallout was calculated according to the “137Cs-90Sr” ratio (Section 3.4.1). Figure 3-14 shows some correlation between the average annual rainfall and 137Cs deposition where a positive linear relationship with 0.69 (R2 = 0.48) correlation factor.
2500
2000
1500 (Bq/m2) NB 1000 137Cs 500
0 0 100 200 300 400 500 600 700 Annual Rainfall (mm/yr)
137 Figure 3-14: CsNB inventories vs. sites average annual rainfall.
128 3.3.4 Correlation between the Sites Altitudes and 137Cs Inventories
The deposition of 137Cs was expected to correlate with the altitudes of sites where for higher places the possibility for the wet and/or the dry precipitation 137Cs is higher. Since the rainfall is dependent on the topography of an area, the 137Cs deposition should be similarly influenced. Fallout of 137Cs from Chernobyl was almost in the form of wet deposition in 1986 [Clarke 1988], which was also the case in the deposition in the Middle East. The hypothesis that deposition of fallout to soils increases with altitude was proven by McGee E. J. et al. [McGee 1992] in his study in northwestern Ireland in 1990. The results of this study provided evidence that greater 137Cs deposition to soils occurred at higher altitudes, which also supported the findings of a previous investigation in the same area. Figure 3-15 shows some correlation between 137Cs inventories and the sites altitudes (R2 = 0.49) where 137Cs deposition increases with increasing altitudes. The results of the first set of samples were used to study the correlation since they are more representative. Actually, the linear relationship used here is not the only possibility to describe the relation between altitudes and 137Cs inventories in this work. A quadratic relationship, for example, could describe them in a good way. However, the correlation between 137Cs inventories and the sites altitudes can be clearly seen in Figure 3-15 regardless the type of that correlation.
4500 ) 4000 3500 3000 2500 2000 1500 1000
137Cs Inventory (Bq/m2 500 0 0 200 400 600 800 1000 1200 Altitude (m)
Figure 3-15: 137Cs inventories vs. sites altitudes.
129 4 Conclusions and Outlook
The main aims of this work were to study the contamination in Jordanian soils due to 137Cs, to determine the fraction of 137Cs from Chernobyl fallout, if any, and the fraction from nuclear bomb tests fallout, to study the migration of 137Cs in the Jordanian soils, to find out if ist presence in soil represents a risk for public health and to compare our results to those from neighboring countries and countries with different climate types from Europe. This study has fulfilled its aims successfully. Apparently, the northwestern part of Jordan has been contaminated by 137Cs from Chernobyl. The contamination of 137Cs in Jordan due to Chernobyl is significant as compared to that from the nuclear bomb tests. The ratio of CsCh-CsNB was calculated according to three methods and was ranging from 0.08-0.17 as a minimum value to 1.13-4.13 as a maximum value. As compared to the neighboring countries, lower contamination of 137Cs was found in the Jordanian soils except in Egypt, which had comparable or lower values. This contamination was also higher in many European countries, where they received higher amounts of 137Cs from Chernobyl. Most of 137Cs is still in the upper layers of soils (0-15 cm), which implies that it is still available for the plants uptake and in turns to animals and humans. Therefore, it represents a possible source for the internal and external doses. The values of the calculated effective dose equivalent due 137Cs in soil at a height of 1 m above the soil surface were low and do not represent a significant hazard on the public health. They were also relatively low as compared to the neighboring countries. A good positive correlation was found between 137Cs depositions, and each of the sites average annual rainfalls and sites altitudes. The migration parameters of 137Cs in soil were calculated by fitting the soil profiles to the convention dispersion equation (CDE) model. These fits were done using two methods (Fit 1 and Fit2). Fit1 was done assuming that the depositions from Chernobyl and from Global nuclear tests have the same migration velocity and the same dispersion coefficient, while Fit2 assumes different migration velocities (vCh and vNB) and different dispersion coefficients (DCh and DNB).
130 In general, comparable high values of dispersion coefficients were found. This implies that the diffusion process was dominant in all sites. This could be attributed to the low contents of water in the Jordanian soils since the dispersivity is inversely proportional to water content in a porous media. The high contents of clay were supposed to play the main role in 137Cs retaining in soil, while the amounts of organic matter and the pH values were considered to have a neglected effect on 137Cs migration in soil. The values of the migration velocity were compared to those from different climate types in Europe, the 30-km zone of ChNPP and Costa Rica in Central America. The average values of the migration velocity were not highly varied (they span a range from 0.1 to 0.2 cm/y) and relatively high values of the migration velocity (more than 0.3 cm/y) were found in some profiles but rarely. Thereafter, we conclude that the migration of 137Cs in soil is a slow process and the influence of the climate type, if any, is very low. In the last comparison, the values of the dispersion coefficient showed a tendency to increase with time after deposition. The fits achieved using Fit2 were visually more representative. Moreover, using the so- called “F-test”, Fit2 was found to be statistically significant better than Fit1 in 4 profiles out of 6. Despite that Fit2 was more descriptive for 137Cs profiles in this work, more work has to be done in this field in order to test the applicability of this method and the assumptions used to justify it. This can be done by applying it on more profiles taken in relatively long time after Chernobyl, from different soil types and different climate types. External dose due 137Cs in soil has been studied in this work but not the internal dose. Studying the internal dose could be an object for another project. For such study, plant samples have to be taken from the areas of interest in addition to meat and milk samples from herds Grazing there.
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