distribution coefficients of caesium, chlorine, iodine, niobium

February 2014

Working Reports contain information on work in progress

or pending completion.

Mervi Söderlund, Merja Lusa, Sinikka Virtanen,

I lkka Väl imaa, Martt i Hakanen, Jukka Lehto

University of Helsinki

Department of Chemistry

Laboratory of Radiochemistry

Anne-Maj Lahdenperä

Saanio & Riekkola Oy

Working Report 2013-68

Distribution Coefficients of Caesium,Chlorine, Iodine, Niobium, Selenium and

Technetium on Olkiluoto Soils

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Distribution coefficients of caesium, chlorine, iodine, niobium, selenium and technetium on Olkiluoto Soils

ABSTRACT

Retention of caesium, chlorine, iodine, niobium, selenium and technetium was investigated on soil samples from Olkiluoto using laboratory batch sorption experiments. Distribution coefficients were measured for both dried and sieved and untreated (wet, not sieved) mineral soil and humus in aerobic and anaerobic conditions. Mineralogical composition of the samples was determined by XRD-analysis.

Caesium was sorbed efficiently on mineral soil samples and less efficiently on humus. Sorption decreased with decreasing cation exchange capacity and clay fraction content. The effect of competing cations decreased in the order Cs+>NH4

+>K+>Ca2+>Na+.

Chlorine was not retained by mineral soil samples, and the sorption was weak on humus.

The sorption of iodine was the strongest on humus and the weakest on the untreated mineral soil samples in the anaerobic conditions. In the mineral soil samples, the sorption decreased with decreasing organic matter content and increasing pH.

The retention of niobium on soil samples was the most efficient among the studied elements. The retention was high regardless of the aeration conditions. Sorption on humus was smaller.

Selenium was retained efficiently on humus. Sorption on mineral soil samples was stronger in aerobic conditions. Sorption increased with time.

Technetium was sorbed well on humus and anaerobic, untreated mineral soil samples. Sorption increased with increasing organic matter content and decreasing redox potential.

The results from the sorption experiments are used in the site specific radionuclide migration modelling.

Keywords: Distribution coefficient (Kd), caesium, chlorine, iodine, technetium, sorption, soil, organic matter.

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Cesiumin, kloorin, jodin, niobiumin, seleenin ja teknetiumin jakaantumiskertoimet Olkiluodon maaperänäytteissä

TIIVISTELMÄ

Cesiumin, kloorin, jodin, niobiumin, seleenin ja teknetiumin pidättymistä Olkiluodon maaperänäytteisiin tutkittiin laboratoriossa kokeellisissa eräsorptiokokeissa. Maaperä-näytteet joko kuivattiin ja seulottiin tai käytettiin käsittelemättöminä. Näytteiden mine-raloginen koostumus määritettiin XRD-analyysillä. Sorptiokokeita tehtiin mineraali-maa- ja humusnäytteille hapellisissa ja hapettomissa olosuhteissa.

Cesium pidättyi tehokkaasi mineraalimaanäytteisiin ja heikommin humukseen. Sorptio heikkeni vähenevän kationinvaihtokapasiteetin ja savifraktion funktiona. Kilpailevien kationien vaikutus pieneni järjestyksessä Cs+>NH4

+>K+>Ca2+>Na+.

Kloori ei pidättynyt mineraalinäytteisiin ja sorptio oli heikkoa humukseen.

Jodin sorptio oli voimakkainta humukseen ja heikointa käsittelemättömiin mineraali-maanäytteisiin hapettomissa olosuhteissa. Mineraalimaassa sorptio väheni orgaanisen aineksen pitoisuuden vähetessä ja pH:n kasvaessa.

Niobiumin pidättyminen maaperänäytteisiin oli tehokkainta tutkittujen aineiden jou-kossa. Niobium pidättyi tehokkaasti mineraalinäytteisiin hapellisissa ja hapettomissa olosuhteissa. Sorptio humusnäytteisiin oli heikompaa.

Seleeni pidättyi tehokkaasti maaperän orgaaniseen ainekseen, kun taas sorptio mine-raalimaanäytteisiin oli voimakkaampaa hapellisissa olosuhteissa. Sorptio kasvoi ajan funktiona.

Teknetium pidättyi hyvin käsittelemättömään humukseen ja hapettomiin mineraali-maanäytteisiin. Sorptio voimistui kasvavan orgaanisen aineksen pitoisuuden ja pienene-vän redox potentiaalin funktiona.

Sorptiokoetuloksia käytetään radionuklidien paikkakohtaisessa migraatiomallinnuk-sessa.

Avainsanat: Jakaantumiskerroin (Kd), cesium, kloori, jodi, teknetium, (ad)sorptio, maaperä, orgaaninen aines.

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TABLE OF CONTENTS

ABSTRACT

TIIVISTELMÄ

1 INTRODUCTION ..................................................................................................... 5

1.1 Posiva’s sorption experiment programme in HYRL .......................................... 5

1.2 Relevance of radionuclides in sorption studies ................................................ 7

1.3 Distribution coefficient Kd ................................................................................. 8

1.4 Overburden properties at Olkiluoto .................................................................. 9

2 MATERIALS .......................................................................................................... 13

2.1 Sampling ........................................................................................................ 13

2.1.1 Sampling locations .................................................................................. 13

2.1.2 General description of soil sampling ....................................................... 14

2.2 Characterisation of the soil samples .............................................................. 23

2.2.1 Chemical characteristics ......................................................................... 23

2.2.2 Geotechnical characteristics ................................................................... 27

2.2.3 Mineralogy ............................................................................................... 32

2.3 Water samples ............................................................................................... 34

2.3.1 Soil solution sampling and chemical characteristics ............................... 34

2.3.2 Soil solution simulant .............................................................................. 40

2.3.3 Electrolyte and buffer solutions ............................................................... 40

2.3.4 Radioactive tracers ................................................................................. 41

3 METHODS ............................................................................................................. 43

3.1 Analysis of waters .......................................................................................... 43

3.1.1 Cation and anion analysis ....................................................................... 43

3.1.2 pH measurement ..................................................................................... 43

3.1.3 Redox potential measurement ................................................................ 43

3.1.4 Tracer activity measurement ................................................................... 44

3.2 Batch sorption studies .................................................................................... 44

3.2.1 Experiment series programme ................................................................ 45

3.2.2 Aerobic soil samples ............................................................................... 48

3.2.3 Anaerobic soil samples ........................................................................... 48

3.2.4 Untreated humus samples ...................................................................... 50

3.2.5 Irradiated humus samples ....................................................................... 50

4 RESULTS .............................................................................................................. 51

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4.1 Kd values for caesium ..................................................................................... 51

4.1.1 Dried and fractioned soil samples of OL-KK14, OL-KK15 and OL-KK16 51

4.1.2 Aerobic untreated soil samples of OL-KK20 and OL-KK21 .................... 55

4.1.3 Anaerobic untreated soil samples of OL-KK20 ....................................... 57

4.1.4 Untreated humus samples of OL-KK21 .................................................. 58

4.1.5 Effect of competing cations on the sorption of caesium .......................... 59

4.1.6 Recommendations of the Kd values for caesium ..................................... 64

4.2 Kd values for chlorine ..................................................................................... 66

4.2.1 Aerobic untreated soil samples of OL-KK20 ........................................... 66

4.2.2 Dried and fractioned humus samples of OL-KK15 .................................. 67


4.2.4 Irradiated humus samples of OL-KK21 ................................................... 69

4.2.5 Recommendations of the Kd values for chlorine ..................................... 69

4.3 Kd values for iodine ........................................................................................ 70


4.3.2 Aerobic untreated soil samples of OL-KK20 and OL-KK21 .................... 73




4.3.6 Recommendations of the Kd values for iodine ........................................ 77

4.4 Kd values for niobium ..................................................................................... 79




4.4.4 Recommendations of the Kd values for niobium ..................................... 85

4.5 Kd values for selenium .................................................................................... 86




4.5.4 Recommendations of the Kd values for selenium .................................... 91

4.6 Kd values for technetium ................................................................................ 92



4.6.3 Anaerobic untreated soil samples of OL-KK20 and OL-KK21 ................ 95

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4.6.6 Recommendations of the Kd values for technetium ................................ 98

5 SUMMARY .......................................................................................................... 101

REFERENCES ........................................................................................................... 107

APPENDIX 1. OL-KK20 and OL-KK21 soil pH ........................................................... 123

APPENDIX 2. OL-KK20 and OL-KK21 organic and dry matter content ..................... 124

APPENDIX 3. OL-KK20 and OL-KK21 mineralogy .................................................... 125

APPENDIX 4. OL-KK14, OL-KK20 and OL-KK21 specific surface areas .................. 127

APPENDIX 5. Kd values of caesium ........................................................................... 129

APPENDIX 6. Kd values of chlorine ............................................................................ 136

APPENDIX 7. Kd values of iodine ............................................................................... 138

APPENDIX 8. Kd values of niobium ............................................................................ 142

APPENDIX 9. Kd values of selenium .......................................................................... 145

APPENDIX 10. Kd values of technetium ..................................................................... 148

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1 INTRODUCTION

A KBS-3-type repository for the spent fuel from the Finnish nuclear power reactors in Olkiluoto and Loviisa is to be built in the bedrock at the Olkiluoto site (Posiva 2009). In the current repository design, the repository is constructed on a single level and the floor of the deposition tunnels is at the depth of between 400-450 m in the Olkiluoto bedrock.

The repository programme includes a long-term safety case, where also the potential radiological impacts on humans and other biota are evaluated (Hjerpe et al. 2010).

The biosphere behaviour evaluation of the radionuclides includes aspects concerning the migration of radionuclides in soils, sorption on soil particles and transfer from soil to plants. In the transport modelling calculations, the sorption of radionuclides on the solid phase is taken into consideration through the radionuclide distribution coefficient (Kd) value. Kd is dependent upon the element, soil properties and hydrogeochemical conditions.

The main mineralogical and chemical characteristics of the different soil layers, including the soil solution, that determine the sorption of radionuclides in the overburden consist of the type of the clay minerals, clay and organic matter content, particle size distribution, redox potential and pH. Radionuclides having a high tendency of forming anionic species in aqueous solutions, such as chlorine, iodine, selenium and technetium, are weakly sorbed on soil mineral constituents. This leads to high potential mobility in the overburden. The sink for the anionic radionuclides is typically the organic matter. On the other hand, cationic radionuclides such as caesium, are sorbed effectively on the clay minerals. Thus, it is important to understand the behaviour of the specific elements in the defined soil type, and to be able to evaluate their behaviour in the future in similar soils based on the current knowledge.

1.1 Posiva’s sorption experiment programme in HYRL

A research project on the overburden migration of long-lived radionuclides originating from the spent nuclear fuel was launched in the laboratory of Radiochemistry (HYRL) in 2008. The responsible research director of the “biosphere project” in HYRL is professor Jukka Lehto. The responsible researchers during the project include MSc. Mervi Söderlund, MSc. Merja Lusa and Dr. Kaisa Vaaramaa. The research is funded by Posiva Oy. The coordinating person in Posiva is Ari T. K. Ikonen, and Anne-Maj Lahdenperä from Saanio & Riekkola Oy is involved as a consult.

The research scope of the biosphere project is to gain information on the safety assessment of spent nuclear fuel, and especially on the migration of radionuclides in the biosphere. The behaviour of radionuclides in the chosen material, e.g. soil or mire samples, is studied in the laboratory by sorption experiments in the circumstances relevant to the final disposal conditions. Distribution coefficients, Kd values, determined in the experiments can be used in the modelling of the mobility and migration of radionuclides. The sorption research includes the steps of sampling and material characterisation; sorption experiments in the desired conditions e.g. as a function of pH; the use of computational methods in the planning of the experiments and evaluation of the results; the speciation analysis of the redox sensitive elements (I and Se) and

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estimation of the effect of speciation on the sorption; analysis, evaluation and reporting of the results.

By the year 2012, the sorption experiments were mainly done for the Olkiluoto soil samples. The sorption experiments with the soil samples of a well-developed podzol soil of an inland reference site will be done on a later occasion, during 2015. With these samples, fewer tests are to be conducted and the special interest will lie in the eluvial (leaching) and illuvial (enriching) layers typical of podzols.

In 2011, the first mire samples were taken from Lastensuo in Eurajoki, and their characterisation is almost complete. The sorption experiments with the mire samples began in 2012 and are still ongoing. More samples of Posiva’s mire reference sites (Haapanen et al. 2010) are to be collected, characterised and used in the experiments. In 2014 or 2015, sediment samples are collected of Posiva’s reference lakes (Haapanen et al. 2010) and from the surface sea sediments in the vicinity of Olkiluoto Island (Lahdenperä & Keskinen 2011). The sorption experiments with the sea and lake sediment samples are to be conducted in 2015 and 2016. The research and sampling schedule of the biosphere project is presented in Table 1.

Table 1. Research and sampling schedule of the Posiva’s biosphere project in HYRL.

Year Soil Mire Lake Sea

2008

sampling of OL-KK14, OL-KK15 and OL-KK16 & sorption experiments with Cs, I

and Tc

2009 sampling of OL-KK20

& sorption experiments with Cs, I

and Tc

2010 sampling of OL-KK21

& sorption experiments with Cs,

Cl, I, Nb and Tc

2011 sorption experiments

with I, Mo, Nb, Se and Tc & speciation

analysis of I

sampling of Lastensuo


with Cs, I, Mo and Nb & selectivity

coefficients of Cs

sorption experiments with Cs, Cl, I, Se and

Tc

2013 sampling

& sorption experiments


2015 podzol sampling & sorption experiments

sampling


sampling


2016 sorption experiments sorption experiments

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1.2 Relevance of radionuclides in sorption studies

In the long-term dose assessment of spent nuclear fuel certain key radionuclides (elements) have been identified (Hjerpe et al. 2010). These are the nuclides expected to be transported to the biosphere within the dose assessment window of 10 000 years and induce the major radiation dose to humans in the biosphere calculation cases.

Key radionuclides are further divided into top priority and high priority radionuclides. The top priority radionuclides C-14, Cl-36 and I-129, are the ones to dominate the radiation dose induced to humans in the most realistic biosphere model calculations. The high priority radionuclides are expected to give a substantial contribution to the radiation doses in a few calculation cases. The high priority radionuclides include three subgroups (I-III). Table 2 summarises the radionuclides included in the top and high (I-III) priority groups in the 2009 dose assessment (Hjerpe et al. 2010); in the forthcoming assessments, however, some changes are expected (based on the preliminary information, silver isotope(s) are rising into high priority groups whereas some nuclides will become less significant).

Table 2. The priority grouping of radionuclides based on their safety relevance in the long-term safety of spent nuclear fuel (Hjerpe et al. 2010).

Top priority High priority (I) High priority (II) High priority (III)

C-14 Cl-36 I-129

Mo-93 Nb-93m Nb-94 Cs-135

Ni-59 Se-79 Sr-90 Y-90

Pd-107 Sn-126 Sb-126

In the safety assessment, C-14 is handled with specific activity models not requiring Kd estimates because of its abundance in the biosphere and major role in biochemistry (Hjerpe et al. 2010). On the other hand, the contribution of fission nuclide I-129 to the intermediate and high level radioactive waste is significant due to its high yield on the fission of U-235 and Pu-239 atoms, 0.6 % and 1.6 % respectively, long physical half-life (1.57x107 y) and high potential mobility in soil (Ashworth et al. 2003; Firestone et al. 1998; Hu et al. 2005; Robens & Aumann 1988). Like I-129, Cl-36 has long physical half-life (3.01x105 y) and high mobility in soil, but also high soil-to-plant transfer and bioavailability (Ashworth & Shaw 2006b; Colle et al. 2005; Firestone et al. 1998; Sheppard et al. 1996). Cl-36 is formed in the neutron capture of stable Cl-35 by the reaction 35Cl(n,γ)36Cl in the fuel and surrounding metal parts (Sheppard et al. 1996).

Nb-93m, Nb-94 and Cs-135 are included in the high priority (I) group of radionuclides in the long-term safety of spent nuclear fuel (Hjerpe et al. 2010). Nb-93m and Nb-94 are produced in the neutron activation of stable Nb-93 in the metal parts surrounding the fuel by the reaction Nb-93(n,n)Nb-93m and Nb-93(n,γ)Nb-94, but also in the fission of U-235 and Pu-239 (fission yields 1.8x10-11 and 2.3x10-10 % for Nb-93m and 9.4x10-4 and 2.5x10-7 % for Nb-94, respectively) (Andersson et al. 1979; Firestone et al. 1998). The main source of Nb-94 is the activation reaction, but the primary source of Nb-93m is the decay of Mo-93 (Firestone et al. 1998). The physical half-life of Nb-93m and Nb-94 are 16.13 and 2.03x104 years, respectively. Unlike previously mentioned

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radionuclides, Cs-135 exist solely in the cationic form (Torstenfelt et al. 1982). Even so, the importance of Cs-135 in the safety analysis arises from its high yield on the fission of U-235 (6.5 %) and long physical half-life of 2.3x106 years (Firestone et al. 1998). The migration rate of caesium in soils is very low due to the efficient retention on clay minerals (Bürmann et al. 1994, Kónya et al. 2005, Zygmunt et al. 1998).

The most important radioisotope of selenium in the spent nuclear is Se-79, which is classified to the high priority (II) group (Hjerpe et al. 2010). Se-79 is a fission product and a neutron activation product formed by the reaction Se-78(n,γ)Se-79 from stable Se-78. It has the physical half-life of 3.77x105 years and high potential mobility in soil due to its anionic nature (Shaw & Ashworth 2010).

Tc-99 is one of the main fission products formed in the nuclear fission of U-235 and Pu-239 with a total fission yield of 6.13 % (Ashworth & Shaw 2005; Denys et al. 2003; Lieser & Bauscher 1987; Sheppard et al. 1983). Because of its high yield, Tc-99 is one of the most abundant elements in the intermediate and high activity level nuclear waste, with substantial contribution to the activity level after 100 000 years (Abdelouas et al. 2005; Ashworth & Shaw 2005; Denys et al. 2003; Lieser & Bauscher 1987; Palmer & Meyer 1981). Tc-99 has been previously presumed to be a major radiation source for humans along with I-129 because of its long physical half-life (2.13x 105 y) and high mobility in soil as pertechnetate (TcO4

-) ion (Ashworth & Shaw 2005; Denys et al. 2003; Lieser & Bauscher 1987; Palmer & Meyer 1981), but in Posiva’s long-term safety analysis of spent nuclear fuel Tc-99 has not been a key radionuclide of concern (Hjerpe et al. 2010) due to its minor release to the biosphere within the 10 000-year dose assessment time window.

1.3 Distribution coefficient Kd

The sorption of Cs, Cl, I, Nb, Se and Tc to the Olkiluoto soil samples was studied by batch sorption experiments. The used tracers were Cs-134, Cl-36, I-125, Nb-95, Se-75 and Tc-99. Part of the soil samples were dried and sieved (< 2 mm), whereas others were untreated (wet, not sieved). Sorption was studied in aerobic and anaerobic conditions.

The biosphere assessment of the repository long-term safety includes aspects concerning the migration of radionuclides in soils, sorption to soil particles and transfer from soil to plants. The distribution coefficient, Kd, is used to describe the retention of substances on the solid phase. Kd value is defined (e.g. Harjula 2000) as the concentration ratio of the radionuclide between solid and liquid phases in the equilibrium state (Equation 1):

[1]

Equation 1. The definition of the distribution coefficient Kd. Cs = concentration of the element in the solid phase and Cl = concentration of the element in the liquid phase.

High Kd value refers to a considerable retention of radionuclide in soil or sediment systems and low concentration in the liquid phase, whereas for low Kd values the retention is weak and radionuclide concentration in solution is high. Kd is widely used as an indicator of the potential mobility of radionuclides in the environment since at a given water flow rate the distribution coefficient is proportional to the transport rate of

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the radionuclide. At Kd values ≤0.1 ml/g an element can be considered as mobile as water, whereas an element with Kd value of ≥10 000 ml/g is practically immobile within the specified time boundaries (Sheppard et al. 2009b).

High Kd value refers to a considerable retention of radionuclide in soil or sediment systems and low concentration in the liquid phase, whereas for low Kd values the retention is weak and radionuclide concentration in solution is high. Kd is widely used as an indicator of the potential mobility of radionuclides in the environment since at a given water flow rate the distribution coefficient is proportional to the transport rate of the radionuclide. At Kd values ≤0.1 ml/g an element can be considered as mobile as water, whereas an element with Kd value of ≥10 000 ml/g is practically immobile within the specified time boundaries (Sheppard et al. 2009b).

The most long-lived radionuclides have elongated interaction times with soils ranging from centuries to millennia. Thus, it has been proposed that desorption Kd values measured for extracted indigenous stable elements from field moist soils give a more truthful description of the slow retention processes at steady state conditions (Sheppard et al. 2009a, 2009b; Sheppard 2011). Slow processes are, for example, the fixation of caesium on the clay minerals’ frayed edge sites or incorporation of chlorine on organic matter (Bastviken et al. 2007; Hird et al. 1995). Desorption Kd values of indigenous elements are typically higher than sorption Kd values of radioactive tracers because of the longer interaction time with the solid phase and formation of less soluble forms (Sheppard et al. 2009b). On the other hand, the retention of short-lived radionuclides may be rather well described by the experimental sorption experiments (Sheppard et al. 2009b; Sheppard 2011).

Kd measured in the batch experiment is dependent on the experimental setup in the sense that Kd varies as the ambient temperature, rexod conditions, pH, organic matter content, particle size distribution, presence of competing ions, microbial activity etc. changes (e.g. Ashworh & Shaw 2006a; Begg et al. 2007; Cha et al. 2006; Gil-García et al. 2009; Lieser & Bauscher 1987; Sheppard et al. 1990; Sheppard & Thibault 1990; Zygmunt et al. 1998). All these factors are effective in the natural conditions as well.

Kd values are typically used in the geosphere transport modelling, when the radionuclide input rates to biosphere are calculated. Radionuclide release rates are, in turn, used to model the distribution of radionuclides within the biosphere. These simulations give activity concentrations in environmental media, which are the basis for evaluating the radiological impacts on humans and biota (Hjerpe & Broed 2010).

1.4 Overburden properties at Olkiluoto

Olkiluoto (Figure 1), a moderately sized island with current surface area of about 12 km2, situates on the eastern coast of the Baltic Sea. The bedrock comprises mostly of high-grade metamorphic supracrustal rocks (Kärki & Paulamäki 2006). Due to the continuous post-glacial crustal rebound (“land uplift”), currently at the rate of 6 mm/y (Eronen et al. 1995; Kahma et al. 2001), about a thousand years ago many of the initially small islands were interconnected into a bigger island and the Olkiluoto Island had begun to get its present shape (Mäkiaho 2005). The effects of crustal rebound are accentuated by a rather flat topography and paludification.

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The bedrock surface at Olkiluoto is quite variable, but the ground surface is rather smooth even in the places where the bedrock topography changes abruptly. As a result of the last glaciation, bedrock depressions are filled with a thick layer of overburden. The current mean elevation on Olkiluoto Island is +5 meters above the sea level, and the highest points are Liiklankallio (+18 m), Selkänummenharju (+13 m) and Ulkopäänniemi (+12 m) (Lahdenperä et al. 2005).

Figure 1. Olkiluoto Island. Topographic database, National Land Survey, permission 41/MYY/12 (Layout Posiva Oy/Jani Helin).

The most common soil types on Olkiluoto Island are fine-textured till (53 %), sandy till (39 %), gravelly till (4 %), peat (3.4 %) and outcrops (0.6 %) (Rautio et al. 2004). The thickness of the till cover is commonly 2-4 m and it is rich in clay fractions in some sites (Huhta 2005, 2007, 2009, 2010; Lintinen et al. 2003; Lintinen & Kahelin 2003; Lahdenperä et al. 2005, Lahdenperä 2009). Some Littorina and Anculys clay areas exist, in addition, to the areas of recent mud cover, especially on the northern and southern sides of the island (Haapanen et al. 2009). Soils at Olkiluoto are weakly developed (Tamminen et al. 2007) due to the short time span of land-uplift and development of 0-3000 years (Mäkiaho 2005). According to Tamminen et al. (2007), the typical soil types are undeveloped Arenosols and Regosols, thin or coarse-grained Leptosols and groundwater-type Gleysols. The predominant soil forming process in Finland is podsolization (Koljonen 1992), a slow process as the formation of a mature podzol profile can take up to 500-1500 years (Starr 1991). Typical podzol soil horizons are weakly developed at Olkiluoto (e.g. Lintinen et al. 2003; Lintinen & Kahelin 2003; Lahdenperä 2009; Lusa et al. 2009).

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The most common soil types on Olkiluoto Island are fine-textured till (53 %), sandy till (39 %), gravelly till (4 %), peat (3.4 %) and outcrops (0.6 %) (Rautio et al. 2004). The thickness of the till cover is commonly 2-4 m and it is rich in clay fractions in some sites (Huhta 2005, 2007, 2009, 2010; Lintinen et al. 2003; Lintinen & Kahelin 2003; Lahdenperä et al. 2005, Lahdenperä 2009). Some Littorina and Anculys clay areas exist, in addition, to the areas of recent mud cover, especially on the northern and southern sides of the island (Haapanen et al. 2009). Soils at Olkiluoto are weakly developed (Tamminen et al. 2007) due to the short time span of land-uplift and development of 0-3000 years (Mäkiaho 2005). According to Tamminen et al. (2007), the typical soil types are undeveloped Arenosols and Regosols, thin or coarse-grained Leptosols and groundwater-type Gleysols. The predominant soil forming process in Finland is podsolization (Koljonen 1992), a slow process as the formation of a mature podzol profile can take up to 500-1500 years (Starr 1991). Typical podzol soil horizons are weakly developed at Olkiluoto (e.g. Lintinen et al. 2003; Lintinen & Kahelin 2003; Lahdenperä 2009; Lusa et al. 2009).

Ca2+, Mg2+, K+, Na+, Al3+ and H+ are the main exchangeable cations in the Olkiluoto soils (Lahdenperä 2009; Lintinen et al. 2003; Lusa et al. 2009). Surface soils are naturally acidic and the pH increases with soil depth (Lahdenperä et al. 2005; Lahdenperä 2009; Lusa et al. 2009). The pH in the humus layer, surface soils and C layer (unaltered soil material) are typically 3.7-4.4, 4.0-8.0 and 6.5-8.0, respectively (Lahdenperä 2009; Lusa et al. 2009).

The typical minerals in order of abundance for gravel (Ø > 2 mm), sand (Ø 0.02-2 mm) and silt fractions (Ø 0.002-0.02 mm) are quartz, plagioclase, potassium feldspar, micas, chlorite and hornblende (Lintinen et al. 2003; Lusa et al. 2009). Illite, hornblende and chlorite are the typical minerals in the clay fraction (Ø < 0.002 mm) (Lintinen et al. 2003; Lusa et al. 2009). Specific surface area of soil particles varies between 2.3-15 m2/g and density between 2.8-3.0 g/cm3 (Lintinen & Kahelin 2003; Lintinen et al. 2003).

The results from the recharge computations carried out with the surface hydrology model indicate that the Olkiluoto bedrock groundwater system is transport-limited, whereas the overburden is supply-limited. The water supply from the overburden to the bedrock is higher than the bedrock system can transmit, which is due to the lower hydraulic conductivity of the bedrock compared with the overburden soils. The supply-limited overburden groundwater system can result in greater runoff and evapotranspiration with increased precipitation (Karvonen 2008). The average groundwater surface usually lies approximately in 2 meters below the ground surface (Lahdenperä et al. 2005; Lahdenperä 2009; Karvonen 2008, 2009).

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2 MATERIALS

2.1 Sampling

2.1.1 Sampling locations

Soil and soil solution samples were taken from the deep soil pits excavated at the Olkiluoto Island for the overall characterisation of the overburden at the site. Soil samples were taken from five soil pits, namely OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21. Soil water and soil samples of OL-KK14, OL-KK15 and OL-KK16 were taken in May 2008, OL-KK20 in 2009 and OL-KK21 in 2010. The sampling locations at the Olkiluoto Island are presented in Figure 2. In Table 3 are the coordinates of the sampling locations according to the Finnish National Coordinate system (KKJ).

Figure 2. The locations of the excavator pits OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21 at the Olkiluoto Island. Topographic database, National Land Survey, permission 41/MYY/12 (Layout Posiva Oy/Jani Helin).

Table 3. The coordinates of the excavator pits OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21 according to the Finnish National Coordinate System (KKJ).

Excavator pit Coordinates

N / latitude E / longitude OL-KK14 6791925 1525630 OL-KK15 6793022 1526567 OL-KK16 6792108 1528024 OL-KK20 6803452 3205461 OL-KK21 6803358 3205299

14

2.1.2 General description of soil sampling

The vertical soil sampling extended from the soil surface to the immediate vicinity of the bedrock surface. In the sampling of OL-KK20, the bedrock surface was not reached because groundwater filled the bottom parts of the pit. Humus and mineral soil samples were taken from the soil pits OL-KK14, OL-KK15, OL-KK16 and OL-KK21, whereas only mineral soil layers were sampled in OL-KK20. Humus was sampled at three different locations in the vicinity of OL-KK21. Aerobic mineral soil samples were taken from all the studied pits, whereas anaerobic samples were taken only from the soil pits OL-KK20 and OL-KK21. Table 4 presents a summary of the sample depths and the soil types of the studied excavator pits.

Table 4. A summary of the sampled soil layers, depths, soil types, soil types used in the Biosphere Description and Assessment and the sample types taken of the excavator pits OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21 at the Olkiluoto Island. Soil types of OL-KK20 and OL-KK21 are based on the results of the grain size distribution analysis ordered by Posiva Oy, whereas the soil types of the soil layers of OL-KK14, OL-KK15 and OL-KK16 are from Lusa et al. (2009).

Soil sample Soil layer Sampling depth (cm)

Soil type Soil type in biosphere assessment

Sample type

OL-KK14

humus 0-5 humus humus

aerobic, dried and fractioned

soil samples

MS1 5-20 sandy till medium mineral

soil


soil


soil MS4 105-240 clay clay

OL-KK15



soil samples

MS1 7-50 sand medium mineral

soil

MS2 50-80 fine grained

silty/clayish till fine mineral soil

MS3 80-110 fine sand fine mineral soil

MS4 110-160 clay clay

MS5 160-300 coarse grained

sandy till coarse mineral

soil

OL-KK16

humus 0-10 humus


soil samples

MS1 10-30 coarse sand coarse mineral

soil

MS2 30-50 sand medium mineral

soil

MS3 50-110 fine to coarse

sand fine to coarse mineral soil

MS4 110-300 sand/sandy till medium mineral

soil

15

Table 4. Continued.

Soil sample Soil layer Sampling depth (cm)

Soil type Soil type in biosphere assessment

Sample type

OL-KK20

not specified

70 sandy till medium mineral

soil

aerobic and anaerobic untreated

mineral soil samples

not specified

130 sandy till medium mineral

soil

MS2 205 sandy till medium mineral

soil


soil


soil

OL-KK21


aerobic untreated

humus samples;

aerobic and anaerobic untreated

mineral soil samples

MS1 40 fine sandy till fine mineral soil




soil


soil


soil

Aerobic soil samples

Aerobic soil samples were taken from all the studied soil pits with a shovel and conserved in plastic bags. OL-KK14, OL-KK15 and OL-KK16 samples were dried and sieved into grain sizes <0.063 mm, 0.063-0.125 mm, 0.125-0.25 mm, 0.25-0.50 mm, 0.50-1.0 mm, 1.0-2.0 mm and >2.0 mm at the Geological Survey of Finland. The grain size fraction <0.063 mm was further analysed at the laboratory of Labtium Oy using Sedigraph 5100-instrument (Lusa et al. 2009). Untreated samples of OL-KK20 and OL-KK21 were preserved in plastic bags in the room temperature at the laboratory of Radiochemistry, the University of Helsinki (Helsingin yliopiston Radiokemian laboratorio, HYRL).

Figures 3-7 presents the photographs and graphics of the soil profiles OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21. A photo of the humus sampling sites 1, 2 and 3 of OL-KK21 are presented in Figure 8.

16

Figure 3. Soil profile of OL-KK14 (upper picture; by Anne-Maj Lahdenperä/ Saanio & Riekkola Oy) and soil profile graphic (lower picture; by Teea Penttinen/ Pöyry Finland Oy).

17

Figure 4. Soil profile of OL-KK15 (by Anne-Maj Lahdenperä/ Saanio & Riekkola Oy) and soil profile graphic (lower picture; by Teea Penttinen/ Pöyry Finland Oy).

18

Figure 5. Soil profile of OL-KK16 (upper picture; by Anne-Maj Lahdenperä/ Saanio & Riekkola Oy) and soil profile graphic (lower picture; by Teea Penttinen/ Pöyry Finland Oy).

19

Figure 6. Soil profile of OL-KK20 (upper pictue; by Sinikka Virtanen/HYRL) and soil profile graphic (lower picture; by Teea Penttinen/ Pöyry Finland Oy). The black boxes in the soil profile graphic indicate the sampling depths of Posiva Oy.

20

Figure 7. Soil profile of OL-KK21 (upper picture; by Jani Helin/Posiva Oy) and soil profile graphic (lower picture; by Teea Penttinen/ Pöyry Finland Oy). The black boxes in the soil profile graphic indicate the sampling depths of Posiva Oy.

21

Figure 8. The sampling sites of the humus samples 1 (A), 2 (B) and 3 (C) from the vicinity of OL-KK21 (Photos by Sinikka Virtanen/HYRL). Humus samples 1 and 3 were taken from the edges of the soil pit OL-KK21, whereas humus sample 2 was from the vicinity of Olkiluoto main road.

Anaerobic soil samples

Anaerobic soil samples were taken from different mineral soil layers of OL-KK20 and OL-KK21 in 2009 and 2010. A special sampler designed for the determination of volatile compounds from soil samples, bought from Eijkelkamp (04.16 Soil coring kit for chemical soil research) (Figure 9), was used.

Samples were taken into stainless steel tubes (Figure 10), which were closed with PE-plugs at both ends. Plugs were taped with duct tape to ensure the tightness of the samples and to prevent the possible oxidation due to air penetration. Tubes were packed into plastic bags and taped securely. Samples were transported to the laboratory of Radiochemistry and transferred into nitrogen filled glovebox. The preservation and further handling of the samples was done in the glovebox. Samples were characterized and used as untreated (e.g. no drying or sieving).

22

Figure 9. Soil coring equipment used in the sampling of the anaerobic soil samples. The length of the ruler is 20 cm. (Photo by Mervi Söderlund/HYRL).

Figure 10. Handle, stainless steel sample tube and coring apparatus taken apart and assembled together (Photos by Mervi Söderlund/HYRL).

23

2.2 Characterisation of the soil samples

The chemical and geotechnical analyses of the soil samples included the determination of pH, organic matter and dry matter content, cation exchange capacity (CEC) and mineralogy. Specific surface area was determined for samples of OL-KK14, OL-KK20 and OL-KK21 by nitrogen gas adsorption. Grain size distribution was measured for samples of OL-KK14, OL-KK15 and OL-KK16. The detailed results from the analysis of OL-KK14, OL-KK15 and OL-KK16 samples are given in Lahdenperä (2009) and Lusa et al. (2009) excluding the specific surface area measurements. Only a rewiev of the main findings is given here. The analysed parameters, methods and standards used are summarized in Table 5.

Table 5. Analysed parameters from the soil samples, methods and used standards.

Parameter Unit Method Standard Number of

samples

pH Milli-Q-water extraction and

CaCl2 extraction

see Lusa et al. 2009, ISO 10390

14

Lost on ignition (LOI) = organic matter content

% Gravimetrically in

550 °C CEN 15407 14

Dry matter content % Gravimetrically in

105 °C ISO 11465 14

Cation exchange capacity (CEC)

mmol/kg Ammonium

acetate extraction in pH 4.5

not specified 14

Mineralogy X-ray diffraction

analyser ISO 3310/1 11

Specific surface area

m2/g

Micrometrics Flowsorb

instrument – N2 adsorption

not specified 35

2.2.1 Chemical characteristics

Soil pH was measured both in Milli-Q-water and in dilute CaCl2 solution with soil to liquid ratio 1:1. Samples were stirred for 30 seconds and left to stand for three minutes (e.g. Lusa et al. 2009). This was repeated five times, after which the samples were let to settle for ten minutes before pH measurement. A pH test series with the anaerobic soil sample of OL-KK20 sample depth 0.7 m was conducted both in aerobic and anaerobic conditions in order to found out if a change in the ambient conditions has a significant affect on pH. As the pH values showed no dependence upon the measuring conditions (aerobic/anaerobic), it was decided that the pH values of the anaerobic samples are measured in the aerobic conditions to simplify the sample treatment and pH measuring. The pH values of the aerobic and anaerobic soil samples of OL-KK20 and OL-KK21 are given in Appendix 1.

The pH measured in water is typically 0.5-1 pH units higher than in dilute CaCl2 or other electrolyte solution (Westman 1991). The use of an electrolyte solution induces lower pH values because Ca2+ and other cations are able to replace part of the acidic

24

cations (e.g. H+, Al3+) present on the soil surfaces’ ion exchange sites into the liquid. When pure water is used in the pH measurement, the salt concentration of the liquid is dependent on the readily or easily soluble salts present on the soil surfaces. The majority of the ions affecting the pH, consisting mainly of H+, are not transferred to the solution from the ion exchange sites, and thus higher pH values are measured.

The variation in the humus and clay mineral content has the greatest impact on the soil pH (Birkeland 1984; Kähkönen 1996; Räisänen 1989). The pH typically decreases with depth as the amount of organic matter and organic acids produced in its decomposition decreases.

Figure 11 presents the soil pH for the samples of OL-KK14, OL-KK15 and OL-KK16 measured in Milli-Q-water and dilute CaCl2 solution. The pH increases with depth, having the smallest value in the humus layer, typically 3.5-5.0 in dilute CaCl2 and 4.3-5.3 in water. In mineral soil layers the pH increases from 4.8 and 5.3 in the uppermost layers to 7.8 and 8.8 in the bottom soil layers. The pH remains rather constant throughout the soil pit OL-KK20 (Figure 12). The total increase in pH is only 0.25 units between the top and bottom soil layers. It is possible that a systematic error occurred when pH was measured, but a pH measurement from new samples verified the observation. The pH in the humus layer of OL-KK21 is 5.6 and 6.3 when measured in dilute CaCl2 and Milli-Q-water, respectively. These values are about two pH units lower compared with the pH values measured for the humus layers of OL-KK14, OL-KK15 and OL-KK16. In mineral soil layers of OL-KK21 pH increases with depth from 6.6 and 7.5 to 7.8 and 8.6 as measured in water or CaCl2, respectively.

3,5 4,0 4,5 5,0 5,5 6,0 6,5 7,0 7,5 8,0 8,5 9,0

300

250

200

150

100

50

0

So

il d

epth

(cm

)

Soil pH

OL-KK14 water OL-KK14 CaCl

2


2


2

Figure 11. The pH of the soil samples of OL-KK14, OL-KK15 and OL-KK16 measured in Milli-Q-water (Lusa et al. 2009) and dilute CaCl2 solution (Lahdenperä 2009).

25

5,5 6,0 6,5 7,0 7,5 8,0 8,5400

350

300

250

200

150

100

50

0

So

il d

epth

(cm

)

Soil pH


2


2

Figure 12. The pH of the aerobic soil samples of OL-KK20 and OL-KK21 as measured in Milli-Q-water and dilute CaCl2 solution. Figure 13 presents the pH values for the anaerobic OL-KK20 and OL-KK21 soil samples. The pH increases with depth, and are approximately 0.5 pH units higher than for the aerobic samples.

6,5 7,0 7,5 8,0 8,5 9,0 9,5400

350

300

250

200

150

100

50

0

So

il d

ep

th (

cm

)

Soil pH


2


2

Figure 13. The pH of the anaerobic soil samples of the pits OL-KK20 and OL-KK21 as measured in Milli-Q-water and dilute CaCl2 solution. Figure 14 presents the potential cation exchange capacity (CEC) for the soil samples of OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21. Potential CEC indicates the cation exchange capacity of the soil in a certain, specified pH value, such as 4.5 or 7, determined by buffered salt solutions (Brady & Weil 2002). Effective cation exchange capacity is determined by neutral salt solutions, such as NH4Cl, in the soil’s own pH. The potential CECs for OL-KK14, OL-KK15 and OL-KK16 samples were

26

calculated as a sum of Na, K, Mg and Ca ions extracted by 1 M NH4Ac solution in pH 4.5 (Lahdenperä 2009). For OL-KK20 and OL-KK21 soil samples same method was used with solid to liquid ratio 1:5. Stable Na, K, Mg and Ca concentrations were measured with ICP-MS (inductively coupled plasma mass spectrometry). Al3+ and Fe2+ were omitted in the calculation of CEC because their concentration was below the limit of detection of ICP-MS (Table 6). Table 6. The detection limits for Na, K, Mg, Ca, Al and Fe on ICP-MS.

Element Detection limit

(µg/l) Element

Detection limit (µg/l)

Sodium (Na) 19.9 Calcium (Ca) 59.6

Potassium (K) 115.9 Aluminium (Al) 3.7

Magnesium (Mg) 0.7 Iron (Fe) 3.7

The highest CECs are typically found in the humus layers where the high concentrations of ion exchange groups, especially –COOH groups, are found in the organic matter creating CEC capacities as high as 1500-5000 mmol H+/kg dry weight (Birkeland 1984). Typically in mineral soil CEC increases with increasing clay mineral and clay fraction content. For the soil pits OL-KK14 and OL-KK15 the CEC of the uppermost mineral soil layer (MS1) was lower than the CEC of the underlying mineral soil layer (MS2) due to the higher clay fraction in MS2. In the bottom soil layers the CEC decreased with depth. In the pits OL-KK20 and OL-KK21 CECs varied considerable between the sampling depths, and there was no clear dependence between CEC and the soil depth. In OL-KK20, the CEC for the uppermost sampling depth 0.7 m (40 mmol/kg) was the lowest among the studied sampling depths, and the highest value was calculated for the underlying sampling depth 1.30 m (68 mmol/kg). In OL-KK21, the CEC was the highest at 1.75 m depth (124 mmol/kg) and the lowest at 1.1 m depth (32 mmol/kg). The CECs of the humus samples taken from the vicinity of OL-KK21 ranged from 130 mmol/kg to 282 mmol/kg.

Calcium was the main exchangeable cation on soil surfaces followed by magnesium and sodium. Potassium concentration in the extraction solution was the smallest among the measured cations indicating relatively low concentration on soil sorption sites. The reported main exchangeable cations in the Olkiluoto soils are Ca2+, Mg2+, K+, Na+, Al3+ and H+ (Lahdenperä 2009; Lintinen et al. 2003; Lusa et al. 2009).

Table 7 presents the cation exchange capacities for the samples taken of OL-KK20 and OL-KK21 calculated as a sum of Na, K, Mg and Ca ions extracted by 1 M NH4Cl solution buffered in pH 4.5.

27

0 100 200 300 400400

350

300

250

200

150

100

50

0

So

il d

ep

th (

cm)

CEC (mmol/kg)

OL-KK14 OL-KK15 OL-KK16 OL-KK20 OL-KK21

Figure 14. The potential cation exchange capacities (CECs) for the soil samples of OL-KK14, OL-KK15 and OL-KK16 (Lahdenperä 2009), OL-KK20 and OL-KK21 calculated as a sum of Na, K, Mg and Ca ion concentrations in 1 M NH4Ac extraction solution. Table 7. The potential cation exchange capacity (CEC, mmol/kg of the dry weight) in the soil samples of OL-KK20 and OL-KK21 determined by NH4Ac buffered in pH 4.5. CEC was calculated as a sum of Na, K, Mg and Ca ions.

OL-KK20 OL-KK21 Soil depth (cm) CEC (mmol/kg)

(standard deviation) Soil depth (cm) CEC (mmol/kg)

(standard deviation) 70 40 (1) 12 (humus 1) 282 (9) 130 68 (-) 12 (humus 2) 130 (3) 205 51 (2) 12 (humus 3) 207 (2) 300 60 (4) 40 43 (1) 340 66 (2) 110 32 (5)

175 124 (69) 240 43 (3) 295 43 (4) 360 56 (1)

- standard deviation was not calculated because the CEC value is based on the Na, K, Mg and Ca concentrations determined from one soil sample

2.2.2 Geotechnical characteristics

Organic and dry matter content was determined from all the soil samples. Specific surface area was measured for the samples of OL-KK14, OL-KK20 and OL-KK21. The results of the organic and dry matter contents of OL-KK14, OL-KK15 and OL-KK16 are presented in Lusa et al. (2009). The results of these soil pits are only reviewed here.

28

Grain size distribution was determined from the soil samples of OL-KK20 and OL-KK21 in the Laboratory of ALS, Luleå, Sweden, but the sampling depths were somewhat different than used in the sorption studies. Thus, the analysis results form the grain size distribution are not presented in this report.

The dry and organic matter content was determined by weighting three to five grams of humus or soil sample into a beaker glass. The beakers were placed to a drying oven at the temperature of 105 °C for 24 hours. Samples were cooled to room temperature and weighted to define the amount of water. The amount of organic matter was determined as a loss of ignition (LOI) by incinerating the samples in an oven at 550 °C degrees from five to six hours. After incineration the samples were cooled to room temperature and weighted. The dry and organic matter content of the samples were calculated using Equation 2 and 3.

% 100% [2]

Equation 2. The calculation of the dry matter content. m20 = sample mass in room temperature and m105 = sample mass after annealing at 105°C.

% 100% [3]

Equation 3. The calculation of the organic matter content. morg = mass of the organic matter in the sample and m550 = the sample mass after annealing at 550 °C.

The organic matter (OM) content decreases rapidly below the humus layer in the mineral soil layers underneath (Figure 15, Appendix 2). In the humus layer of OL-KK14, OL-KK15 and OL-KK16 the organic matter content was 59 %, 63 % and 54 %, respectively, whereas in OL-KK21 the corresponding values were 11 %, 14 % and 10 % for humus samples 1, 2 and 3. The humus layer of OL-KK21 was very thin and had considerable amounts of mineral matter mixed up, descriptive of an uppermost mineral soil layer enriched with organic matter (Ah). This might explain the low organic matter content of the OL-KK21 humus samples. Instead of humus, the layer might be representative of an uppermost mineral soil layer enriched with organic matter (Ah). In the mineral soil layers organic matter content was <1 %.

The dry matter content of the mineral soil samples of OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21 is presented in Figure 16 and for the aerobic and anaerobic soil samples of OL-KK20 and OL-KK21 in Appendix 2. The dry matter content for the humus samples of OL-KK14, OL-KK15 and OL-KK16 was 48 %, 43 % and 27 %, respectively. For the OL-KK21 humus samples 1, 2 and 3 the dry matter content was 29 %, 34 % and 39 %. Figure 17 presents the organic matter (OM) content for the anaerobic mineral soil samples of OL-KK20 and OL-KK21. The OM contents were similar to the aerobic samples, but the dry matter content was typically 1.5-3.0 % higher for the anaerobic samples.

29

0,0 0,5 1,0 1,5 2,0 2,5 3,0400

350

300

250

200

150

100

50

0

So

il d

epth

(c

m)

Organic matter content (%)


Figure 15. The organic matter content for the aerobic mineral soil samples of the excavator pits OL-KK14, OL-KK15 and OL-KK16 (Lusa et al. 2009), OL-KK20 and OL-KK21.

84 86 88 90 92 94 96400

350

300

250

200

150

100

50

0

So

il d

ep

th (

cm

)

Dry matter content (%)


Figure 16. The dry matter content for the aerobic soil samples of the excavator pits OL-KK14, OL-KK15 and OL-KK16 (Lusa et al. 2009), and OL-KK20 and OL-KK21.

30

0,4 0,6 0,8 1,0 1,2 1,4400

350

300

250

200

150

100

50

0

So

il d

epth

(m

)

Organic matter content (%)

OL-KK20 OL-KK21

Figure 17. The organic matter content for the anaerobic mineral soil samples of the excavator pits OL-KK20 and OL-KK21.

Specific surface areas were determined from the sieved mineral soil samples of OL-KK14, and bulk soil samples of OL-KK20 and OL-KK21. Samples were sent to the Tampere University of Technology, where the measurements were done using Micromeritics Flowsorb instrument. Soil samples were dried to constant weight, cooled to room temperature and specific surface area was measured based on nitrogen adsorption (BET/N2). Results for specific surface area measurements are presented in Figures 18 and 19 and Appendix 3.

For OL-KK14, the specific surface area was the lowest in the topmost mineral soil layer (MS1 at the depth of 20 cm) ranging from 2.1 to 0.6 m2/g, and the highest, 9.0 and 4.6 m2/g, in the mineral soil layer underneath (MS2 at the depth of 60 cm). Also, the CEC is higher for MS2 layer than for MS1 (Figure 14) (Lusa et al. 2009). Grain size did not have a great impact on the specific surface area except for the smallest grain size fraction, <0.063 mm, were the highest values were seen in every soil layer among the studied grain size fractions. These results are in a good agreement with the previous measurements of the specific surface areas of soil particles, even though the smallest values are below earlier results from Olkiluoto (2.3-15 m2/g) (Lintinen & Kahelin 2003; Lintinen et al. 2003).

The specific surface area of the OL-KK20 and OL-KK21 samples decreases with sample depth (Figure 19) from 7 m2/g at 0.7 m depth to 1 m2/g at 3.4 m depth and from 4 m2/g at 0.4 m depth to 1 m2/g at 3.6 m depth. This may be attributed to the formation of large specific surface area secondary (clay) minerals and weakly crystalline aluminium and iron oxides as the weathering products of the soil primary minerals (Birkeland 1984; Brady & Weil 2002). The weathering processes mainly take place in the uppermost mineral soil layers underneath the humus layer, where hydration, hydrolysis, acid reactions and oxidation-reduction reactions act simultaneously changing the soil geological material. Also, the soil texture has an impact on the specific surface area due to the fact that the particle surface area increases with decreasing particle size (Brady & Weil 2002).

31

<0.063 0.063-0.125 0.125-0.25 0.25-0.5 0.5-1.0 1.0-2.00

1

2

3

4

5

6

7

8

9

Sp

ecifi

c su

rfa

ce a

rea

(m2 /g

)

Grain size (mm)

soil depth: 5-20 cm 20-60 cm 60-105 cm 105-240 cm

Figure 18. The specific surface areas of the MS1, MS2, MS3 and MS4 samples of OL-KK14 as a function of the grain size. The specific surface area was measured by nitrogen adsorption.

0 1 2 3 4 5 6 7400

350

300

250

200

150

100

50

0

So

il d

epth

(cm

)

Specific surface area (m2/g)

OL-KK20 OL-KK21

Figure 19. The specific surface area of the soil samples of OL-KK20 and OL-KK21 measured by nitrogen adsorption.

When the distribution coefficients of radionuclides are determined for bedrock samples, the samples are typically ground and Kd values are determined for a specific grain size fraction. In this case the correct way to present the Kd values is to normalize to the specific surface area, because mass based Kd value typically increase with decreasing grain size fraction. In the case of soil samples, the samples are dry sieved into different grain size fractions or conserved untreated. Thus, no grinding is involved in the sample treatment and mass based Kd values determined in the sorption tests give an accurate picture of the distribution of the radionuclide between the liquid and solid phases. In the

32

following Chapters, the presented Kd values are mass based values and normalisation to the specific surface area was not carried out.

2.2.3 Mineralogy

Mineralogical analyses of the soil samples of OL-KK20 and OL-KK21 were done in the Geological Survey of Finland. The samples were dried and grain size fractions <0.01 mm and 1.0-2.0 mm were separated from the bulk sample by sieving. Mineral composition was examined by X-ray diffraction (XRD) using Philips X’Pert MPD instrument. XRD spectra were recorded in angle interval 2-70°2θ with step size 0.02°. Time per step was 1 s. Based on XRD-analysis, IR-spectra was measured for two samples in accordance to identify kaolinite, chlorite and other clay minerals present in samples. Swelling clay minerals were determined by heating four samples in 550 °C for one hour, after which XRD spectra was rerecorded in angle interval 2-35°.

Samples were mainly composed of quartz, potassium feldspar, plagioclase and kaolinite. Quartz was the main mineral, followed by plagioclase. Some samples contained minor amounts of micas, amphibole, hematite and chlorite. No swelling clay minerals were found in the samples. The mineralogy was similar to <0.01 mm and 1.0-2.0 mm samples differing only in the mineral proportions. The abundance of clay minerals was higher for the smaller grain size fraction. Difference with regard to the mineralogical composition of the soil samples of OL-KK14, OL-KK15 and OL-KK16, amphibole was not one of the main minerals and kaolinite content was much higher. Figures 20-23 present the main mineral compositions of grain size fractions <0.01 mm and 1.0-2.0 mm samples separated from bulk soil samples of soil pits OL-KK20 and OL-KK21. Mineralogy of OL-KK20 and OL-KK21 is given in Appendix 4.

350

300

250

200

150

100

50

0 10 20 30 40 50 60 70 80 90 100

Mineral content (%)

So

il d

epth

(cm

)

Other minerals

Kaolinite Plagioclase Potassium

feldspar Quartz

Figure 20. Quartz, potassium feldspar, plagioclase, kaolinite and other mineral (micas, amphibole, hematite and chlorite) content for mineral soil samples in the grain size fraction <0.01 mm of OL-KK20.

33

350

300

250

200

150

100

50

0 10 20 30 40 50 60 70 80 90 100

Mineral content (%)

So

il d

epth

(c

m)

Other minerals


feldspar Quartz

Figure 21. Quartz, potassium feldspar, plagioclase, kaolinite and other mineral (micas, amphibole, hematite and chlorite) content for mineral soil samples in the grain size fraction 1.0-2.0 mm of OL-KK20.

350

300

250

200

150

100

50

0 10 20 30 40 50 60 70 80 90 100

Mineral content (%)

So

il d

ep

th (

cm

)

Other minerals


feldspar Quartz

Figure 22. Quartz, potassium feldspar, plagioclase, kaolinite and other mineral (micas, amphibole, hematite and chlorite) content for mineral soil samples in the grain size fraction <0.01 mm of OL-KK21.

34

350

300

250

200

150

100

50

0 10 20 30 40 50 60 70 80 90 100

Mineral content (%)

So

il d

ep

th (

cm

)

Other minerals


feldspar Quartz

Figure 23. Quartz, potassium feldspar, plagioclase, kaolinite and other mineral (micas, amphibole, hematite and chlorite) content for mineral soil samples in the grain size fraction 1.0-2.0 mm of OL-KK21.

2.3 Water samples

The sorption of radionuclides is usually examined in two-component electrolyte solutions, such as NaCl or CaCl2. For the direct use of the determined Kd values, it is recommended to use a liquid phase representing the site-specific composition of soil solution or groundwater to simulate the circumstances on the field conditions (Camps et al. 2003; Sauvé et al. 2000; Yasuda et al. 1995). To create a suitable composition of the liquid phase to be used in the sorption tests, it is important to gain knowledge on the soil solution composition on Olkiluoto Island. Therefore soil solution samples for the main cation and anion analysis were taken from humus and mineral soil layers of OL-KK14, OL-KK15 and OL-KK16 in 2008. The pH, DOC, anion and cation analysis results have been reported previously for two forest intensive monitoring plots of OL-FIP4 (Scots pine forest) and OL-FIP10 (Norway spruce forest) (Haapanen 2006, 2007, 2008, 2009). Soil solution analysis results formed the basis for the composition of the soil solution simulant.

2.3.1 Soil solution sampling and chemical characteristics

Soil solution samples were taken from the soil layers of OL-KK14, OL-KK15 and OL-KK16. Sampling for soil solution separation was done by inserting a weighted plastic tube with an inner diameter of 4.6 cm into the wall of the excavator pit at the desired depths. Soil filled tubes were frozen and sent to the Finnish Forest Research Institute (Metla, Rovaniemi) for soil solution separation with high speed centrifuge. The pH and DOC concentration was measured at Metla, whereas the cation and anion analyses were performed at the Department of Geography (the University of Helsinki). The detailed

35

description of the soil solution sampling and analysis results is presented in Lusa et al. (2009).

Soil solution sampling of the forest intensive monitoring plots OL-FIP4 and OL-FIP10 was started in 2004 and 2007 (Aro et al. 2010). Samples were taken from different depths down the soil profile to approximately 30 cm with suction cup lysimeters (tension lysimetre) installed in the mineral soil layers, and plate lysimeters (zero-tension lysimetre) typically situated immediately underneath the humus layer. The pretreatment of the water samples was done in the field and the chemical analyses were done at Metla, Rovaniemi.

Figure 24 presents the compilation of the average soil and soil solution pH values as a function of soil depth. Soil pH values were calculated based on the pH values determined for the aerobic soil samples of OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21 (original pictures Figures 11 and 12), whereas soil solution pH values were based on the pH values for OL-KK14, OL-KK15, OL-KK16, OL-FIP4 and OL-FIP10 (Figure 25).

4,0 4,5 5,0 5,5 6,0 6,5 7,0 7,5 8,0 8,5400

350

300

250

200

150

100

50

0

So

il d

epth

(cm

)

pH

soil solution soil; water soil; CaCl

2

Figure 24. The average pH values of the soil and soil solution samples as a function of soil depth. Soil pH was determined in Milli-Q-water and in dilute CaCl2 solution.

Figure 25 and 26 presents the pH and DOC concentration of soil water and soil solution samples. The solution pH increases with depth from approximately 4 in the topmost soil layers (humus) to 7.5 in the deepest mineral soil layers. DOC concentrations reach the highest values, 120-160 mg/l in the uppermost soil layers (humus) and decrease rapidly with depth to circa 5 mg/l.

36

3,5 4,0 4,5 5,0 5,5 6,0 6,5 7,0 7,5 8,0

300

250

200

150

100

50

0

So

il d

epth

(cm

)

pH

OL-KK14 OL-KK15 OL-KK16 FIP4 FIP10

Figure 25. The pH of the soil water samples (OL-KK14, OL-KK15 and OL-KK16) and the soil solution samples (Haapanen 2006, 2007, 2008, 2009).

0 20 40 60 80 100 120 140 160 180

300

250

200

150

100

50

0

So

il d

epth

(cm

)

DOC (mg/l)

OL-KK14 OL-KK15 OL-KK16 FIP4 FIP10

Figure 26. The DOC concentration of the soil water samples (OL-KK14, OL-KK15 and OL-KK16) and the soil solution samples (Haapanen 2006, 2007, 2008, 2009).

The cation (Na, K, Ca and Mg) concentrations of OL-KK14, OL-KK15, OL-KK16, OL-FIP4 and OL-FIP10 soil solution samples are presented in Figure 27. The main cation concentrations varied between the samples and depths. In the OL-FIP10, the maximum sodium concentration was found in the 30 cm depth, 17.5 mg/l, and the minimum concentration in MS2 of OL-KK14, 1.5 mg/l. Potassium concentrations seem to behave somewhat similarly, as the highest values were in the humus layers and lowest in the

37

uppermost mineral soil layers. Calcium concentrations varied greatly between the study sites. The highest Ca concentration was in the humus layer of OL-KK16 (25 mg/l), and the lowest in MS2 mineral soil layer of OL-KK14 (1.7 mg/l). The maximum and minimum concentration of magnesium was in MS2 layer of OL-KK15 (11.5 mg/l) and OL-KK14 (0.6 mg/l), respectively (Haapanen 2006, 2007, 2008, 2009; Lusa et al. 2009). Cation concentrations from the soil solution were measured with ICP-MS.

The main anion (Cl, NO3, F and SO4) concentrations of OL-KK14, OL-KK15, OL-KK16, OL-FIP4 and OL-FIP10 soil solution samples are presented in Figure 28. Chloride and nitrate concentrations depended on the sampling depth; decreasing with increasing soil depth. The highest concentrations for Cl and NO3 were in the humus layer, 58 mg/l in OL-KK15 and 7.4 mg/l in OL-KK16, respectively. The minimum chloride concentration 2 mg/l was in the uppermost sampling depth of OL-FIP4 and OL-FIP10, and in MS2 layer of OL-KK14. Nitrate concentration was typically below 1 mg/l in all the studied sampling depths. The average fluoride concentration in the soil water samples of OL-KK14, OL-KK15 and OL-KK16 was 2.92 mg/l. The highest concentration was in the humus sample of OL-KK15 (8.3 mg/l), and the lowest in the OL-KK16 (0.8 mg/l). Sulphate concentrations varied greatly, the lowest concentration was only 0.6 mg/l in the uppermost sample of OL-FIP4, and the highest 204 mg/l in MS2 layer of OL-KK15. Anion concentrations from the soil solution were measured with ion chromatography (IC).

38

Fig

ure

27.

Con

cent

rati

ons

of th

e m

ain

cati

ons

sodi

um (

Na)

, pot

assi

um (

K),

mag

nesi

um (

Mg)

and

cal

cium

(C

a) in

the

soil

sol

utio

n sa

mpl

es

of O

L-K

K14

, O

L-K

K15

and

OL

-KK

16 (

Lus

a et

al.

2009

) an

d fo

rest

int

ensi

ve m

onit

orin

g pl

ots

OL

-FIP

4 an

d O

L-F

IP10

(H

aapa

nen

2006

, 20

07, 2

008,

200

9).

38

39

Fig

ure

28.

Con

cent

rati

ons

of t

he m

ain

anio

ns c

hlor

ide

(Cl)

, flu

orid

e (F

), n

itra

te (

NO

3) a

nd s

ulph

ate

(SO

4) i

n th

e so

il s

olut

ion

sam

ples

of

OL

-KK

14,

OL

-KK

15 a

nd O

L-K

K16

(L

usa

et a

l. 20

09)

and

fore

st i

nten

sive

mon

itor

ing

plot

s O

L-F

IP4

and

OL

-FIP

10 (

Haa

pane

n 20

06,

2007

, 200

8, 2

009)

.

39

40

2.3.2 Soil solution simulant

The liquid phase used in the sorption experiments was synthetic soil solution simulant. The soil solution simulant was composed after the cation and anion analysis of the soil solution samples taken from the humus and mineral soil layers of OL-KK14, OL-KK15 and OL-KK16 and from the forest intensive monitoring plots OL-FIP4 and OL-FIP10 (Haapanen 2006, 2007, 2008, 2009; Lusa et al. 2009).

Cations in the soil solution simulant were sodium (Na+), potassium (K+), calcium (Ca2+) and magnesium (Mg2+). The cation concentration was equilibrated with counter anions chloride (Cl-), sulphate (SO4

2-) and silicate (SiO32-). Fluoride (F-), nitrate (NO3

-) and phosphate (PO4

3-) were omitted from the soil solution simulant due to their small concentrations in the soil solution samples compared with chloride and sulphate (Haapanen 2006, 2007, 2008, 2009; Lusa et al. 2009).

The soil solution simulant was prepared from J.T. Baker and Merck p.a. and Suprapur grade chemicals using ultrapure 18 MΩ Milli-Q water. Table 8 presents the reagents used in the preparation of the soil solution simulant and calculated concentrations of the cations and anions.

Table 8. The reagents used in the preparation of the soil solution simulant and the calculated concentrations of the ions.

Ion Reagent Calculated concentration

(mg/l)

Calculated concentration

(mmol/l) Na+ Na2SiO3 x 9 H2O 4.0 0.17 K+ KCl 2.7 0.07

Mg2+ MgSO4 x 7 H2O 2.7 0.11 Ca2+ CaCl2 x 2 H2O 6.5 0.16

SiO32- Na2SiO3 x 9 H2O 6.6 0.08

Cl- KCl and CaCl2 x 2 H2O 16.5 0.36 SO4

2- MgSO4 x 7 H2O 10.7 0.11

2.3.3 Electrolyte and buffer solutions

CsCl, NaCl, KCl, NH4Cl and CaCl2 electrolyte solutions were used in the sorption experiments of caesium when the effect of competing cations was investigated. The cation concentrations of CsCl solutions used were 0.01, 0.1, 1 and 10 mM, whereas the cation concentrations of other electrolyte solutions were 0.1, 1.0, 10, 100 and 1000 mM.

Electrolyte solutions were prepared from Merck and Riedel-de Häen p.a. and Suprapur grade reagents in ultrapure 18 MΩ Milli-Q water.

CHES buffer solution (2-(cyclohexylamino)ethanesulfonic acid) was used in the sorption experiments of Nb-95 in the anaerobic conditions to buffer the solution pH to 9-10. CHES concentration in the final solution was 1.0x10-3 M. CHES was purchased from Fluka (ultra purity).

41

2.3.4 Radioactive tracers

Radioactive tracers used in the sorption studies were commercially available Cs-134, Cl-36, I-125, Nb-95, Se-75 and Tc-99.

The initial chemical form of Cs-134 was caesium chloride (CsCl) in 0.1 M HCl. The tracer was bought from Eckert & Ziegler. Solution for the sorption experiments was prepared from the stock tracer solution by diluting with 2x10-4 M HCl and by later on with ultrapure 18 MΩ Milli-Q-water and used without pH adjustment.

The initial chemical form of Cl-36 was sodium chloride (NaCl) in aqueous solution. The tracer was bought from Amersham Biosciences. Solution for the sorption experiments was prepared from the tracer stock solution by diluting with ultrapure 18 MΩ Milli-Q-water and used without pH adjustment.

The initial chemical form of I-125 was sodium iodide (NaI) in phosphate buffer solution. The tracer was bought from MAP Medical Technologies. Solution for the sorption experiments was prepared from the tracer stock solution by diluting with ultrapure 18 MΩ Milli-Q-water and used without pH adjustment.

The initial chemical form of Nb-95 was niobium chloride (NbCl5) in 6.0 M HCl. The tracer was bought from Perkin Elmer as a mixture of Nb-95 and its mother nuclide Zr-95. Nb-95 and Zr-95 have similar x-ray and γ-energies, for which reason Nb-95 had to be separated from its mother nuclide Zr-95. The separation was done by the following scheme (Weiss et al. 1987):

A small volume (100-300 µl) of Nb-95/Zr-95 solution was drawn via 0.8 injection needle into 1 ml syringe, and the syringe was emptied into a 25 ml beaker glass. Tracer solution was evaporated to near dryness and dissolved to 0.5-1.0 ml of 10 M HNO3. This procedure was repeated three times.

The separation column was made of 5 ml syringe by inserting seven layers of Whatman GF6 glass fibre filters on to the other. A plastic tube was jointed to a plastic faucet and the faucet was attached to the syringe. The column was pre-conditioned by withdrawing 2x1 ml of 10 M HNO3 slowly through the filters into a 5 ml syringe connected to the plastic tube.

Tracer solution, 1 ml of HNO3 containing Nb-95 and Zr-95, was transferred into the column. The solution was drawn into a clean 5 ml syringe. The beaker glass was rinsed with 1 ml of 10 M HNO3 and the acid was transferred to the column and drawn slowly to the syringe. The solution was moved to a weighted 20 ml plastic liquid scintillation bottle. This solution was so called ‘absorption solution’ and contained the majority of the mother nuclide Zr-95.

The column was washed with 2x1 ml of HNO3 to ensure that all of the non-sorbed Zr-95 had come through. After this stage, the plastic tube was changed to a clean one and Nb-95 was eluted from the column with 2x1 ml of 0.5 M HNO3 + 0.3 M HF solution. The elution solution containing Nb-95 was withdrawn into a 10 ml syringe and transferred to a weighted 20 ml plastic liquid scintillation bottle or to a clean 20 ml beaker glass for the change of the tracer matrix from

42

0.5 M HNO3 + 0.3 M HF to 0.1 M HCl. The tracer matrix was changed by evaporating the solution until dry and dissolving the precipitate into 1 ml of 0.1 M HCl. This procedure was repeated three times, after which the solution was transferred to a weighted 20 ml (HD PE) plastic liquid scintillation bottle. The bottle was weighted again to find out the volume of Nb-95 tracer solution. The purity of the Nb-95 tracer solution was ensured with gammaspectrometry.

The initial chemical form of Se-75 was selenious acid (H2SeO3) in 0.1 M HCl. The tracer was bought from Eckert & Ziegler. Solution for the sorption experiments was prepared from the tracer stock solution by diluting with 0.1 M HCl and used without pH adjustment.

The initial chemical form of Tc-99 was 0.01 M ammonium pertechnetate (NH4TcO4). The tracer was bought from Amersham International. Solution for the sorption experiments was prepared from the tracer stock solution by diluting with 0.01 M NH4OH and used without pH adjustment.

43

3 METHODS

3.1 Analysis of waters

3.1.1 Cation and anion analysis

Cation and anion analysis were conducted to the soil solution samples separated at Metla and soil solution simulant test samples at the beginning of the sorption experiments. Analysis of the soil solution samples was conducted in order to find out the main cations and anions present in the samples. Based on the results, the composition of the soil solution simulant was created. The required contact time for the liquid phase to reach a steady-state with the solid state for the main ion composition and pH was characterised by measuring the changes as a function of time.

From the soil solution samples the main cations Na, K, Ca, Mg and the trace cations Al, As, Cd, Cr, Cs, Cu, Fe, Mn, Ni, S, Pb, Zn and U were analyzed with Agilent 7500ce/cx ICP-MS equipped with octopole reaction chamber. Anions F, Cl, NO3, PO4 and SO4 were measured with IC (Metrohm MIC-12 ion chromatogram and ASupp5/100 column). Elemental concentration standards and background (Milli-Q-water) samples were measured together with the solution samples.

For the soil solution simulant test samples, only Na, K, Ca and Mg were analysed from the cations, and Cl, SiO3 and SO4 were measured from the anions.

3.1.2 pH measurement

Equilibrium pH was measured from every sorption sample and soil solution simulant test samples. Before measurements, pH electrode was calibrated with pH buffer standards (Orion) 4.01, 7.00 and 10.01 and the slope-value was recorded. After calibration pH electrode was inserted into the solution and the pH value was let to stabilise. For the aerobic samples, pH was mainly measured from filtered liquid samples and from some unfiltered, centrifuged samples. In the anaerobic conditions, pH was solely measured from the unfiltered liquid samples. Gel-filled low maintenance electrode Orion 9107BNMD was used in the aerobic conditions. In the nitrogen atmosphere (glovebox), glass electrode Orion 9104APWP was applied.

The pH of the soil solution simulant test samples was determined after the soil and soil solution simulant reached the steady-state.

3.1.3 Redox potential measurement

Redox potential was measured from the anaerobic sorption samples in the nitrogen atmosphere of the glovebox. Eh-electrode (Orion 9778BNWP) was calibrated with ORP-solution (Orion) to correlate the measured values to the corresponding Eh mV (millivolt) values given by standard/normal hydrogen electrode (SHE or NHE). After calibration, Eh-electrode was inserted into the soil/water mixture and the redox potential was let to stabilize. During the Eh-measurement, samples were gently shaken. Stabilisation was usually achieved in 30-45 minutes.

44

3.1.4 Tracer activity measurement

The tracer activity in the liquid phase was determined by gammaspectrometry or liquid scintillation counting depending on the radiation properties of the studied nuclide.

The activity of the β- -emitters Cl-36 and Tc-99 was determined by liquid scintillation counting with 1219 Rackbeta (LKB Wallac) or TriCarb 2910TR (PerkinElmer). Equilibrium activity was measured from the sorption samples as follows: 8 ml aliquot of filtered liquid was pipetted into a 20 ml plastic scintillation bottle. 12 ml of scintillation cocktail (PerkinElmer‘s Optiphase ’Hisafe’ 3 or Ultima Gold AB) was added, and the bottle was gently shaken to ensure the mixing of the phases. Activity counting samples were taken to the scintillation counter and let to stabilize from 0.5 hours to 12 hours before the counting to avoid uncertainties created by the presence of air bubbles. Simultaneously with the sorption samples, tracer standards and Milli-Q water background samples. Equation 4 was used to calculate the sample Kd value. All the Kd values were based on the sample dry weight and are thus comparable with each other.

[4]

Equation 4. The calculation of the Kd value. Ai = measured initial activity (ie. the activity of the tracer standard), Af = measured final activity of the solution, Vams = volume (ml) of the activity measurement sample, Vi = initial volume (ml) of the sample and m = dry mass (g) of the sample.

The γ–emitters Cs-134, I-125, Nb-95 and Se-75 were measured by gammaspectrometry with Wizard™ 3” (PerkinElmer Wallac). The tracer final activity in the liquid phase was determined from the sorption samples by pipetting 10 or 15 ml of centrifuged (Nb-95 and Se-75) or typically filtered liquid into a 20 ml plastic scintillation bottle. The samples were taken to the γ-counter and measured immediately if possible. The final activity of Nb-95 and Se-75 in the sorption samples was determined from duplicates. Tracer standard samples and Milli-Q-water background samples were measured together with the sorption samples. The volume of the standard and background samples was equal with the volume of the activity measurement samples. Kd was calculated as presented in Equation 4. The Kd uncertainties presented in the following Chapters and Appendices are geometric standard deviations or standard deviations.

3.2 Batch sorption studies

The sorption of radionuclides on different materials is usually examined and estimated by batch sorption technique (e.g. Aldaba et al. 2010). In this method the solid sample, e.g. soil or sediment, is suspended to a substantially larger volume of liquid. Radioactive tracer is added, suspension is equilibrated for the desired time and radionuclide concentration is measured in the liquid phase after phase separation and liquid filtering (e.g. Hsu & Chang 1994; Lee et al. 2008; Sheppard et al. 2009b, 2011).The distribution of radionuclide between solid and liquid phases is determined and the distribution coefficient (Kd) is calculated.

45

The sorption of Cs-134, Cl-36, I-125, Nb-95, Se-75 and Tc-99 was studied on humus and mineral soil samples in aerobic and/or anaerobic conditions as a function of time. Soil samples were from the excavator pits OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21.

3.2.1 Experiment series programme

Batch sorption tests were done in different phases depending upon the choice of the material and radionuclide, thus creating an experiment series programme. The experiments and programme were started with the samples of OL-KK14, OL-KK15 and OL-KK16. The nuclides Cs-134, I-125 and Tc-99 were examined first. In the second phase, OL-KK20 aerobic and anaerobic mineral soil samples were used in the studies for the same nuclides. Thirdly, the sorption of Cl-36 was examined on mineral soil and humus samples. In the final stage of the experiments Nb-95 and Se-75 were included. The sorption experiments are still continued and the results will be reported later. Table 9 presents the studied radionuclides and soil samples in accordance with the grouping of the sorption experiment series. All the conducted batch series are included. Table 9. Sorption experiment sequence and the grouping of the materials and radioactive tracers Cs-134, Cl-36, I-125, Nb-95, Se-75 and Tc-99.

Sorption experiment series Sampling

site Nuclide

Cs-134 Cl-36 I-125 Nb-95 Se-75 Tc-99

on dried and fractioned soil samples

OL-KK14 x x x OL-KK15 x x x OL-KK16 x x x

as a function of time on aerobic untreated soil samples

OL-KK20 x x x x x x

OL-KK21 x x as a function of time on anaerobic untreated soil

samples

OL-KK20 x x x x x

OL-KK21 x

on dried and fractioned samples of humus layer

OL-KK15 x

as a function of time on untreated soil samples of humus

layer OL-KK21 x x x x x x

as a function of time on irradiated soil samples of

humus layer OL-KK21 x x x

effect of competing cations OL-KK20 x

Table 10 presents a detailed and descriptive summary of the experimental series programme conducted with the radioactive tracers Cs-134, Cl-36, I-125, Nb-95, Se-75 and Tc-99. Figures, corresponding page numbers and Appendices attributed to a certain radionuclide and sorption experiment series are given.

46

Tab

le 1

0. S

umm

ary

of th

e so

rpti

on e

xper

imen

ts c

ondu

cted

wit

h ra

dioa

ctiv

e tr

acer

s C

s-13

4, C

l-36

, I-1

25, N

b-95

, Se-

75 a

nd T

c-99

.

Nu

clid

e S

orp

tion

ex

per

imen

t se

ries

S

amp

lin

g si

te

Fig

ure

s P

age

nu

mb

er

Ap

pen

dix

(p

age

nu

mb

er)

Cs-

134

on d

ried

and

fra

ctio

ned

soil

sa

mpl

es

OL

-KK

14

OL

-KK

15

OL

-KK

16

30, 3

1, 3

2 52

, 53,

54

5.1

(127

), 5

.2 (

128)

as a

fun

ctio

n of

tim

e on

aer

obic

un

trea

ted

soil

sam

ples

O

L-K

K20

O

L-K

K21

33

, 34

55, 5

7 5.

3 (1

29),

5.4

(13

0; O

L-

KK

20),

5.5

(13

1; O

L-

KK

21)

as a

fun

ctio

n of

tim

e on

an

aero

bic

untr

eate

d so

il

sam

ples

O

L-K

K20

34

, 35

57, 5

8 5.

3 (1

29),

5.4

(13

0)

as a

fun

ctio

n of

tim

e on

un

trea

ted

hum

us s

ampl

es

OL

-KK

21

36

59

-

effe

ct o

f co

mpe

ting

cat

ions

O

L-K

K20

37

, 38,

39,

40,

41

61, 6

2, 6

3, 6

3, 6

4 5.

6 (1

32-1

33)

Cl-

36

on u

ntre

ated

soi

l sam

ples

in

aero

bic

cond

itio

ns

OL

-KK

20

- -

-

on d

ried

and

fra

ctio

ned

hum

us

sam

ples

O

L-K

K15

43

67

6.

1 (1

34)

as a

fun

ctio

n of

tim

e on

un

trea

ted

hum

us s

ampl

es

OL

-KK

21

44

68

6.2

(135

)

as a

fun

ctio

n of

tim

e on

ir

radi

ated

hum

us s

ampl

es

OL

-KK

21

- -

6.2

(135

)

I-12

5

on d

ried

and

fra

ctio

ned

soil

sa

mpl

es

OL

-KK

14

OL

-KK

15

OL

-KK

16

45

, 46

71

, 72

7.1

(136

), 7

.2 (

137)

as a

fun

ctio

n of

tim

e on

aer

obic

un

trea

ted

soil

sam

ples

O

L-K

K20

O

L-K

K21

47

74

7.

3 (1

38)

as a

fun

ctio

n of

tim

e on

an

aero

bic

untr

eate

d so

il

sam

ples

O

L-K

K20

-

- 7.

3 (1

38)

as a

fun

ctio

n of

tim

e on

un

trea

ted

hum

us s

ampl

es

OL

-KK

21

48

76

7.4

(139

)

as a

fun

ctio

n of

tim

e on

ir

radi

ated

hum

us s

ampl

es

OL

-KK

21

- -

7.4

(139

)

- F

igur

e/A

ppen

dix

not a

dded

to th

e re

port

46

47

Tab

le 1

0. S

umm

ary

of th

e so

rpti

on e

xper

imen

ts c

ondu

cted

wit

h ra

dioa

ctiv

e tr

acer

s C

s-13

4, C

l-36

, I-1

25, N

b-95

, Se-

75 a

nd T

c-99

.

Nu

clid

e S

orp

tion

ex

per

imen

t se

ries

S

amp

lin

g si

te

Fig

ure

s P

age

nu

mb

er

Ap

pen

dix

(p

age

nu

mb

er)

Nb

-95

as a

fun

ctio

n of

tim

e on

ae

robi

c un

trea

ted

soil

sa

mpl

es

OL

-KK

20

50, 5

1, 5

2 80

, 81,

82

8.1

(140

), 8

.2 (

141)

as a

fun

ctio

n of

tim

e on

an

aero

bic

untr

eate

d so

il

sam

ples

O

L-K

K20

52

, 53

82, 8

3 8.

1 (1

40)

, 8.2

(14

1)

as a

fun

ctio

n of

tim

e on

un

trea

ted

hum

us s

ampl

es

OL

-KK

21

54

84

8.3

(142

)

Se-

75

as a

fun

ctio

n of

tim

e on

ae

robi

c un

trea

ted

soil

sa

mpl

es

OL

-KK

20

56, 5

7 87

, 88

9.1

(143

), 9

.2 (

144)

as a

fun

ctio

n of

tim

e on

an

aero

bic

untr

eate

d so

il

sam

ples

O

L-K

K20

58

89

9.

1 (1

43)

, 9.2

(14

4)

as a

fun

ctio

n of

tim

e on

un

trea

ted

hum

us s

ampl

es

OL

-KK

21

59

90

9.3

(145

)

Tc-

99

on d

ried

and

fra

ctio

ned

soil

sa

mpl

es

OL

-KK

14

OL

-KK

15

OL

-KK

16

61

93

10.1

(14

6)

as a

fun

ctio

n of

tim

e on

ae

robi

c un

trea

ted

soil

sa

mpl

es

OL

-KK

20

62

94

10.2

(14

7)

as a

fun

ctio

n of

tim

e on

an

aero

bic

untr

eate

d so

il

sam

ples

OL

-KK

20

OL

-KK

21

63

95

10.2

(14

7)

as a

fun

ctio

n of

tim

e on

un

trea

ted

hum

us s

ampl

es

OL

-KK

21

64

96

10.3

(14

8)

as a

fun

ctio

n of

tim

e on

ir

radi

ated

hum

us s

ampl

es

OL

-KK

21

- -

10.3

(14

8)

- F

igur

e/A

ppen

dix

not a

dded

to th

e re

port

47

48

3.2.2 Aerobic soil samples

Sorption experiments with radioactive tracers in the aerobic conditions were usually conducted as follows: 1.00 g of soil sample was weighted into a 50 ml Sorvall® polypropylene centrifuge tubes. Soil solution simulant (25 ml) and tracer were added and the tubes were placed onto a reciprocal shaker. After the selected time had elapsed, ranging from 1 day to several weeks, solid and liquid phases were separated by centrifugation (Sorvall®, rotor SS-34 or Beckmann Coulter Avanti J-26 XPI, rotor JA-25.50), and the liquid was further filtered through 0.2 µm syringe filter (Life Sciences, Supor® Membrane) into a 50 ml PE-centrifuge tube (VWR). Equilibrium pH was measured from the filtered solutions. Samples for the activity measurement were done as described in Section 3.1.4.

Batch tests to study the effects of competing cations on the sorption of caesium were done similarly, but the soil solution simulant was replaced with electrolyte solution.

The equilibrium times varied depending on the radionuclide and sampling site. Table 11 presents the equilibrium times used in the sorption experiments of radionuclides.

Table 11. The equilibrium times used for the sorption experiments of Cs-134, Cl-36, I-125, Nb-95, Se-75 and Tc-99 on soil samples of OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21.

Equilibrium time (days)

Nuclide OL-KK14 OL-KK15 OL-KK16 OL-KK20 OL-KK21

humus mineral

soil humus

mineral soil

humus mineral

soil

aerobic mineral

soil

anaerobic mineral

soil humus

aerobic mineral

soil

anaerobic mineral

soil

Cs-134 91 21 91 21 135 30 1-245 1-28 7-23 1-112 -

Cl-36 - - 7-93 - - - 7 - 7-365 - -

I-125 92 14 92 14 92 15 7-147 1-21 1-93 7-63 -

Nb-95 - - - - - - 7-63 1-21 1-28 - -

Se-75 - - - - - - 7-63 7-33 1-63 - -

Tc-99 91 14 91 14 91 14 7-217 7-33 7-93 - 32-364

- sorption experiments not done

Typically two or three tracer activity standard samples, i.e. tracer samples containing the same activity as in the sorption samples but within a known volume (10 ml or 15 ml for γ-emitters and 8 ml for β--emitters) were done when the radioactive tracer was added to the sorption samples. Activity standards were used to calculate the fraction of the radionuclide sorbed on the soil samples, and thus define the Kd value.

3.2.3 Anaerobic soil samples

To study the effect of the redox conditions on the sorption of radionuclides, part of the sorption tests were carried out in the nitrogen filled glovebox. Oxygen and carbon dioxide levels in the N2- atmosphere were below 4 and 0.1 ppm throughout the test series. For many redox sensitive elements, such as technetium and selenium, the change

49

to more reducing conditions typically enhances their sorption (Ashworth & Shaw 2006a; Ashworth et al. 2008; Lieser & Bauscher 1987; Sheppard et al. 1990).

At the beginning of the anaerobic sorption test, a set of 20 ml glass scintillation bottles or polypropylene centrifuge tubes were weighted. Bottles or tubes were transferred to the glovebox where approximately one gram of the soil sample was added to the containers. The containers were brought outside the glovebox, sample mass was determined and the soil solution simulant and tracer solutions, equilibrated with the nitrogen atmosphere prior to the addition, were added to the samples in the glovebox. Samples were left to stand in the glovebox for a chosen time period, during which the samples were shaken daily to ensure the mixing of the phases. Equilibrium times ranged from one day to a year, depending on the nuclide. Equilibrium pH and redox potential were measured, sorption samples in the centrifuge tubes were centrifuged and the liquid phase was filtered. For the samples in the glass bottles, the liquid phase was transferred into a 50 ml PE-centrifuge tube after the pH and Eh measurement and the fine suspended solid material was let to settle down overnight before filtration. Samples for the tracer final activity measurement were done in the glovebox as described in Section 3.1.4.

Figure 29. A picture of AtmosBagTM containing soil filled steel tubes and sorption samples (Photo by Sinikka Virtanen/HYRL).

The sorption experiments for Tc-99 on the anaerobic OL-KK21 mineral soil samples were done similarly, but in the field conditions using AtmosBagTM. 20 ml of soil solution simulant was added to weighted glass scintillation bottles in the glovebox, and the bottles were weighted again. Tc-99 tracer was equilibrated with the nitrogen atmosphere in the glovebox prior to the addition into the water/soil mixture. The equipment needed in the preparation of the anaerobic sorption samples were packed in a plastic box, which was transferred into AtmosBagTM. In the field, freshly taken anaerobic soil samples were transported into AtmosBagTM and the bag was filled with nitrogen gas (Figure 29). Soil samples were added to the bottles until the liquid surface was on the desired level. Tc-99 tracer was added and Tc-99 activity standard samples were done simultaneously. Bottles were closed as tightly as possible and transported to

50

the Laboratory of Radiochemistry in the nitrogen gas filled AtmosBagTM. At the laboratory, bottles were weighted to determine the soil sample mass. After weighting the samples were transferred into the glovebox. The rest of the treatment was done as described above for other samples.

3.2.4 Untreated humus samples

To study if sample drying has an effect on the sorption of radionuclides on humus, an additional test series was done with untreated humus samples of OL-KK21. Tests were done accordingly: 25 ml of soil solution simulant was added into 50 ml polypropylene centrifuge tubes and the tubes were weighted. Humus was added from two to four grams, so that the liquid surface reached the desired level. The tubes were weighted again to define the exact sample mass. After tracer addition the tubes were placed on a reciprocal shaker, equilibrated a chosen time and centrifuged. The liquid phase was filtered through 0.2 µm syringe filter and the equilibrium pH and tracer activity were measured.

3.2.5 Irradiated humus samples

Microbial activity has documented to have a great impact on the sorption of elements (Bunzl & Schimmack 1988; Février et al. 2007). By reducing the microbial activity in the samples, the sorption of radionuclides may be altered drastically. Therefore a complementary sorption series was carried out with γ-irradiated, sterile humus samples. In addition to the microbe population size and their activity, irradiation may cause changes in the physical properties of the samples, such as structure, particle size or solubility.

The γ-irradiated sterilized humus samples were done by weighting 2-3 grams of untreated humus of OL-KK21 into 50 ml polypropylene centrifuge tubes. 25 ml of soil solution simulant filtered through 0.25 µm filter (PES, Nalgene) was added, and the tubes were placed in the vicinity of Co-60 source to receive 20 kGy radiation dose.

After irradiation, the tubes were transferred into the glovebox where the radiotracer was added. The centrifuge tubes were brought outside and placed on a reciprocal shaker, equilibrated a chosen period and centrifuged. The liquid phase was filtered through 0.2 µm syringe filter and the equilibrium pH and tracer activity were measured.

51

4 RESULTS

The results of the Kd values of Cs-134, Cl-36, I-125, Nb-95, Se-75 and Tc-99 in humus and mineral soil layers of the deep soil pits OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21 at the Olkiluoto Island are presented in the following Chapters. The Kd values are compared with the Kd values reported in the literature.

4.1 Kd values for caesium

Caesium has only one oxidation state of +I. It does not hydrolyze nor form complexes in solutions, and occurs solely as hydrated Cs+ ion. Its sorption on mineral surface hydroxyl groups and functional groups of organic matter takes place via an ion exchange mechanism, even though the uptake is not efficient because of the large size and low charge of Cs+ ions. Clay and mica minerals sorb caesium ions very effienctly on their interlamellar ion exchange sites. Concentrations of caesium in natural waters are typically very low.

4.1.1 Dried and fractioned soil samples of OL-KK14, OL-KK15 and OL-KK16

The sorption of caesium on dried and fractioned soil samples of different soil layers of the deep excavator pits OL-KK14, OL-KK15 and OL-KK16 is presented in Figure 30 and Appendix 5. Humus tends to retain caesium poorly compared with the mineral soil layers MS1-MS5, even though there are differences between the pits and the soil layers arising from the variance in the CEC and the clay fraction content. The Kd values for the humus layers of soil pits OL-KK14, OL-KK15 and OL-KK16 were 11 759 ml/g, 198 ml/g and 964 ml/g, respectively. The great variability in the Kd values is due to the variation in the organic matter (OM) and dry matter contents, and especially in the mineralogy of the dry matter. The organic matter content was the highest for the soil pit OL-KK15 (62.6 %) and lowest for OL-KK16 (54.2 %), whereas the highest and lowest dry matter contents were in OL-KK14 (48.0 %) and OL-KK16 (26.5 %), respectively (Lusa et al. 2009). High OM content decreases the sorption of caesium, whereas an increase in the dry matter content has a positive effect on the sorption because caesium is mainly retained by the soil mineral matter (e.g. Absalom et al. 1995; Cremers et al. 1988; Dyer & Cho 1999; Giannakopoulou et al. 2007; Shenbar & Eriksson 1993).

The sorption mechanism of caesium on organic matter is ion exchange, and the concentration of other cations, especially divalent and trivalent cations in the solution restrains the sorption of caesium due the competition of the sorption sites (Bondar & Zabrodskii 2001; Dumat et al. 1997; Dunigan & Francis 1972; Rigol et al. 2002; Staunton & Levacic 1999). The large variability in the Kd values cannot be, however, explained by the differences in the OM and dry matter contents. The most possible reason is the contents of the clay and mica minerals in the humus layers. Even small amounts of highly caesium-selective mica and clay minerals among the organic matter control the level of the sorption. Mineralogical composition of the mineral matter in the humus layers was not determined and therefore no direct evidence to this is available.

52

100 1000 10000 100000

MS5

MS4

MS3

MS2

MS1

humusS

oil

lay

er

Kd (ml/g)

OL-KK14 OL-KK15 OL-KK16

Figure 30. Kd values of Cs-134 for the soil layers of OL-KK14, OL-KK15 and OL-KK16. The Kd values for each sample depth was calculated as a weighted average of the Kd values determined for the grain size fractions of >0.063 mm, 0.063-0.125 mm, 0.125-0.25 mm, 0.25-0.50 mm, 0.50-1.0 mm and 1.0-2.0 mm.

A major fraction of Cs+ is sorbed in the mineral soil layers compared with the humus layer, excluding OL-KK14 (Figure 30). In the mineral soil layers of OL-KK14 and OL-KK15 the Kd values decrease with increasing soil depth and decreasing clay fraction content. The lowest Kd values and clay fraction contents in the mineral soil layers were in the topmost layer MS1. In MS2 soil layers the Kd values were the highest, and so was the amount of clay fractions. The Kd values for MS1 and MS2 soil layers were 3 041 ml/g and 32 787 ml/g for OL-KK14, and 982 ml/g and 18 949 ml/g for OL-KK15. Kd values for the other mineral soil layers vary between these ranges. The result is quite expected, because topsoil layers usually show lower sorption of caesium compared with the subsoil layers. This is due to the increased weathering and the action of soil forming processes influencing the surficial parts of soils. Chemical weathering takes place as a counteract for the soil acidifying caused by the action of the acid load produced by the natural organic acids liberated in the decomposition of organic matter and carbonic acid, nitric acid and sulphuric acid originating from the atmosphere and precipitation. As a consequence the amount of more easily weathered minerals (e.g. olivine and Ca-rich plagioclase) decrease and the amount of resistant minerals (e.g. quartz and potassium feldspar) with relatively low sorption capacities increase (Birkeland 1984; Campbell & Davies 1995; Kónya et al. 2005; Mollah & Ullah 1998; Torstenfelt et al. 1982).

In the soil pit OL-KK16, the highest and lowest Kd values were in the soil layers MS1 (11 288 ml/g) and MS4 (3 157 ml/g). In OL-KK16, the clay fraction content increased as a function of depth, but the Kd values decreased. Equilibrium pH increased with depth in all the soil pits, being the lowest in the humus layer (4.5-5.6) and highest in the bottom mineral soil layers (5.2-8.6). In the mineral soil the most important factors

53

governing the sorption of caesium are the clay mineral content, the nature of the present clay minerals (expanding or non-expanding, CEC) and the amount of clay sized fraction (Absalom et al. 1995, Cremers et al. 1988, Giannakopoulou et al. 2007, Shenbar & Eriksson 1993).

Altogether a general trend to the soil pits OL-KK14, OL-KK15 and OL-KK16 is that the Kd of caesium has a maximum, on average 20 000 ml/g, at the depth of about 60 cm and it systematically decreases towards lower depths, being almost an order of magnitude lower in the deepest sampling depths. Also, at about 30 cm the Kd is considerably lower than at 60 cm. The reason for this behaviour is the most possibly the varying clay contents of the layers presented in Figure 31, which shows the correlation between the Kd values and the clay fraction (< 0.002 mm) content.

Figure 31. The correlation between Cs-134 Kd values and the clay fraction content of the soil samples.

References concerning the determination of caesium Kd values for soil samples taken at different sampling depths as done as extensively as in this study were limited in the literature. For example, Gillham et al. (1990) defined that Cs Kd decreased from 26 000 ml/g to 6 300 ml/g as sample depth increased from 0.9 m to 6.0 m. For another Canadian soil samples, the decrease was from 7 000 ml/g in sample depth 1.0 m to 4 100 ml/g for sampling depths 4.0 m and 10.0 m. The nature of the samples was not described in detail. Compared with these values, the experimental values of the studied soil samples are in a good agreement. The overall average Kd values for the soil pits OL-KK14, OL-KK15 and OL-KK16 were 13 150 ml/g, 6 054 ml/g and 6 239 ml/g, respectively. The overall Kd values were calculated by taking an average of the individual Kd values of the soil layers of each soil pit. The calculation of the Kd values of the soil layers is described in the caption of Figure 30.

In Figure 32 and Appendix 5 are presented the Kd values of Cs-134 on the dried and fractioned soil samples of different soil layers of the excavator pit OL-KK15 as a function of the grain size. Even though the retention typically increases with decreasing particle size (Baeza et al. 1995; Giannakopoulou et al. 2007; Hsu & Chang 1995; Kamel & Navratil 2002; Li et al. 2004; Zygmunt et al. 1998), the effect was not so evident to OL-KK15 samples. Typically the increase in the sorption with decreasing grain size is

y = 1872,2x + 62,373R² = 0,7845

0

5000

10000

15000

20000

25000

0 2 4 6 8 10 12

Kd (ml/g)

Clay fraction content (%)

54

related to the surface sorption site density of the material; the majority of the sorption sites are situated on the particle surfaces, and smaller particles have larger surface-area-to-volume ratio thus increasing their reactivity (Baeza et al. 1995). The retention of caesium on mineral soil particles increase in the order of clay (<2 µm) ~ very fine silt (2-5µm) > fine silt (5-10 µm) > coarse silt (10-50 µm) (Li et al. 2004).

<0.063 0.063-0.125 0.125-0.25 0.25-0.50 0.50-1.0 1.0-2.0

100

1000

10000

Kd (

ml/g

)

Grain size (mm)

humus

MS1

MS2

MS3

MS4

MS5

Figure 32. Kd values of Cs-134 for OL-KK15 sample layers as a function of the grain size.

For the mineral soil layers MS1, MS3 and MS4 the highest Kd values were for the smallest grain size <0.063 mm, and the Kd values decreased with increasing particle size or were rather constant. For MS5, the Kd values decreased from the grain size <0.063 mm to 0.25-0.50 mm and increased again towards larger grain sizes. In the soil layers MS1, MS3 and MS4 changes in the Kd values followed the changes in the cation exchange capacity (Figure 13). The respective Pearson’s correlations for the soil layers of MS1, MS3 and MS4 were 0.9978, 0.9759 and 0.9442. However, for the soil layers of MS2 and MS5 the highest Kd values were in the grain size fractions of 0.25-0.50 mm (20 457 ml/g) and 1.0-2.0 mm (6 093 ml/g). The correlation between the Kd and the cation exchange capacity in MS2 and MS5 were smaller: 0.7359 for MS2 and 0.1944 for MS5.

In the case that the sorption of caesium would take place only on the outer surfaces of the mineral grains, a clear correlation between the Kd and the grain size would have been noticed. As this is not the case, it is assumed that the mineral matter is porous and has sorption sites on the outer surfaces and in its internal space. In addition, dry-sieving, used in the separation of different grain size fractions for another, does not discriminate between the same sized individual mineral particles and soil aggregates consisting of smaller particles cemented together by organic matter and/or weakly crystalline aluminium and iron oxides for instance. Also, it is possible that the surface of the larger grains is covered by an attached dust layer comprised of smaller particles.

55

4.1.2 Aerobic untreated soil samples of OL-KK20 and OL-KK21

The retention of caesium on the aerobic untreated soil samples of the excavator pits OL-KK20 and OL-KK21 decreases as a function of soil depth (Figure 33, Appendix 5) from the topmost sampling depths, 0.7 m (5 456 ml/g) and 0.4 m (7 256 ml/g) to 3.4 m (1 315 ml/g) and 3.6 m (2 742 ml/g). Also, the Kd values for the samples of OL-KK21 were higher than for OL-KK20. Since sample mineralogy, pH, OM and dry matter contents were rather similar and the CEC did not correlate with the Kd values, differences may be partly due to the different grain size distributions (clay fraction contents) of these two pits. For OL-KK14, OL-KK15 and OL-KK16 the Kd increased with increasing clay fraction content (Figure 31), and it is possible that a similar trend is true for OL-KK20 and OL-KK21. However, direct evidence to this is not available since the grain size distribution analysis of the soil samples of OL-KK20 and OL-KK21 was not done.

2000 4000 6000 8000 100004,0

3,5

3,0

2,5

2,0

1,5

1,0

0,5

0,0

So

il d

ep

th (

m)

Kd (ml/g)

OL-KK20 OL-KK21

Figure 33. Kd values of Cs-134 on the aerobic untreated soil samples of the excavator pits OL-KK20 and OL-KK21. The Kd value for a specific sample depth was calculated as an average of all the sorption tests done to the sample depth in question.

The retention of caesium on OL-KK20 and OL-KK21 samples increased with increasing equilibrium time (Appendix 5). This may be explained as an initial fast sorption on the external surfaces of the particles, which is followed by a slow diffusion into the aggregate cores (e.g. Wang & Staunton 2010). The increase in the equilibrium time thus increases the fraction of caesium in fixed, scarcely exchangeable form in soils.

The Kd values determined for the aerobic untreated soil samples of OL-KK20 and OL-KK21 were smaller than for dried and fractioned samples of OL-KK14, OL-KK15 and OL-KK16 at approximately the same depths. For example, Kd for OL-KK14 MS3 layer at the depth of 0.6-1.05 m was 13 970 ml/g, whereas the respective value for OL-KK20 at sample depth 0.7 m was about 5 460 ml/g. Correspondingly, Kd for OL-KK21 at sample depth 1.75 m was 3 670 ml/g, whereas for MS4 (1.05-2.4 m) layer of OL-KK14

56

the respective value was 4 190 ml/g. The average Kd values for the soil pits of OL-KK20 and OL-KK21 were 2 606 ml/g and 4 158 ml/g, while the respective values for OL-KK14, OL-KK15 and OL-KK16 were 13 150 ml/g, 6 054 ml/g and 6 239 ml/g. Mineralogy, CEC and grain size distribution are probably important factors explaining the differences when evaluating the Kd values.

Compared with the literature Kd values for mineral soils including the typical Finnish soil type till, the results of this study are in a good agreement. Generally, the Kd values for caesium on soils range from 31 ml/g to 26 000 ml/g (Absalom et al. 1995; Beneš et al. 1994; Bunzl & Schimmack 1988; Cremers et al. 1988; Gillham et al. 1980; Leqoux et al. 1992; Mollah & Ullah 1998). For (USDA) soil texture types clay, loam and sand, the Kd values decreased from 5 500 ml/g to 3 500 ml/g and to 530 ml/g, respectively (Vidal et al. 2009).

The sorption of caesium on minerals and soils could be affected by pH via the protonation and deprotonation degree of surface functional groups (Baeza et al. 1995; Fujikawa & Fukui 1997; Giannakopoulou et al. 2007; Kim et al. 1996; Torstenfelt et al. 1982). As the pH rises, the proportion of negatively charged, deprotonated surface functional groups acting as cation exchange sites increases. This leads to a higher negative surface charge, which attracts cations from the solution to balance the charge for zero (Avery 1996, Baeza et al. 1995, Fujikawa & Fukui 1997, Giannakopoulou et al. 2007, Kamel & Navratil 2002, Torstenfelt et al. 1982). Thus, the sorption of caesium favours elevated pH values.

Figure 34 and Appendix 5 presents the Kd values determined for the soil samples of OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21 as a function of equilibrium pH. It is evident from Figure 34 that pH does not have a drastic effect on the sorption of caesium. Instead, relatively high and low Kd values are noticed approximately at the same pH values; for example in pH 7.7 Kd values were 901 ml/g (low) and 4 900 ml/g (high). Therefore, it is assumed that the pH has only a minor effect on the sorption of caesium on mineral soils at the relevant pH range of 5-9, and the differences obtained in the Kd values are of variation in the mineralogy and grain size distribution. The equilibrium pH of the anaerobic samples was higher than for the aerobic samples due to the absence of carbon dioxide in the nitrogen filled glovebox.

57

5,5 6,0 6,5 7,0 7,5 8,0 8,5 9,0 9,5 10,0100

1000

Aerobic samples (time 1-7 days)

Aerobic samples (time 18-245 days)

Anaerobic samplesK

d (

ml/g

)

Solution equilibrium pH

Figure 34. Kd values of Cs-134 on the aerobic and anaerobic untreated soil samples of the excavator pit OL-KK20 as a function of equilibrium pH.

4.1.3 Anaerobic untreated soil samples of OL-KK20

The sorption of caesium on the anaerobic untreated soil samples of excavator pit OL-KK20 is presented in Figure 35 and Appendix 5. The retention of caesium decreases as a function of soil depth from 3 319 ml/g at the depth of 0.7 m to 649 ml/g at 3.4 m, which is possibly due to decrease in the clay fraction content and CEC. A similar trend was also noticed for the aerobic soil samples (Figures 30 and 33). The average Kd for the untreated and anaerobic samples from mineral soil layers of soil pit OL-KK20 was 1 629 ml/g.

The Kd values for the anaerobic samples of OL-KK20 were typically smaller than those of the aerobic samples at the same equilibrium time. For example at five days equilibrium time the Kd values for the sample depth of 3.4 m were 1 100 ml/g and 600 ml/g for the aerobic and anaerobic samples, respectively. The reason for the decrease in the retention of caesium in the anaerobic conditions is proposed to be the production of ammonium ions (NH4

+) in reductive redox conditions (Wang & Staunton 2010). Ammonium ions are capable of competing effectively with caesium of the sorption sites and to release sorbed caesium into solution (see Section 4.1.2). This has also been verified in the previous studies of Cremers et al. (1988), Comans et al. (1989) and Hird et al. (1995). In these studies the Kd values for caesium on the anaerobic soil samples ranged from 10 ml/g to 4 700 ml/g when measured in natural groundwater, and from 1 ml/g to 2 000 ml/g in synthetíc groundwater (Holgersson 2009). The Kd values in this study are within these large ranges.

As for the aerobic soil samples, the sorption of caesium in the anaerobic conditions increased as a function of time. The data is presented in Appendix 5. According to Wang & Staunton (2010), for instance, the reason is a fast initial sorption on the particle external surfaces and slow diffusion into the aggregate cores.

58

500 1000 1500 2000 2500 3000 3500

3,5

3,0

2,5

2,0

1,5

1,0

0,5

So

il d

epth

(m

)

Kd (ml/g)

Figure 35. Kd values of Cs-134 on the anaerobic untreated soil samples of OL-KK20. The Kd value for a specific sample depth was calculated as an average of all the sorption tests done to the sample depth in question.

4.1.4 Untreated humus samples of OL-KK21

The sorption of caesium on untreated humus samples of OL-KK21 is presented in Figure 36 and Table 12. The retention decreases with increasing equilibrium time, which is opposite to the behaviour of caesium on the mineral soil samples. The Kd values for the humus samples range from 6 558 ml/g on day seven to 1 822 ml/g on day 28. The sorption on humus on day seven is at the same level as that on the topmost untreated mineral soil samples of 0.7 m in OL-KK20 (3 983 ml/g), whereas compared with the sample depth at 0.4 m of OL-KK21 (7 166 ml/g) the sorption is lower. This is in a good agreement with the previous studies showing that the sorption of caesium on organic matter is typically smaller than on soils’ mineral components (Dumat et al. 1997; Rigol et al. 2002).

The sorption of caesium on humic acids and humin increases with increasing pH and decreasing concentration of competing cations (Dunigan & Francis 1972; Helal et al. 1998). Association with humic acids and humin is non-specific and the sorption is partly or completely reversible (Helal et al. 1998; Hsu & Chang 1994, 1995; Nakamaru et al. 2007; Lofts et al. 2002; Rigol et al. 2002). Micro-organisms contribute to the retention of caesium in organic soils (Parekh et al. 2008, Sanchez et al. 2000), and the retention of caesium is stronger in biotic environments than in abiotic (Tamponnet et al. 2008). Micro-organisms may increase the retention of caesium on soils by increasing the surface area of organic matter and by inducing more sorption sites, removing competing cations from the solution or by biosorption and bioaccumulation through active K+ transport system (Kakiuchi et al. 2002; Tamponnet et al. 2008).

Based on the results of the sorption experiments, it is possible that soil micro-organisms are mainly responsible for the sorption of caesium on OL-KK21 humus samples. The decrease in the sorption over time might be caused by the diminishing of micro-

59

organism population as oxygen and probably nutrient level decreased. Another reason may be the formation of anaerobic conditions and the production of ammonium ions (NH4

+) leading to enhanced competition of the sorption sites.

5 10 15 20 251000

2000

3000

4000

5000

6000

7000K

d (

ml/g

)

Time (days)

Figure 36. Kd values of Cs-134 on the aerobic untreated humus samples of OL-KK21 as a function of time.

According to Bunzl & Schimmack (1988), Holgersson (2009) and Sheppard et al. (2009b), the Kd values determined for peat range from 10 ml/g to 6 000 ml/g. The respective value for organic matter is approximately 270 ml/g (Vidal et al. 2009). When comparing the literature Kd values with the determined values for the untreated humus samples, they are within the range. Table 12. Kd values, standard deviations and equilibrium pH values of Cs-134 on humus samples of OL-KK21.

Time (days) Kd (ml/g) Standard

deviation (ml/g) pH

7 6 558 724 6.91 14 3 465 2 324 6.69 23 1 822 187 7.42

4.1.5 Effect of competing cations on the sorption of caesium

The sorption of caesium is restrained in high salinity solutions due to enhanced competition of the sorption sites with the other dissolved cations, whereas in the solutions of lower ionic strength the competition is weaker (Fujikawa & Fukui 1997; Hsu et al. 2002; Hurel et al. 2002; Kamel & Navratil 2002; Lee et al. 2008; Lieser & Steinkopff 1989a; Torstenfelt et al. 1982; Yllera de Llano 1998). The ability of

60

dissolved cations to inhibit the sorption of caesium on soil decreases in the order Cs+>NH4

+>K+>Na+>Ca2+>Mg2+ (Kamel & Navratil 2002; Shenbar & Eriksson 1993). Table 13 presents the effect of caesium, ammonium and potassium on radiocaesium sorption on Swedish sand, loamy sand, clay loam and clay soils. Stable caesium has a greater effect on the sorption of radiocaesium compared with potassium or ammonium.

Table 13. The effect of stable caesium, ammonium and potassium on the sorption of radiocaesium on different soil types (Shenbar & Eriksson 1993). The initial activity of the Cs+-134 tracer was 5 kBq.

Soil type Kd (ml/g) Kd (ml/g)

0.03 mmol Cs+/l

Kd (ml/g) 0.153 mmol Cs+/l

Kd (ml/g) 0.767 mmol Cs+/l

Kd (ml/g) 100 mmol

NH4+/l

Kd (ml/g) 100 mmol

K+/l

Sand 1 961 96 45 19 15 94 Loamy

sand 3 512 24 7 4 10 36

Clay loam 4 081 58 18 7 14 99 Clay 6 189 1 274 338 103 39 125

Figures 37-41 and Appendix 5 present the effect of competing cations on the retention of Cs-134 on the aerobic soil samples of OL-KK20. Stable Cs+ competes effectively with radiocaesium of sorption sites and can remove radiocaesium from ion exchange sites by isotopic exchange (Avery 1996; Beneš et al. 1994; Comans et al. 1991; Cornell 1991; Gillham et al. 1980; Hsu & Chang 1995; Lujanienė et al. 2006; Wang & Staunton 2010). When soil contains relatively high amounts of stable Cs and clay minerals, radiocaesium is sorbed mainly reversibly via ion exchange on the clay minerals regular sorption sites, termed as RES (Kamei-Ishikawa et al. 2008; Lujanienė et al. 2006). If competing stable caesium is not present, radiocaesium may be sorbed irreversibly on high affinity interlayer sites on the edges (frayed edge sites; FES) (Lujanienė et al. 2006).

According to Figure 37, Appendix 5 and Table 11, relatively low concentrations of stable caesium in solution have a great impact on the sorption of radiocaesium. In soil solution simulant with one week equilibration time Kd values ranged from 4 184 ml/g at the sample depth 0.7 m to 580 ml/g at the sample depth of 3.4 m, the average Kd being 1 827 ml/g. The Kd values for the same sampling depths in 0.01 mM CsCl were 215 ml/g and 49.3 ml/g, whereas in 10 mM CsCl the respective values were only 1.05 ml/g and 0.92 ml/g. In the caesium solutions and the soil solution simulant the Kd values decreased as a function of depth. In the analysed samples of soil solution from Olkiluoto, caesium concentration was 0.03 µM at maximum (Lusa et al. 2009). The lowest concentration in this experiment series was approximately at the same level, as the carrier concentration (stable caesium) for Cs-134 in soil water simulant was 0.02 µM. The concentration of stable caesium in 10 µM solution was more than 3400-times higher compared with the natural level. Thus, it is quite possible that the Kd values for radiocaesium in Olkiluoto soils are similar to the values determined for Cs-134 in pure soil water simulant.

61

0,1 1 10 100 1000

3,5

3,0

2,5

2,0

1,5

1,0

0,5

Kd (ml/g)

So

il d

epth

(m

)

Soil water simulant

0.01 mM CsCl

0.1 mM CsCl

1 mM CsCl

10 mM CsCl

Figure 37. Kd values of Cs-134 on the soil samples of OL-KK20 in the presence of 0.01 mM, 0.1 mM, 1 mM and 10 mM of stable Cs+ as a competing cation.

Radiocaesium is retained much better than sodium on the soil sorption sites in 0.1 mM and 1 mM NaCl solutions (Figure 38, Appendix 5). When sodium concentration increases to 10 mM, the radiocaesium Kd values decrease below the Kd values determined in the soil solution simulant. The Kd values for sampling depths 0.7 m and 3.4 m in 0.1 mM NaCl were 8 892 ml/g and 1 719 ml/g, which are considerably higher than the respective values determined in the soil solution simulant (4 184 ml/g and 580 ml/g). The soil solution simulant contained minor concentrations of potassium, which competes efficiently with caesium of the sorption sites and thus leads to smaller Kd values. In 1000 mM NaCl the Kd values were 163 ml/g and 55.3 ml/g for sampling depths 0.7 m and 3.4 m, respectively. The Kd values for Cs-134 decreased to some extent as a function of depth in NaCl solutions. The maximum sodium concentration in the analysed samples of Olkiluoto soil solution was approximately 0.70 mM (Lusa et al. 2009). It can be concluded that sodium is not very efficient in the competition of the soil sorption sites at the low concentrations found in Olkiluoto soils.

62

1 10 100 1000 10000

3,5

3,0

2,5

2,0

1,5

1,0

0,5

So

il d

epth

(m

)

Kd (ml/g)

Soil water simulant

0.1 mM NaCl

1 mM NaCl

10 mM NaCl

100 mM NaCl

1000 mM NaCl

Figure 38. Kd values of Cs-134 on the soil samples of OL-KK20 in the presence of 0.1 mM, 1 mM, 10 mM, 100 mM and 1000 mM of stable Na+ as a competing cation.

K+ and NH4+ ions can compete efficiently with Cs+ ions of the sorption sites situated on

the soil surfaces and clay mineral interspaces due to their chemical similarity and uniform behaviour (Avery 1996; Brouwer et al. 1983; Comans et al. 1989; Konopleva et al. 2009; Rajec & Shaw 1994; Rigol et al. 2002; Shenbar & Eriksson 1993). K+ ions have been found to be the main competitor to Cs+ in soils and NH4

+ mainly in anoxic sediments and soils (Cremers et al. 1988; Comans et al. 1989; Wang & Staunton 2010). Figures 39 and 40 present the Kd values of Cs-134 on the aerobic soil samples of OL-KK20 in the presence of the increasing concentrations of potassium (Fig. 39) and ammonium (Fig. 40). The Kd values, uncertainties and equilibrium pH values are presented in Appendix 5.

Potassium and ammonium inhibit the sorption of caesium on soil samples much better than sodium. Lower Kd values than in the soil solution simulant were determined in 0.1 mM solutions. The Kd values for sampling depths 0.7 m and 3.4 m in 0.1 mM KCl were 3 032 ml/g and 521 ml/g, whereas the respective values in 0.1 mM NH4Cl were 1 281 ml/g and 399 ml/g. For both, potassium and ammonium ions, zero Kd values for Cs-134 were obtained partly in 10 mM and completely in 100 mM and 1 000 mM solutions. Cs-134 Kd values decreased as a function of depth in KCl and NH4Cl solutions. The maximum potassium concentration in the analysed samples of Olkiluoto soil solution was approximately 0.42 mM (Lusa et al. 2009). Based on the research carried out, potassium and ammonium have considerable affinity to the soil sorption sites, and they compete effectively with caesium; ammonium slightly better than potassium.

63

100 1000 10000

3,5

3,0

2,5

2,0

1,5

1,0

0,5

Soil water simulant

0.1 mM KCl

1 mM KCl

10 mM KCl

100 mM KCl

1000 mM KCl

So

il d

ep

th (

m)

Kd (ml/g)

Figure 39. Kd values of Cs-134 on the soil samples of OL-KK20 in the presence of 0.1 mM, 1 mM, 10 mM, 100 mM and 1000 mM of stable K+ as a competing cation. Negative Kd values are not shown on the logarithmic scale.

10 100 1000 10000

3,5

3,0

2,5

2,0

1,5

1,0

0,5

So

il d

epth

(m

)

Kd (ml/g)

Soil water simulant

0.1 mM NH4Cl

1 mM NH4Cl

10 mM NH4Cl

100 mM NH4Cl

1000 mM NH4Cl

Figure 40. Kd values of Cs-134 on the soil samples of OL-KK20 in the presence of 0.1 mM, 1 mM, 10 mM, 100 mM and 1000 mM of stable NH4

+ as a competing cation. Negative Kd values are not shown on the logarithmic scale.

64

The competitive effect of calcium ions on the sorption of caesium on samples of OL-KK20 is more comparable to that of sodium than of potassium or ammonium. The Kd values for sampling depths 0.7 m and 3.4 m in 0.1 mM CaCl2 were 5 371 ml/g and 862 ml/g. In 1000 mM CaCl2 the respective Kd values were 238 ml/g and 61.4 ml/g (Figure 41). The radiocaesium Kd values decreased as a function of depth in CaCl2 solutions. The maximum calcium concentration in the analysed samples of Olkiluoto soil solution was approximately 0.62 mM (Lusa et al. 2009). Thus, calcium does not compete efficiently with radiocaesium of the soil sorption sites at low calcium concentrations found in Olkiluoto soils.

10 100 1000

3,5

3,0

2,5

2,0

1,5

1,0

0,5

So

il d

epth

(m

)

Kd (ml/g)

Soil water simulant

0.1 mM CaCl2

1 mM CaCl2

10 mM CaCl2

100 mM CaCl2

1000 mM CaCl2

Figure 41. Kd values of Cs-134 on the soil samples of OL-KK20 in the presence of 0.1 mM, 1 mM, 10 mM, 100 mM and 1000 mM of stable Ca2+ as a competing cation.

Based on the results of the competing cations carried out with OL-KK20 mineral soil samples, the efficiency of dissolved cations to inhibit the sorption of radiocaesium on soils decreases in the order Cs+ > NH4

+ > K+ > Ca2+ > Na+. This order was obtained by calculating one single Kd value for each cation, which included all the studied solutions and sample depths. According to Kamel & Navratil (2002) and Shenbar & Eriksson (1993) the effect of competing cations decreased in the order Cs+ > NH4

+ > K+ > Na+ > Ca2+ > Mg2+.

4.1.6 Recommendations of the Kd values for caesium

Figure 42 presents the Cs-134 Kd values as a function of soil depth on soil samples of humus and mineral soil layers of the excavator pits OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21. These values were used in the estimation of the recommended conservative and realistic Kd values for Cs-134 (Table 14). The conservative Kd value was the smallest value in a given data group, e.g. among the humus samples, whereas the realistic Kd value was calculated as the geometric mean (GM) value of the data group. The log-normally distributed log-transformed GM values

65

are typically preferred above arithmetic mean (AM) values in the description of the Kd data (Vidal et al. 2009). The Kd values determined in the experiment serie ‘Competing cations’ were omitted in the estimation procedure because these do not describe the sorption process taking place in the natural conditions. Also, the Kd values determined for the irradiation sterilized samples were not included in the humus data set because microbial activity is typically ubiquitous everywhere in soils. The range of determined Kd values for humus and mineral soil samples in the sorption tests at the specific soil depths are given in Table 14.

100 1000 10000400

350

300

250

200

150

100

50

0

Soil depth: humus 5-30 cm 30-100 cm 100-200 cm >200 cm

So

il d

epth

(c

m)

Kd (ml/g)

Figure 42. Kd values of Cs-134 on the aerobic and anaerobic soil samples of the excavator pits OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21 as a function of soil depth used in the estimation of the recommended conservative and realistic Kd values. The Kd values determined in the experiment serie ‘Competing cations’ and irradiation sterilized humus samples were omitted.

The conservative Kd values of Cs-134 are higher in the mineral soil, 350-863 ml/g, compared with humus, 109 ml/g. The similar behaviour was also seen for the realistic Kd values. The conservative and realistic Kd values did not show a systematic correlation with the soil depth. According to Vidal et al. (2009) the GM Kd value of caesium increased with the clay content being 5 500 ml/g for clay soil. The lowest value was in the organic soil, 270 ml/g (Vidal et al. 2009).

66

Table 14. The range of the Kd values determined for Cs-134 in the sorption tests and the recommended conservative and realistic Kd values.

Soil depth (cm)

Soil type Kd range

(ml/g) N

Conservative Kd value

(ml/g)

Realistic Kd value (ml/g)

humus 109 – 21 223

36 100 1 000

5-30 top mineral soil 863 – 19 498

24 860 5 200

30-100 intermediate mineral soil

573 – 40 039

108 570 7 300

100-200 deep mineral

soil 835 – 13 874

103 830 4 000

>200 deep mineral

soil 350 – 25 853

180 350 1 800

4.2 Kd values for chlorine

Chlorine occurs in solutions as chloride ion Cl-, which is a common anion in soil solutions regardless of the solution pH or the redox conditions. In the soil solution simulant used in this study chloride concentration was 0.36 mmol/l. Interaction of Cl- with mineral surfaces is very weak and in general it is considered as a non-retaining ion. Possible sorption takes place by the uptake on protonated M-OH2

+ groups on the mineral surfaces, and the sorption reaction is favoured by low pH values. It has been proposed that chloride can be incorporated into organic substances, i.e. chlorinated organic compounds are formed.

4.2.1 Aerobic untreated soil samples of OL-KK20

The sorption of Cl-36 on mineral soil samples taken from the depths 0.7 and 3.4 m of OL-KK20 was remarkably low and the calculated average Kd values (± geometric standard deviation) were 0.13±0.26 and 0.10±0.13 ml/g, respectively. Thus, within the uncertainty limits these values are identical. The Kd values for the sample depth 0.7 m varied between 0.00 and 0.64 ml/g, when the respective variation for the sample depth 3.4 m was 0.00-0.28 ml/g.

The retention of chlorine in the chloride form (Cl-) on mineral soils takes place through non-specific sorption (i.e. anion exchange) due to electrostatic attraction between positively charged surfaces and negatively charged chloride ions (Katou et al. 1996, Viers et al. 2001, Yoshida et al. 1992). Sorption is favoured at low pH values below the pHpzc (pH point of zero charge) of the material. At these pH values the surface charge of the soil particles becomes more positive, and a higher proportion of the surface functional –M-OH groups are protonated into –M-OH2

+ form (Viers et al. 2001). In Finnish mineral soils uptake by anion exchange is practically non-existing due to rather high soil pH, thus decreasing the sorption of chloride (e.g. Yli-Halla et al. 2000; Viers et al. 2001). Organic matter is the main sorbent for chlorine (Sheppard et al. 1996, 2006).

67

The Kd values of chloride in mineral soils has been found to be close to 0 ml/g (Baston et al. 1992; Kashparov et al. 2007; Sheppard, et al. 1993) and experimental results of OL-KK20 are in a good agreement with these studies. The typical Kd value for Cl-36 on mineral soil has been in the range of 0-0.63 ml/g (Gil-García et al. 2009; Rumynin et al. 2004; Sheppard et al. 1993). Higher values for indigenous chloride range from 0.036 ml/g to 12 ml/g (Sheppard et al. 2009b).

4.2.2 Dried and fractioned humus samples of OL-KK15

The sorption of Cl-36 on OL-KK15 humus samples dried and sieved into different grain size fractions increased with the equilibrium time (Figure 43, Appendix 6). After one week, the Kd was approximately 0 ml/g for all the fractioned samples. In three months time the Kd ranged from 20 to 61 ml/g. The highest Kd was found for the grain size <0.063 mm, and the lowest for the grain size 0.063-0.125 mm. In the other grain size classes there was little or no variation in the Kd value. Change in the sample pH between one week and three months time was within 0.2 pH units, as the average pH increased from 4.29 to 4.36.

0 10 20 30 40 50 60

2.0-

1.0-2.0

0.5-1.0

0.25-0.5

0.125-0.25

0.063-0.125

<0.063

Gra

in s

ize

(mm

)

Kd (ml/g)

Time 0.25 months Time 3 months

Figure 43. Kd values of Cl-36 on the dried and fractioned humus samples of OL-KK15.

There is no differences in the sorption behaviour of chlorine on mineral soil and dried and fractioned organic soil in one week equilibrium time, but after three months the Kd values for the humus samples are noticeably higher proving the role of organic matter as the main sorbent for chlorine in soils. This indicates that chlorine is sorbed on humic substances dissolved from the organic matter during the equilibration time or chlorination of the solid material took place. In the nature, chloride associates preferably with low molecular weight humic substances compared with high molecular weight humic substances (molecular size >0.01 µm), which may be due to the differences in the functional groups and their amounts (Lee et al. 2001). Also, in soils the chlorination

68

reaction, i.e. the formation of chemical bonds between chlorine and the organic compounds, favours small fulvic and humic substances (Asplund & Grimwall 1991; Bastviken et al. 2007), which are often relatively mobile and bioavailable.

However, the results for the dried and fractioned humus samples of OL-KK15 are for pretreated samples and thus they are not representative to the soil organic matter in the natural conditions. Typically humus samples can be dried but not sieved into different grain size fractions, which was done in order to find differences in the sorption behaviour between the grain size fractions. Thus, the Kd values for humus samples of OL-KK15 should be handled with a precaution.


Untreated humus samples retain Cl-36 rather poorly (Figure 44, Appendix 6); the Kd values are in the range of 0.31 and 1.67 ml/g. The highest Kd values was for the humus sample No. 2, for which the OM content was the highest (14.2 %) and equilibrium pH the lowest (5.5-6.73) among the studied samples. For the humus samples No. 1 and 3, the OM content was 11.2 % and 9.6 %, whereas the equilibrium pH values were 6.68-8.45 and 6.96-7.65, respectively. Longer equilibration decreases the sorption of chlorine and the Kd values drop within one year. This may be due to a decrease in the microbial activity of the samples with time.

0 50 100 150 200 250 300 350 400

1

10

humus 1 humus 2 humus 3

Kd (

ml/g

)

Time (days)

Figure 44. Kd values of Cl-36 on the aerobic untreated humus samples 1, 2 and 3 of OL-KK21 as a function of time.

The test series was conducted with the samples of humus layers of the excavator pits OL-KK15 and OL-KK21 located in the vicinity of the harbour road (OL-KK15) in the northern part of the island, and the main road (OL-KK21) in the southeastern corner of the island (Figure 2). The different locations may induce differences in the chemical and physical composition of the soils by variation e.g. in the vegetation cover. Compared with the dried and fractioned humus samples of OL-KK15, the sorption of chlorine on the untreated humus was very low, only slightly higher than on the mineral soil

69

samples. Reasons in the sorption difference of chloride between OL-KK15 and OL-KK21 samples may be smaller OM content, 62.6% in OL-KK15 compared with 9.6-14.2% in OL-KK21, and higher equilibrium pH (average 4.4 versus 6.0-7.5). Untreated soil samples can be seen as more representative of the organic matter found in the field conditions, and thus Kd values determined for OL-KK21 samples can be considered as more relevant.

The results indicate that the organic matter is only slightly more efficient sorbent for chlorine than mineral soil, even though it can not retain Cl-36 or reduce its migration significantly.

4.2.4 Irradiated humus samples of OL-KK21

Micro-organisms control the concentration of inorganic chloride (Clinorg) in soil solution (Bastviken et al. 2007) and they have an active role in the production of organochloride compounds in the nature (e.g. Asplund & Grimwall 1991, Bastviken et al. 2007; Hjelm et al. 1995; Keppler et al. 2000; Rohlenova et al. 2009; Öberg & Grøn 1998). The organochloride compounds are present everywhere in soil, organic matter and surface waters, and the amount of organic chloride (Clorg) is higher than the amount of inorganic chloride (Clinorg) in soils (Asplund & Grimwall 1991; Rohlenová et al. 2009; Öberg & Grøn 1998). In spite of the multiple roles of the micro-organisms in the cycle of chlorine in soils, in the experiments with the humus samples of OL-KK21, the role of micro-organisms seems to be minor. The determined Kd values are in the same range for the untreated and irradiation sterilized samples, 0.31-1.67 ml/g and 0.80-1.55 ml/g, respectively. The Kd values, their uncertainties and equilibrium pH values are given in Appendix 6 for irradiation sterilized humus samples 1, 2 and 3 of OL-KK21.

4.2.5 Recommendations of the Kd values for chlorine

The data set of the determined Kd values in the sorption tests and used in the calculation of the conservative and realistic Kd values of chlorine consisted of the untreated humus samples of OL-KK21 and mineral soil samples of OL-KK20. The humus samples of OL-KK15 were omitted due to the pretreatment (drying and sieving) and relatively high Kd values compared with the fresh humus samples of OL-KK21. Also, the Kd values for the irradiation sterilized samples were excluded because microbial activity is typically ubiquitous in soils. The range of the determined Kd values for humus and mineral soil samples in the sorption tests are given in Table 15, which also includes the recommended conservative and realistic Kd values for chlorine.

The conservative and realistic Kd values of chlorine are higher on the humus compared with the mineral soil, e.g. the realistic values being 0.77 ml/g and 0.13 ml/g, respectively. The desorption GM Kd value for indigenous chlorine on organic soil has been estimated to be 150 ml/g, whereas the respective value for mineral soil was 1.4 ml/g (Sheppard et al. 2009b). The values determined in this study are far lower, which may arise from the differences in the way of determination; Sheppard et al. (2009b) estimated the Kd values based on the desorption of indigenous chlorine, whereas the Kd values gained in this study are established on the radiotracer sorption. Sheppard et al. (2009b) stated that the desorption GM Kd values would be more presentative because of the formation of less soluble forms of organochlorine by the action of soil microbes or plants, for instance.

70

Table 15. The range of the Kd values determined for Cl-36 in the sorption tests and the recommended conservative and realistic Kd values.

Soil depth (cm)

Soil type Kd range

(ml/g) N

Conservative Kd value (ml/g)


humus 0.0 - 1.94 (0-61)*

39 0.20 0.77**

5- >200 cm

top mineral soil – deep mineral soil

0.0 - 0.64 8 0.0 0.13

* range if OL-KK15 samples are included ** if OL-KK15 organic humus soil samples are taken into consideration when GM value for humus is calculated, the corresponding value would be 1.5 ml/g (n=59). 4.3 Kd values for iodine

Iodine is a redox-sensitive element and is met in oxidation states –I, O, +I and +V in the natural conditions. The typical form of iodine in reducing conditions is iodide ion (I-, oxidation state –I) while oxidising conditions and alkaline solutions favour iodate ions (IO3

-, oxidation state +V). In turn, molecular iodine (I2) is favoured by low pH and high redox potential. It is slightly soluble in water; 2.3 mmol/l at 20 oC. Sorption of iodide on minerals is very weak and takes place in the similar manner to chloride, by uptake to protonated M-O-H2

+ groups on minerals surfaces. The sorption of iodate is in general higher than that of iodide and takes place by ligand exchange mechanism. Sorption of both, iodide and iodate, increases with decreasing pH. As in the case of chlorine, organic matter retains iodine better than mineral soil and formation of iodinated organic compounds is possible.


The sorption of iodine on the dried and fractioned soil samples of the different layers of excavator pits OL-KK14, OL-KK15 and OL-KK16 decrease with increasing sample depth (Figure 45, Appendix 7) corresponding to a decrease in the organic matter content and increase in the pH. In all the studied soil pits, the Kd values are the highest in humus, ranging from 77.6 ml/g for OL-KK16 to 2700 ml/g for OL-KK15. In OL-KK14, OL-KK15 and OL-KK16 the Kd values for MS1 were 54.9 ml/g, 2.14 ml/g and 18.3 ml/g, respectively. These values are higher than determined for the underlying soil layers, where the Kd values typically ranged between 0 ml/g and 2.72 ml/g. Grain size did not have a marked effect on the retention of iodine on soil (see Appendix 7) even though previously sorption has found to increase with decreasing particle size (Fukui et al. 1996; Yuita & Kihou 2005).

71

0,01 0,1 1 10 100 1000

MS5

MS4

MS3

MS2

MS1

humus

Soi

l lay

er

Kd (ml/g)


Figure 45. Kd values of I-125 for the soil layers of OL-KK14, OL-KK15 and OL-KK16. The Kd values for each sample depth was calculated as a weighted average of the Kd values determined for the grain size fractions of >0.063 mm, 0.063-0.125 mm, 0.125-0.25 mm, 0.25-0.50 mm, 0.50-1.0 mm and 1.0-2.0 mm.

The results of this study give support to the previous findings of Whitehead (1973, 1978) of iodine being mostly retained by the soil organic matter. A high positive correlation was found between the iodine Kd values and the soil organic matter content, as the Pearson correlation coefficient ranged from 0.98001 to 0.9999. The equilibrium pH had also a visible effect on the sorption of iodine, the Kd values decreasing with increasing pH. The Pearson correlation coefficient for the Kd value and equilibrium pH ranged from -0.7804 to -0.9139. The influence of the rising pH has been verified for organic and mineral soils previously, and has been explained as a change in the surface charge and the protonation degree of the functional groups (Fukui et al. 1996; Kaplan 2003; Muramatsu et al. 1990; Nishimaki et al. 1994; Yoshida et al 1992; Um et al. 2004; Viers et al. 2001). When soil pH is approximately 3, the sorption is at its maximum because of the electrostatic attraction between negatively charged iodide (I-) ions and positively charged soil sorption sites (Ashworth et al. 2003; Evans & Hammad 1995; Fukui et al. 1996). The decrease in the retention is quite abrupt when the pHpzc (pH point of zero charge) of the soil or mineral is approached (Kaplan 2003). When the pH value rises above the pHpzc, the sorption reaches steady state being in practise non-existing. The sorption mechanism of iodide on soil mineral constituents is typically ion exchange (Lieser & Steinkopff 1989b).

The protonation degree of the surface sorption sites is affected by the pH, but also conversions in the speciation of iodine can be seen; for example in the very acidic pH value of 3, 3 % of iodide (I-) was converted into iodate (IO3

-) and 60 % to molecular iodine (I2) (Fukui et al. 1996). The conversion stage to molecular iodine decreased as pH increased, whereas the conversion to iodate remained within 3-5 % regardless of the pH. Molecular iodine can remain in the liquid phase.

72

In literature only a few references concerning the determination of iodine Kd values for soil samples taken at different sampling depths were found. For example, geometric mean Kd values for soil samples with differing OM contents varied from 2.3 ml/g to 34 ml/g as the OM content increased from 2 % to 10 % (Vidal et al. 2009). In mineral soils Kd values ranged from 0.0041 ml/g to 73 ml/g (Bors et al. 1991; Holgersson 2009; Sheppard et al. 2009b; Vidal et al. 2009). Compared with the literature, the experimental Kd values in this study are in a good agreement for the mineral soil samples. The values for the humus samples are considerably higher than previously obtained. Furthermore, it should be taken into consideration that the samples used in this study were dried and fractioned and the Kd values for the soil layers were calculated as a weighted average of the Kd values of the studied grain sizes (>0.063 mm, 0.063-0.125 mm, 0.125-0.25 mm, 0.25-0.50 mm, 0.50-1.0 mm and 1.0-2.0 mm). The overall average Kd values for the soil pits of OL-KK14, OL-KK15 and OL-KK16 were 132 ml/g, 445 ml/g and 21.5 ml/g, respectively.

Sorption of iodine on the humus layers of OL-KK14, OL-KK15 and OL-KK16 as a function of equilibrium pH is presented in Figure 46. Three separate groups are distinguished on the basis of the equilibrium pH, and each group correspond to a certain soil pit. The highest Kd values were seen for the OL-KK15 samples, which had the lowest equilibrium pH values (4.4-4.5). The Kd values ranged from 1 867 ml/g to 3 800 ml/g. In OL-KK14 pH and Kd values ranged from 4.7 to 4.8 and from 359 ml/g to 1 242 ml/g, respectively. The lowest Kd values were measured for OL-KK16 humus samples, 10.1-163 ml/g, where also the pH values were the highest (5.4-5.6). The average Kd values for the humus layers were 604 ml/g, 2 670 ml/g and 77.6 ml/g in OL-KK14, OL-KK15 and OL-KK16, respectively. The average pH values were 4.7, 4.4 and 5.5. The equilibrium pH seems to affect the sorption of iodine on humus to some extent as Pearson correlation between pH and Kd was -0.7422. Sorption of iodine does not show clear dependence on the grain size distribution. The Kd values, uncertainties and equilibrium pH values are given in Appendix 7.

4,4 4,5 4,6 4,7 4,8 4,9 5,0 5,1 5,2 5,3 5,4 5,5 5,6 5,7

10

100

1000

OL-KK14 humus OL-KK15 humus OL-KK16 humus

Kd (

ml/g

)


Figure 46. Kd values of I-125 for the humus samples of the excavator pits OL-KK14, OL-KK15 and OL-KK16 as a function of equilibrium pH.

73

4.3.2 Aerobic untreated soil samples of OL-KK20 and OL-KK21

Sorption of iodine on untreated aerobic mineral soil samples of OL-KK21 decrease with depth from 136 ml/g for the topmost sample at the depth of 0.4 m to 5.89 ml/g in the second down most sample at the depth of 2.95 m (Figure 47, Appendix 7). The Kd values for the samples taken of OL-KK20 show no clear dependence on depth; the range is from 26.4 ml/g at 2.05 m to 14.2 ml/g at 3.4 m. The Kd values determined for the OL-KK20 samples were higher than for the samples of OL-KK21, excluding the sample depth 0.4 m. The higher Kd values for OL-KK20 can possibly be explained with a higher organic matter content. Equilibrium pH was about the same for the samples of both soil pits and increased slightly with increasing sample depth. The average Kd values for OL-KK20 and OL-KK21 were 20.8 ml/g and 31.5 ml/g, respectively.

Soil micro-organisms affect the sorption of iodine. According to Evans & Hammad (1995), the role of micro-organisms on the sorption of iodine is dual; they increase the sorption and may decrease it by altering the oxidation states of iodine. Iodate is possibly retained better on soil constituents compared to iodide (Ashworth & Shaw 2006a; Ashworth & Shaw 2006b; Fukui et al. 1996; Gil-García et al. 2009), as the reduction of iodate to iodide as a reason for the difference was ruled out by experiments with iodine added in the iodide (I-) form. Furthermore, isotopic exchange between stable and radioactive iodine in soils may take place and have a slight effect on the sorption of iodine on mineral soil (Bors et al. 1991, Muramatsu et al. 1990, Sheppard et al. 1995).

The Kd values determined for the untreated soil samples are somewhat higher than the values for the dried and fractioned samples. The average Kd values for the dried and fractioned mineral soil samples of the layers MS1-MS5 of OL-KK14, OL-KK15 and OL-KK16 were 14.2 ml/g, 0.57 ml/g and 7.45 ml/g, respectively. The corresponding values for the untreated samples of OL-KK20 and OL-KK21 were 20.8 ml/g and 31.5 ml/g. The lower Kd values for the samples of OL-KK14, OL-KK15 and OL-KK16 might be caused by a decrease in the population of micro-organisms during sample pretreatment. For example, Muramatsu et al. (1990) found that sample heating decreases the sorption of iodine. The organic matter content and equilibrium pH fell in the same range for all the studied soil samples and thus do not explain the observed behaviour.

The Kd values determined for the untreated aerobic soil samples were in the range of 0.0041-73 ml/g reported by Bors et al. (1991), Holgersson (2009), Sheppard et al. (2009b) and Vidal et al. (2009).

74

1 10 1004,0

3,5

3,0

2,5

2,0

1,5

1,0

0,5

0,0 OL-KK20 OL-KK21

So

il d

epth

(m

)

Kd (ml/g)

Figure 47. Kd values of I-125 on the aerobic untreated soil samples of the excavator pits OL-KK20 and OL-KK21. Kd values are calculated as an average of all the sorption test done to a specific sample depth.

The Kd values determined for dried and fractioned soil samples are not recommended to be used in the migration modelling of iodine because they may not represent the natural conditions. It is possible that the drying and fractioning of the samples have induced changes in the solid phase. Instead, Kd values for the untreated soil samples may give more relevant results.


The retention of iodine on the anaerobic untreated soil samples of OL-KK20 was smaller than on the aerobic samples. The average Kd for the aerobic samples of soil pit OL-KK20 was 0 ml/g, whereas the respective value for the aerobic samples was 20.8 ml/g. In Appendix 7 is presented the Kd values of iodine on the anaerobic soil samples of OL-KK20.

The reasons for lower sorption on anaerobic soil samples may be the changes in the speciation of iodine, equilibrium pH and population of micro-organisms. In the anaerobic, reducing conditions iodine is mainly present as iodide (I-), which is poorly retained by soil mineral components (Ashworth et al. 2003; Ashworth & Shaw 2006a; Ashworth & Shaw 2006b; Fukui et al. 1996; Um et al. 2004; Yuita 1992). The effect of speciation is typically seen in soils at depths below the groundwater surface, as iodine concentration drastically decreases due to low sorption and relatively high migration with groundwater (Yuita & Kihou 2005). The equilibrium pH for the anaerobic samples was 1.5-2.0 pH units higher than for the aerobic samples, ranging from 9.0 to 9.8. The higher pH values increase the negative surface charge of minerals by increasing the fraction of deprotonated surface sorption sites (M-O-) thus declining the sorption of iodide. The population of anaerobic soil micro-organisms decreases more readily at high

75

iodine concentrations (110-980 mg/l) than the population of aerobic micro-organisms (Sheppard & Hawkins 1995). This may possibly decrease the sorption of iodine in the anaerobic conditions. Organic matter content was the same for the aerobic and anaerobic samples of OL-KK20 and OL-KK21.

The Kd values for iodine (I-) determined in the anaerobic conditions by Holgersson (2009) ranged from -16 ml/g to 19 ml/g. The Kd values for iodine in this study are within these limits. Since sorption of iodine does not take place in the anaerobic conditions (see Appendix 7), the Kd value for iodine can be approximated to be zero for the modelling purposes.


Iodine is typically retained in the soil layer with the highest content of organic matter and the highest degree of microbial activity (Bostock et al. 2003; Whitehead 1973; Whitehead 1978). Iodine is sorbed on organic matter on a large pH scale from 2 to 9 and the sorption is enhanced as the decomposition degree of OM increases (Evans & Hammad 1995; Whitehead 1974a; Whitehead 1974b). Sorption is proposed to take place via the amino acid (H2NRCOOH) and phenolic (R-(aromatic ring) -OH) functional groups (Whitehead 1974b). Sorption mechanism is partly ascribed as a simple anion exchange reaction, because an increase in the concentration of competing anions, e.g. Cl-, NO3

- and F- decrease the sorption of iodine (Sheppard et al. 1995).

Organic matter can retain considerable amounts of iodine, but the sorption takes place only on insoluble humic substances, excluding fulvic acids on which the sorption takes place even in the dissolved form (Bunzl & Schimmack 1988; Fukui et al. 1996; Muramatsu et al. 1990; Sheppard & Hawkins 1995; Sheppard et al. 1995). On the other hand, micro-organisms may initiate the incorporation of iodine to organic matter and the reaction involves extracellular enzymes (Bunzl & Schimmack 1988; Evans & Hammad 1995; Gil-García et al. 2009; Koch-Steindl & Pröhl 2001; Sheppard & Hawkins 1995; Yoshida et al. 1998).

Figure 48 presents the Kd values of iodine as a function of equilibrium time on humus samples taken from the vicinity of excavator pit OL-KK21. The retention of iodine increases for the first five days and decreases afterwards. The highest Kd value of 475 ml/g was found on day five, whereas the lowest Kd 9.59 ml/g was on day 93. Equilibrium pH ranged from 7.0 to 7.7. The decrease in the sorption of iodine on untreated humus samples could be explained as a decrease in the population of micro-organisms with increasing equilibrium time as oxygen and nutrient level drops. Iodine Kd values on untreated humus samples, uncertainties and equilibrium pH values are given in Appendix 7.

76

0 20 40 60 80 1001

10

100

1000

Kd (

ml/g

)

Time (days)

Figure 48. Kd values of I-125 on the aerobic untreated humus samples of OL-KK21 as a function of time.

Sorption of iodine on the dried and fractioned humus layer samples of OL-KK14 and OL-KK15 were considerably higher than on the samples of OL-KK21. The respective Kd values were 604 ml/g, 2 670 ml/g and 220 ml/g. The difference can be explained by the variation in the organic matter content and equilibrium pH. The OM content for the humus samples of OL-KK14, OL-KK15 and OL-KK21 were 58.9 %, 62.6 % and 11.7 %, whereas equilibrium pH values were 4.7, 4.4 and 7.4, respectively.

According to Vidal et al. (2009), the Kd values for iodine on organic soil fall between 2.9-36 ml/g, for peat soils 0-400 ml/g (Bunzl & Schimmack 1988; Sheppard et al. 2009b) and for humic acids 15-40 ml/g (Muramatsu et al. 1990). The Kd values determined for iodine on humus samples in this study are within the ranges reported in the literature.


Sterilization of the samples may destroy the living micro-organisms and their by-products, enzymes (Muramatsu et al. 2004), usually leading to decreased sorption of iodine (Assemi & Erten 1994; Bird & Schwartz 1996; Devivier et al. 2004; Evans & Hammad 1995; Gil-García et al. 2009; Muramatsu & Yoshida 1999; Muramatsu et al. 2004; Yoshida et al. 1998). For example, the Kd values for nonsterilized and γ-irradiation sterilized organic soil samples were 505 ml/g and 150 ml/g, respectively (Evans & Hammad 1995).

The sorption of iodine on three irradiation sterilized humus samples (1, 2 and 3) taken from the vicinity of OL-KK21, are smaller at seven day equilibrium time compared with the nonsterilized humus samples. The Kd values for nonsterilized and sterilized humus samples were 244 ml/g and 2.51 ml/g, respectively. The increase in the sorption of iodine on sterilized samples as a function of time was opposite to the trend seen for the nonsterilized samples; the average Kd value for three humus samples 1, 2 and 3 of

77

OL-KK21 were 6.2 ml/g and 56.6 ml/g for seven days and three month equilibrium times, whereas the respective Kd values for the nonsterilized samples were 244 ml/g and 9.6 ml/g. Equilibrium pH was approximately 2 units lower for sterilized samples (5.4) than for nonsterilized samples (7.4).

The reason for the increase in the sorption on sterilized samples with time might be the increased association of iodine with organic matter. Also, the sorption may take place on the soil mineral constituents mixed among the humus. The growth of aerobic microbe population with time is not a realistic explanation, since oxygen level in the sample tubes would be quite low even though some diffusion through tube walls would take place.

The Kd values determined in other studies for irradiation sterilized soils range from 14 to 2 000 ml/g (Evans & Hammad 1995; Muramatsu et al. 1990). On average, the Kd values for the humus samples determined in this study are in the range of 26.9 ml/g to 34.8 ml/g, which is in good agreement with the values reported in the literature.

4.3.6 Recommendations of the Kd values for iodine

Figure 49 presents the I-125 Kd values as a function of soil depth on humus and aerobic mineral soil samples of OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21. The data set for the humus comprised of the Kd values determined for the samples of OL-KK14, OL-KK15 and OL-KK16. The Kd values for OL-KK21 samples were excluded due to the rather high equilibrium pH (8.1) measured in the sorption tests. Also, the Kd values determined for the irradiation sterilized samples were excluded from the humus data set because microbial activity is typically ubiquitous everywhere in soils. The data set for the aerobic mineral soil comprised of the Kd values determined for the aerobic soil samples of OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21. The estimation of the recommended conservative and realistic Kd values for iodine in the anaerobic conditions was based on the data set of the Kd values determined for the anaerobic OL-KK20 mineral soil samples in anaerobic conditions. It is very likely that there is no differences in the sorption behaviour of iodine in the anaerobic conditions between different sample pits. The range of the determined Kd values for the humus and mineral soil samples in the aerobic and anaerobic sorption tests are given in Table 16, which also presents the recommended conservative and realistic Kd values for iodine in respective conditions.

The estimated conservative Kd value of iodine was rather high for humus, 10 ml/g, whereas the corresponding value for aerobic and anaerobic mineral soil was typically 0 ml/g at different depths. The realistic Kd values for aerobic soil somewhat decreased as a function of depth, being the highest for the humus layer, 460 ml/g, and the lowest at 100-200 cm (3.6 ml/g). The realistic Kd values calculated for the anaerobic soils were about ten times smaller than the smallest values estimated for the aerobic soils, and the values decreased with soil depth from 0.5 ml/g at 30-100 cm to 0.2 ml/g at 100-200 cm and >200 cm.

78

10 100 1000400

350

300

250

200

150

100

50

0


So

il d

ep

th (

cm

)

Kd (ml/g)

Figure 49. Kd values of I-125 on the aerobic soil samples of OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21 as a function of soil depth used in the estimation of the recommended conservative and realistic Kd values.

Vidal et al. (2009) estimated that the GM Kd value of iodine for organic soil was 32 ml/g, whereas the corresponding values for sand, loam and clay were 4.1 ml/g, 8.0 ml/g and 11 ml/g. Compared with these values, the recommended realistic Kd values for the aerobic mineral soil are within the same range, whereas the value for humus is approximately 10-fold.

Table 16. The range of the Kd values determined for I-125 in the sorption tests and recommended conservative and realistic Kd values.

Aerobic soil

Soil depth (cm)

Soil type Kd range

(ml/g) N Conservative

Kd value (ml/g)


humus 10 - 4 368 34 10 460

5-30 top mineral

soil 12 - 77 24 12 32


0.0 - 182 74 0.0 3.8


soil 0.0 - 54 50 0.0 3.6

>200 deep mineral

soil 0.0 - 85 69 0.0 4.9

79

Table 16. Continued.

Anaerobic soil

Soil depth (cm)

Soil type Kd range


Kd value (ml/g)



0.0 - 0.68 4 0.0 0.5


soil 0.0 - 0.35 4 0.0 0.2

>200 deep mineral

soil 0.0 - 0.63 12 0.0 0.2

4.4 Kd values for niobium

In nature the only oxidation state of niobium is +V. Only in very reducing conditions it can be reduced to +IV state. Pentavalent niobium is expected to be present in water solutions as Nb(OH)5 (or HNbO3) in slightly acidic to neutral media. At higher pH values it may turn into Nb(OH)6

- (or NbO3-), which has been expected to decrease its

sorption. As with other penta- and tetravalent metals niobium hydrolyses readily and is thus fairly insoluble. Therefore, the sorption experiments with niobium are complicated due to its tendency to sorb on various surfaces, not only on mineral surfaces but also on surfaces of sorption experiment vials and filters. Experimental data on the sorption behaviour of niobium in soil and other media is rather limited.


The sorption of niobium on aerobic untreated mineral soil samples of the excavator pit OL-KK20 at about pH 4 is presented in Figure 50 and Appendix 8. The Kd values do not show dependence on soil depth but the variation was rather wide, 15 385-127 414 ml/g. Reason for such behaviour is difficult to explain, as mineralogy, OM content, pH and grain size distribution do not change notably between the samples. The average Kd for the samples of OL-KK20 was 67 652 ml/g. In these experiments, the solution equilibrium pH ranged between 3.2 and 4.5, and the average was 3.9.

Sorption and speciation of niobium in soil is rather poorly studied, but it has been noted that niobium is relatively immobile in soils due to its strong sorption on mineral particles (Echeverria et al. 2005). The sorption increases with decreasing particle size (Sheppard et al. 1990), being effective on clay, and moderate or strong on sand and gravel (Baston et al. 1992; Behrens et al. 1982). Compared with the literature results, the experimental results in this study are in the same range or somewhat higher. The Kd values reported in the literature ranged from 10 ml/g to 41 000 ml/g (Baker et al. 1994; Baston et al. 1992; Behrens et al. 1982), compared with 15 385-127 414 ml/g determined for the OL-KK20 samples at about pH 4.

80

20000 40000 60000 80000 100000 120000140000

3,5

3,0

2,5

2,0

1,5

1,0

0,5

Soi

l dep

th (

m)

Kd (ml/g)

Figure 50. Kd values for Nb-95 on the aerobic untreated soil samples of OL-KK20 at about pH 4. The Kd values are calculated as an average of all the sorption test done to a specific sample depth.

At pH values of 7-8 (Figure 51) the average Kd value of Nb-95 for the soil samples of OL-KK20 was 60 573 ml/g, which is comparable with the value obtained at pH 4 (67 652 ml/g). Due to the tendency of niobium to sorb on various experimental surfaces, additional tests were carried without filtering the liquids after equilibration. This yielded a clearly lower average Kd value of 10 354 ml/g. The reason for this can be either that the filters sorb dissolved niobium or filtering removes colloidal particles on which niobium is adsorbed. In the former case filtering would overestimate Kd while in the latter filtering would give a realistic value.

The Kd values showed slight dependence on soil depth (Figure 51), as Kd was the highest in the topmost sample depth (0.7 m) and the lowest in the bottom sampling depth (3.4 m). This behaviour was evident to both unfiltered and filtered samples. However, the Kd values for filtered samples were noticeably higher than the Kd determined from the unfiltered samples; 26 957-153 621 ml/g opposed to 5 042-20 397 ml/g, respectively. An increase in the Kd value upon filtering has previously been explained as sorption of Nb on small suspended particles and colloids which are removed from the solution upon filtering (Baker et al. 1994; Baston et al. 1992). Therefore, blank experiments, where the sorption experiment procedure was repeated without any solid material, were carried out. The decrease in the Nb-95 solution activity of the blank samples upon filtering was only 2.1 %, whereas for the sorption samples the decrease was from 43.9 % to 71.4 %. Thus, it was concluded that the higher Kd values obtained upon filtering are due to the removal of colloidal particles bearing niobium, and therefore these values are more representative.

This conclusion is also supported by the study of Baston et al. (1992), in which the Kd values of Nb-95 on clay increased from 240-380 ml/g for unfiltered samples to 2100 ml/g and <10 000 ml/g for samples filtered through 0.45 µm and 0.03 µm filters, respectively.

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0 20000 40000 60000 80000 100000 120000 140000 160000

3,5

3,0

2,5

2,0

1,5

1,0

0,5

Soi

l de

pth

(m)

Kd (ml/g)

OL-KK20; unfiltered OL-KK20; filtered

Figure 51. Kd values for Nb-95 for unfiltered, centrifuged and filtered aerobic untreated soil samples of OL-KK20 at pH 7-8. The Kd values are calculated as an average of all the sorption test done to a specific sample depth.

Figure 52 presents Nb-95 Kd values for the untreated aerobic and anaerobic soil samples of excavator pit OL-KK20 as a function of equilibrium pH. The pH does not have a significant effect on the sorption of Nb-95, but a great variation between the different samples is notable. For example, at pH values 3.6, 3.6 and 3.7 Kd values of 510 ml/g, 5 500 ml/g and 52 000 ml/g were observed. References for the sorption of niobium on soil samples at different pH values were not available, but Anderson et al. (1979) found that the sorption of niobium decreased with increasing pH in alkaline solutions when pH was higher than 8. For example, Nb-95 Kd on Swedish granite decreased from 10 000 ml/g to 100 ml/g as the pH increased from 8.0 to 10.5. Such alkaline conditions were not met in these sorption studies, but in the pH range 2.8-8.1 pH was not found to be a dominant factor affecting the sorption of niobium on soil samples.

82

2,0 2,5 3,0 3,5 4,0 4,5 5,0 5,5 6,0 6,5 7,0 7,5 8,0 8,5

1000

10000

100000

aerobic samples

anaerobic samples

Kd (

ml/g

)


Figure 52. Kd values of Nb-95 for filtered aerobic and anaerobic soil samples of OL-KK20 as a function of equilibrium pH. The concentration of pH buffer solution CHES in the anaerobic samples was 1.0x10-3 M.


Sorption of Nb-95 on the untreated anaerobic mineral soil samples of OL-KK20 is presented in Figure 53 and Appendix 8. The Kd is dependent on sample depth, decreasing with increasing depth. For example, the Kd in the topmost sample depth (0.7 m) was 48 271 ml/g, whereas the corresponding value in the lowest sample depth (3.4 m) was 2 709 ml/g. The average Kd for the whole soil pit was 12 089 ml/g, and the average pH was 5.69. Compared with the Kd values determined for the aerobic samples at about pH 4, the anaerobic Kd values are significantly lower excluding sample depth 0.7 m, for which higher Kd was found (30 132 ml/g versus 48 271 ml/g). The average Kd for the whole soil pit in the aerobic conditions was 67 652 ml/g, and the average pH was 3.9. There was no clear correlation between the Kd values and pH (Fig. 52), and thus the difference in the pH could not be reason for the discrepancy between aerobic and anaerobic experiments.

83

1000 10000 100000

3,5

3,0

2,5

2,0

1,5

1,0

0,5

Soi

l dep

th (

m)

Kd (ml/g)

Figure 53. Kd values for Nb-95 for anaerobic untreated soil samples of OL-KK20. Samples contained CHES-buffer in 1.0x10-3 M concentration. The Kd values are calculated as an average of all the sorption test done to a specific sample depth.

References for the sorption of niobium on soil samples in anaerobic conditions were not available, but for granites the sorption preferred aerobic conditions (Andersson et al. 1979). The respective Kd values were approximately 10 000 ml/g and 1 200 ml/g at pH 8 for the aerobic and anaerobic conditions. The experimental results for the untreated OL-KK20 mineral soil samples are similar to the results of the granite samples.


The Kd values for Nb-95 on the untreated humus samples of OL-KK21 are given in Figure 54 and Appendix 8. The Kd values for the humus samples 1 and 3 increase slightly as a function of time, excluding day one. The initial (1 day) and final (28 days) Kd values for the unfiltered humus sample 1 are 990 ml/g and 1447 ml/g, whereas the respective values for the humus sample 3 are 1077 ml/g and 1350 ml/g. The Kd values calculated for the humus sample 2 somewhat decreased with time, as the Kd in day 1 was 781 ml/g and 669 ml/g on day 28. The pH values for the humus sample 2 was 1-1.5 pH units lower than for the humus samples 1 and 3. The Kd values for the filtered samples are higher, which is explained as the sorption of Nb-95 on small, colloidal particles. The difference in the activity of the blank experiments for unfiltered and filtered samples was 2.1 %, whereas for the humus samples the difference ranged from 3.3 % to 6.3 %. Thus, it is concluded that the higher Kd values obtained upon filtering the samples are more representative.

84

0 5 10 15 20 25 30400

600

800

1000

1200

1400

1600

1800

Kd

(ml/g

)

Time (days)

humus 1; unfiltered

humus 1; filtered

humus 2; unfiltered

humus 2; filtered

humus 3; unfiltered

humus 3; filtered

Figure 54. Kd values of Nb-95 for unfiltered and filtered aerobic untreated humus samples of OL-KK21 as a function of time.

The sharp decline in the Kd values after one day equilibrium time might be explained as the bioaccumulation of Nb-95 into micro-organisms and liberation back to solution, after which niobium is retained on present organic matter and mineral components. The release of Nb-95 from micro-organism may be natural excretion or liberation due to the perishing of the micro-organisms. Bioaccumulation is the active uptake of radionuclides by living micro-organisms, which affects the biogeochemical cycle and sorption e.g. of chloride (Bastviken et al. 2007; Kakiuchi et al. 2002). The variance in the sorption behaviour of Nb-95 on the humus samples may be due to the differences in the sample composition; the OM content was the highest for humus sample 2. If the organic matter is the main sorbent for niobium and other factors had insignificant effect, the highest sorption would have been seen for humus 2 samples. On the other hand, if the effect of mineral matter would have been pronounced in the humus samples, the highest sorption would naturally been seen for humus sample 3 having the highest dry matter content. Average equilibrium pH of the unfiltered humus samples 1 and 3 were 7.1 and 7, whereas the corresponding value for humus sample 2 was 5.3. The pH of the filtered samples were typically approximately 0.1-0.5 pH units higher than for the unfiltered samples. It is likely that one sole factor dominating the sorption reaction on humus samples can not be named, but more factors act simultaneously.

Compared with the literature Kd values determined for peat and organic soils, the Kd values in this study were almost at the same scale or considerably lower. The Kd value of 14 000 ml/g for peat has been obtained by analysing the native element (stable Nb-93) content from peat samples and their’s extracted pore water (Sheppard et al. 2009b), while arithmetic mean value for organic soils has determined to be 2000 ml/g (Gil-García et al. 2009; Sheppard & Thibault 1990; Vidal et al. 2009).

The sorption of niobium on humus samples is much lower than on anaerobic or aerobic mineral soil samples. The average Kd for the unfiltered humus samples 1, 2 and 3 were

85

1066 ml/g, 623 ml/g and 915 ml/g, respectively, whereas the corresponding values were 12 089 ml/g (filtered) and 10 354 ml/g (unfiltered) for the anaerobic and aerobic mineral soil samples. It has been suggested that the role of organic matter in the fixation of Nb in soils may be substantial (Gerbazek et al. 1994), but based on the results of this study, niobium is mainly retained in the mineral soil, more efficiently in aerobic soil conditions than in anaerobic.

4.4.4 Recommendations of the Kd values for niobium

Figure 55 presents the Nb-95 Kd values as a function of soil depth on humus and mineral soil samples. The data set for the humus comprised of the Kd values determined for filtered OL-KK21 samples. The data set for mineral soil comprised of the Kd values determined for the filtered soil samples of the excavator pit OL-KK20 conducted at about pH 7-8. Filtered samples were used due to the tendency of niobium to sorb on small suspended particles and colloids. The Kd values determined in pH of about 4 were omitted because such a low pH is not typically met in soils. Also, the sorption tests conducted in the anaerobic conditions were omitted because the formation of niobium-CHES complex could not be ruled out. The range of the determined Kd values for the humus and mineral soil samples in the sorption tests are given in Table 17, which also presents the recommended conservative and realistic Kd values for niobium.

1000 10000 100000

350

300

250

200

150

100

50

0

Soil depth: humus 30-100 cm 100-200 cm >200 cm

So

il d

epth

(cm

)

Kd (ml/g)

Figure 55. Kd values of Nb-95 on the aerobic, filtered soil samples at about pH 7-8 of OL-KK20 as a function of soil depth used in the estimation of the recommended conservative and realistic Kd values.

The conservative and realistic Kd values of niobium on the humus are lower than on the mineral soil in all soil depths. In the mineral soils the Kd values decrease with soil depth. The best estimate values for niobium on Swedish soils ranged from 16 000 ml/g to 940 000 ml/g depending on the soil type (Sheppard et al. 2009b). On the other hand, the GM Kd values for loam and clay were 2 500 ml/g and 2 400 ml/g, respectively

86

(Vidal et al. 2009). Compared with these values, the recommended realistic Kd values are approximately in the same range with GM Kd values for Swedish soils, but considerably higher than the values estimated by Vidal et al. (2009).

Table 17. The range of the Kd values determined for Nb-95 in the sorption tests and recommended conservative and realistic Kd values.

Soil depth (cm)

Soil type Kd range


Kd value (ml/g)


humus 242 -1912 24 240 870


19 587 –4069 345

8 19 500 100 000


soil 11 743 - 112 837

8 11 700 34 400

>200 deep mineral

soil 2 172 -102 055

24 2100 23 500

4.5 Kd values for selenium

Selenium is a non-metal and its chemistry resembles that of sulphur. The chemistry of selenium is complicated due to several possible oxidation states of –II (selenide Se2-), 0 (elemental selenium S), +IV (selenite SeO3

2-) and +VI (selenate SeO42-). Selenate and

selenite occur in oxic conditions, elemental selenium in intermediate redox conditions while selenide is the prevailing form in most reducing conditions. Selenate is sorbed on mineral surfaces by ion exchange reaction whereas selenite can also form stronger bonds by ligand exchange mechanism. Sorption of both forms is favoured by decreasing pH due to an increase of negatively charged sorption sites on mineral surfaces. Selinide, similarly to sulphide, forms sparingly soluble compounds with divalent transition metal cations, especially with ferrous ions. Since the concentration of sulphur in soil is much higher than that of selenium it is probable that selenide is coprecipitated with sulphide compounds. Behaviour of elemental selenium is not well known. The initial speciation of the selenium tracer used in the sorption experiments was selenite, but possible transformations in the speciation during the sorption experiments were not studied.


Sorption of selenium on untreated, aerobic soil samples of OL-KK20 shows no clear dependence on soil depth but lower Kd values in the two lowest sampling depths at 3.0 and 3.4 m were comparable (Figure 56, Appendix 9). The average Kd for the whole soil pit calculated from all the unfiltered centrifuged samples was 95 ml/g. The Kd values ranged from 9.5-244 ml/g and the highest Kd was in the uppermost sample depth at 0.7 m, whilst the smallest was in the lowest sample depth at 3.4 m. The difference between the Kd values for unfiltered and filtered samples ranged from 1 ml/g to 15 ml/g indicating that only minor, if any, fraction of selenium was sorbed on small suspended particles or colloids.

87

0 50 100 150 200 250 300

3,5

3,0

2,5

2,0

1,5

1,0

0,5

Soi

l dep

th (

m)

Kd (ml/g)

OL-KK20; unfiltered OL-KK20; filtered

Figure 56. Kd values for Se-75 for the unfiltered, centrifuged and filtered aerobic untreated soil samples of OL-KK20. The Kd values are calculated as an average of all the sorption tests done to a specific sample depth.

The sorption of selenium on mineral soil increases with increasing organic matter (OM), clay fraction, iron and aluminium oxyhydroxide contents, decreasing pH and with enhanced weathering of the soil material (e.g. Choppin et al. 2009; Dhillon & Dhillon 1999; Dhillon et al. 2010; Keskinen et al. 2009; Neal et al. 1987; Pezzarossa et al. 1999; Pezzarossa & Petruzelli 2001; Vuori et al. 1989; Yläranta 1983). Also, the role of micro-organisms on the sorption of selenium on mineral soils has been found to be important (Ashworth & Shaw 2006a; Darcheville et al. 2008; Février et al. 2007). The OM content of the studied soil samples was rather equal and thus do not explain the noticed sorption pattern. However, it is possible that in spite of the similar grain size distribution of the samples, the composition of the fractions differed e.g. in the content of weakly crystalline aluminium and iron oxides thus establishing the noted difference. Solution equilibrium pH ranged between 6.7 in the uppermost layer and 8.0 in the lowest layers. The Kd values increased with equilibration time up to nine weeks but no essential increase was observed during the lats three weeks.

The effect of solution equilibrium pH on Se-75 Kd values is presented in Figure 57 for all the soil samples of OL-KK20, both aerobic and anaerobic. The Kd values, their uncertainties and equilibrium pH values are given in Appendix 9. There is no clear correlation between increasing Kd and decreasing pH, as would be expected from the increasing protonation degree of –M-OH groups to –M-OH2

+ and arisen positive surface charge of the mineral surfaces with decreasing pH (Choppin et al. 2009; Dhillon & Dhillon 1999, Viers et al. 2001). Instead, rather constant Kd values around 200 ml/g was observed at pH < 7.5, representing the pH values of soil samples at 0.7 m of OL-KK20. A slight increase in the Kd is possible at the lowest measured pH values < 6. At pH values 7.6-8.2, representative of the deeper mineral soil layers, the Kd values varied from 2 ml/g to 650 ml/g. This reveals that another factor or factors than pH is

88

responsible for the variation; probably the cause is in the redox conditions. It is possible that the highest Kd values in the uppermost sampling depth of 0.7 m are due to the lower pH and higher OM content, but the reasons for the variation of the Kd values in the deeper layers still remains unclear.

The experimental results of the Kd values in this study are in a good agreement with the previous results of other researchers for different types of mineral soils. The Kd range of 1.3-1305 ml/g has been reported by Darcheville et al. (2008), Dhillon & Dhillon (1999), Gil-García et al. (2009), Sharmasarkar & Vance (2002) and Vidal et al. (2009). Typically the efficiency of fine textured soils to retain selenium is higher than that of coarse textured soils, but the affinity of selenium towards organic matter is even higher (Ashworth and Shaw 2006). For example, the Kd values of selenium on organic soil, clay and sand were 1000, 240 and 56 ml/g, respectively (Vidal et al. 2009). Yläranta (1983) determined the range for the selenium Kd values for Finnish fine sand soils to be 10.6 - 23.3 ml/g.

6,0 6,5 7,0 7,5 8,0 8,5

1

10

100

Aerobic; centrifuged, unfiltrated

Aerobic; filtrated

Anaerobic; filtrated

Kd (

ml/g

)


Figure 57. Kd values of Se-75 for the aerobic and anaerobic soil samples of OL-KK20 as a function of equilibrium pH.


The sorption behaviour of selenium in the anaerobic conditions does not differ significantly from that in the aerobic conditions. The largest difference was that about 30-times lower Kd values were observed for soil depth 2.05 m in the anaerobic conditions. At the other sample depths the differences were smaller (Figure 58).

Results from the anaerobic sorption tests are interesting, because typically the migration rate and mobility decrease and retention increases in the anaerobic, reducing conditions due to the formation of reduced forms of selenium; elemental selenium (Se(0)) and selenide (Se(-II)) (Ashworth & Shaw 2006a). These forms are usually encountered in the solid phase, whereas selenite (SeO3

2-(+IV)) and selenate (SeO42-(+VI)) are more

89

soluble and normally met in the liquid phase (Masschelyen et al. 1991; Pyrzyńska 1998). The sorption of selenite on minerals, soils and sediments is more efficient than the sorption of selenate (Collins et al. 2006; Duc et al. 2003; Goldberger & Glaubig 1988; Sharmasarkar & Vance 2002). The transformations between the reduced and oxidised forms of selenium in soils are usually relatively slow and can go through microbiological or abiotoc route (Harada & Takahashi 2009; Ryser et al. 2006; Zawilanski & Zavarin 1996). Altogether, micro-organisms have an important role in the biogeochemistry, speciation and distribution of selenium in soils (Darcheville et al. 2008).

The average Eh of the samples of OL-KK20 was only +110 (range +60-200) RmV, which is regarded as oxidizing environment. For example, the redox potential needed to convert selenite to elemental selenium at pH 8-8.5 is 0- -50 mV (Takeno 2005). On average, an increase of +88 RmV was noticed in the redox potential when comparing the Eh values at seven days and 31 days equilibrium time. Based on the measured Eh values and their changes, actually anaerobic, reducing conditions were not achieved. The sorption of selenium in the anaerobic conditions needs more research so that reliable and descriptive Kd values can be used in the migration modelling.

1 10 100

3,5

3,0

2,5

2,0

1,5

1,0

0,5

Soi

l de

pth

(m)

Kd (ml/g)

Figure 58. Kd values of Se-75 for the filtered anaerobic untreated soil samples of OL-KK20. The Kd values are calculated as an average of all the sorption test done to a specific sample depth.

Kd values for mineral soils determined in anaerobic conditions were not found in the literature. The Kd values for sea basin sediments in the anaerobic conditions have been found to range from 412 ml/g to 9520 ml/g (Collins et al. 2006). Comparing with these values, the Kd values in this study are significantly smaller (1.5-100 ml/g). The failure to reach reducing conditions for the anaerobic soil samples is the probable reason for such low Kd values. The relatively short time period used in our sorption tests could be one of the reasons for rather high Eh values. The equilibrium time affects the redox conditions of the samples by lowering the oxygen level in the nitrogen atmosphere of the glovebox with time. Also, it is possible that any redox reactive components or pairs,

90

e.g. Fe3+ and Fe2+, were not present in the samples or their concentration was too low for the redox potential to change. It is also possible that the Eh electrode could not track the changes in the redox potential, and instead, the measured value was affected by the pH.


The sorption of selenium on untreated humus samples 1, 2 and 3 of OL-KK21 increase with time (Figure 59). For example, the Kd values for centrifuged, unfiltered samples of humus sample 3 increased from 57 ml/g on day one to 785 ml/g on day 63. The increase in the Kd values was even more drastic for humus sample 1, whereas for humus sample 2 the increase was only slight. The average Kd values for the humus samples 1, 2 and 3 were 419 ml/g, 298 ml/g and 313 ml/g, respectively. The average equilibrium pH values for the respective samples were 6.9, 4.5 and 6.6.

Compared with the mineral soil samples, the Kd values of the humus samples were fairly higher. The average Kd values for unfiltered/filtered aerobic samples and filtered anaerobic samples were 95/101 ml/g and 35 ml/g, respectively, when the combined average Kd for the unfiltered/filtered humus samples 1, 2 and 3 was 343/356 ml/g. The results suggest that selenium is mainly retained in the soil layers rich with organic matter, which is in good agreement with the previous studies where organic matter was found to be the primary sorbent for selenium (e.g. Choppin et al. 2009; Gustafsson & Johnsson 1992; Pezzarossa et al. 1999). The mechanism of association with organic matter is proposed to be mediated by soil micro-organism, as micro-organisms reduce selenium to lower oxidation states and enhance its sorption to humic substances (Dhillon et al. 2010; Gustafsson and Johsson 1992). In soils, selenium is primarily retained in the soil organic layer, topmost 0-4 cm, where the complex-forming processes occur (Choppin et al. 2009; Gustafsson & Johnsson 1992; Lusa et al. 2009). The retention of selenium on soils decreases with decreasing organic matter content (Pezzarossa et al. 1999).

The Kd value of selenium in organic soils (organic matter content ≥ 20 %) ranged from 230 ml/g to 1800 ml/g (Gil-García et al. 2009; IAEA 1994; Sheppard & Thibault 1990; Vidal et al. 2009). The Kd values attained in this study for humus samples, 57-1394 ml/g, are within these limits.

91

0 10 20 30 40 50 60 70

0

200

400

600

800

1000

1200

1400

1600

Kd

(ml/g

)

Time (days)

humus 1; unfiltered

humus 1; filtered

humus 2; unfiltered

humus 2; filtered

humus 3; unfiltered

humus 3; filtered

Figure 59. Kd values of Se-75 for the unfiltered, centrifuged and filtered aerobic untreated humus samples of excavator pit OL-KK21 as a function of time.

4.5.4 Recommendations of the Kd values for selenium

Figure 60 presents the aerobic Se-75 Kd values as a function of soil depth on humus and mineral soil samples of OL-KK20. The data set for the humus comprised of the Kd values determined from the filtered samples of OL-KK21. The data set for the mineral soil consisted of the Kd values determined from the filtered soil samples of OL-KK20. On the other hand, the estimation of the recommended conservative and realistic Kd values for Se-75 on the anaerobic Olkiluoto soils was based on the data set of the Kd values determined for the anaerobic OL-KK20 mineral soil samples in the anaerobic conditions. The range of the determined Kd values for the humus and mineral soil samples in the sorption tests are given in Table 18, which also presents the recommended conservative and realistic Kd values for selenium in the aerobic and anaerobic conditions.

The realistic Kd value of selenium for humus was approximately equal with the uppermost aerobic mineral soil (30-100 cm), but considerably higher than for deeper mineral soil layers (Table 18). The respective values for anaerobic soil decreased with soil depth. Anaerobic conditions are probably prevailing at soil depths below 100 cm or at least 200 cm. The best estimate values for selenium on Swedish soils ranged from 85 ml/g to 360 ml/g depending on the soil type (Sheppard et al. 2009b). On the other hand, the geometric mean Kd values for sand, loam and clay were 56 ml/g, 220 ml/g and 240 ml/g, respectively (Vidal et al. 2009). Compared with these values, the recommended realistic Kd values are in the same range.

92

0 200 400 600 800 1000 1200 1400 1600 1800

350

300

250

200

150

100

50

0

Soil depth: humus 30-100 cm 100-200 cm >200 cm

So

il d

ep

th (

cm

)

Kd (ml/g)

Figure 60. Kd values of Se-75 on the aerobic, filtered soil samples of OL-KK20 as a function of soil depth used in the estimation of the recommended conservative and realistic Kd values.

Table 18. The range of the Kd values determined for Se-75 in the sorption tests and recommended conservative and realistic Kd values.

Aerobic soil Soil depth

(cm) Soil type

Kd range (ml/g)

N Conservative

Kd value (ml/g) Realistic Kd value (ml/g)

humus 57 – 1 669

36 57 240


180 - 464 16 180 250


soil 2.26 - 408 16 2.2 17

>200 deep mineral

soil 2.44 - 654 48 2.4 17

Anaerobic soil Soil depth

(cm) Soil type

Kd range (ml/g)

N Conservative



85 - 116 4 85 116


soil 20 - 101 4 20 46

>200 deep mineral

soil 1.36 - 28 12 1.4 4.7

93

4.6 Kd values for technetium

Technetium occurs in aerobic conditions as an anionic species of TcO4- (pertechnetate)

with the oxidation state of +VII. At low redox potentials pertechnetate is reduced to tetravalent Tc4+ which is readily hydrolysable and thus forms insoluble oxides/hydroxides or is strongly sorbed on hydroxyl groups on mineral surfaces. Pertechnetate, in turn, is very mobile and does not pratically interact with mineral surfaces due to its large size and low negative charge. In the experiments of this study technetium was initially added in the pertechnetate form.


The sorption of technetium in the pertechnetate (TcO4-) form on dried and fractioned

soil samples of different soil layers of OL-KK14, OL-KK15 and OL-KK16 is presented in Figure 61 and Appendix 10. In OL-KK14 and OL-KK15 the highest Kd values were for the humus layer, 15.1 ml/g and 126 ml/g, respectively, whereas in OL-KK16 the Kd in the humus layer, 1.44 ml/g, was at same level as in the mineral soil layers. The low Kd value for OL-KK16 humus sample may be due to its high pH, 6.2, compared with the pH values 5.0 and 4.6 for OL-KK14 and OL-KK15, respectively. In mineral soil layers the average Kd values were 0.50 ml/g, 0.12 ml/g 1.00 ml/g for OL-KK14, OL-KK15 and OL-KK16, respectively, and there was no trend with soil depth.

1 10 100

MS5

MS4

MS3

MS2

MS1

humus

Kd (ml/g)


So

il la

yer

Figure 61. Kd values of Tc-99 for the soil layers of OL-KK14, OL-KK15 and OL-KK16. The Kd values for each sample depth was calculated as a weighted average of the Kd values determined for the grain size fractions of >0.063 mm, 0.063-0.125 mm, 0.125-0.25 mm, 0.25-0.50 mm, 0.50-1.0 mm and 1.0-2.0 mm. Negative Kd values are not shown on logarithmic scale.

94

Literature references concerning the determination of the Kd values of technetium on soil samples taken at different sampling depths done as extensively as in this study was not available. The Kd values determined for soils, observed in the previous studies, ranged from -0.4 ml/g to 61 ml/g, whilst the typical Kd was approximately 0-0.5 ml/g (Abdelouas et al. 2002; Mousny & Myttenaere 1981; Sheppard et al. 1990; Vidal et al. 2009). The Kd values determined in this study are in good agreement with the literature results. It is worth noting that the samples used in the sorption experiments were dried and fractioned and the Kd values for the soil layers were calculated as a weighted average of the Kd values of the studied grain sizes (>0.063 mm, 0.063-0.125 mm, 0.125-0.25 mm, 0.25-0.50 mm, 0.50-1.0 mm and 1.0-2.0 mm). The overall average Kd values for the deep soil pits were 3.42 ml/g, 21.1 ml/g and 1.08 ml/g for OL-KK14, OL-KK15 and OL-KK16, respectively.


The sorption of technetium on mineral soil is known to be quite low and to be mostly or completely reversible ion exchange on soils’ sorption sites (Bunzl & Schimmack 1988; Denys et al. 2003; Koch-Steindl & Pröhl 2001; Lieser & Bauscher 1987; Sheppard et al. 1983). In mineral soils, aluminium and iron sesquioxides may act as sorbents for technetium (Mousny & Myttenaere 1981; Tagami & Uchida 1996). The presence of competing anions hinders the sorption as the competition of the sorption sites increases (Lieser & Bauscher 1987).

The sorption of technetium (TcO4-) on untreated aerobic soil samples of OL-KK20

shows no clear dependence on soil depth, organic matter content or equilibrium pH (Figure 62, Appendix 10). The average Kd value for the soil pit OL-KK20 was 1.32 ml/g, which is slightly higher than the average Kd values for mineral soil layers of OL-KK14, OL-KK15 and OL-KK16; 0.50 ml/g, 0.35 ml/g and 1.00 ml/g, respectively. The reason might be that the soil samples of OL-KK20 were not dried before the sorption experiments, and thus contained more micro-organisms capable of changing the ambient Eh-pH conditions by decomposing organic matter in the presence of limited amounts of oxygen (Abdelouas et al. 2002, 2005; Koch-Steindl & Pröhl 2001). Shift in the redox conditions to more reducing would induce a change in the speciation of technetium from highly mobile and non-sorbing pertechnetate (TcO4

-) to immobile, sparingly soluble Tc(+IV) (Lieser & Bauscher 1987). In dried and pretreated samples the microbe population might be reduced in comparison with the untreated soil samples.

The typical Kd value for technetium on mineral soil has been reported to be approximately 0-0.5 ml/g with the range from -0.4 ml/g to 61 ml/g (Abdelouas et al. 2002; Mousny & Myttenaere 1981; Sheppard et al. 1990; Vidal et al. 2009). The Kd values determined in this study are within these ranges.

95

1,2 1,3 1,4

3,5

3,0

2,5

2,0

1,5

1,0

0,5

So

il d

epth

(m

)

Kd (ml/g)

Figure 62. Kd values of Tc-99 for the aerobic untreated soil samples of OL-KK20. The Kd values were calculated as an average of all the sorption test done to a specific sample depth.

The Kd values determined for dried and fractioned soil samples are not recommended to be used in the migration modelling of technetium because they may not represent the actual conditions in the natural environment. Instead, the Kd values for untreated soil samples may give more relevant results even though the difference between untreated and pretreated samples was only minor.

4.6.3 Anaerobic untreated soil samples of OL-KK20 and OL-KK21

Technetium is a redox sensitive element and in anaerobic, reducing conditions formation of rather immobile, sparingly soluble Tc(+IV) species is dominating (Koch-Steindl & Pröhl 2001; Lieser & Bauscher 1987). Tc(+IV) has high affinity towards surface sorption sites which increases the Kd value of technetium (Lieser & Bauscher 1987). The mobility of technetium in oxygen deficient soils is very low or non-existing (Ashworth & Shaw 2005; Lieser & Bauscher 1987; Tagami & Uchida 1996; Tagami & Uchida 1999). At pH 7, the conversion from TcO4

- to Tc(+IV) takes place when the redox potential is approximately +100 mV, whereas at pH 9 the conversion occurs at lower redox potentials (0 mV) (Takeno 2005).

Figure 63 and Appendix 10 presents the Kd values of technetium on untreated anaerobic soil samples of OL-KK20 and OL-KK21. Sorption on the samples of OL-KK21 clearly increases with soil depth, whereas on OL-KK20 the sorption shows only slight variation. The retention of technetium on OL-KK21 samples was higher than on the samples of OL-KK20, excluding sample depth 1.1 m. Reason for the higher Kd values for OL-KK21 are probably the lower measured redox potentials compared with OL-KK20; +46.28-133.66 RmV versus +202.48-366.48 RmV. The highest Kd value determined for OL-KK21 was seen for the lowest sample depth 3.6 m (568 ml/g), whereas sample depth 1.1 m had the lowest value (1.05 ml/g). In OL-KK20, the Kd values remained almost constant ranging from 0.86 ml/g to 3.04 ml/g. The average Kd

96

value for the soil pits OL-KK20 and OL-KK21 were 1.96 ml/g and 93.6 ml/g, whereas the respective average Eh values were +253.6 RmV and +93.6 RmV. It is probable that the reduction on pertechnetate to Tc(+IV) took place in the more reducing samples leading to higher Kd values. In this experiment series highly reducing conditions were not met.

1 10 1004,0

3,5

3,0

2,5

2,0

1,5

1,0

0,5

0,0 OL-KK20 OL-KK21

So

il d

ep

th (

m)

Kd (ml/g)

Figure 63. Kd values of Tc-99 for the anaerobic untreated soil samples of OL-KK20 and OL-KK21. The Kd values were calculated as an average of all the sorption test done to a specific sample depth.

For OL-KK20, the difference in the sorption behaviour of technetium between the aerobic and anaerobic samples was practically nonexisting, the average Kd values being 1.32 ml/g and 1.96 ml/g, respectively. This is probably due to the rather high Eh value of +253 RmV in the anaerobic conditions. However, for OL-KK21 the Kd values determined for the anaerobic samples were considerably higher compared with the values of the aerobic samples indicating the substantial role of the redox conditions and speciation on the sorption of technetium. The redox conditions and its effect on the speciation of technetium in soils overwhelm the effect of organic matter in the retention and sorption processes.

In anaerobic soil conditions the Kd values of technetium has reported being in the range of 0.27-440 ml/g (Sheppard et al. 1990). The results of this study are in a good agreement with the literature values, even though some further research is recommendable to carry out, especially in the anaerobic conditions.


Organic matter is known as the main sorbent for oxidised and reduced technetium species in soils (Ashworth & Shaw 2005; Bunzl & Schimmack 1988; Koch-Steindl & Pröhl 2001; Mousny & Myttenaere 1981; Sheppard et al. 1983; Sheppard et al. 1990).

97

The sorption reaction is reversible and not affected by pH in the range of 1.6-5.4 (Wolfrum & Bunzl 1986).

The sorption of technetium (TcO4-) on humus sample 1 increased with increasing

equilibrium time; from 3.54 ml/g on day seven to 366 ml/g on day 31 (Figure 64, Appendix 10). Equilibrium pH remained from neutral (7.0) to slightly alkaline (7.7). It is known that aerobic micro-organisms are not able to reduce or biosorb TcO4

- (Bennett & Willey 2003), so the behaviour for untreated humus samples is due to changes in other properties. It is possible that micro-organisms have induced changes e.g. in the redox conditions since TcO4

- does not form complexes with humic substances, and complexation occurs only when TcO4

- is reduced to Tc(+IV) (Bennett & Willey 2003; Ticknor et al. 1996). The reduction reaction increases the sorption of technetium on humic substances, and also enhances the association of Tc with solid organic matter, which would explain the rather high Kd values were observed with time.

0 20 40 60 80 1001

10

100

1000

Kd (

ml/g

)

Time (days)

Figure 64. Kd values of Tc-99 for the untreated humus sample 1 of OL-KK21 as a function of time.

The Kd values for the dried and fractioned humus samples of OL-KK14, OL-KK15 and OL-KK16 were considerably smaller than the values for OL-KK21, namely 15.1 ml/g, 126 ml/g and 1.44 ml/g compared with 306 ml/g. Reason for the low sorption on dried and fractioned samples may be caused by a decrease in the microbe population during the sample drying. The Kd values determined for organic soils by other researchers have been in the range of 1.5-41 ml/g (Kaplan 2003; Vidal et al. 2009; Wolfrum & Bunzl 1986). The values in this study are substantially higher in comparison with the literature values, which may be e.g. due to longer equilibrium times used in this experiment series.


The sorption of technetium on untreated irradiation sterilized humus samples 1, 2 and 3 of OL-KK21 increase with equilibrium time (Appendix 10), but do not extend to the same level with nonsterilized soil samples. The Kd values for irradiated samples

98

increased from 2.89 ml/g on day seven to 9.50 ml/g on day 93. Equilibrium pH ranged from 4.80 to 5.68. The Kd values determined for irradiated humus samples resemble the values determined for the dried and fractioned humus samples, for which the Kd values ranged from 1.44 ml/g to 126 ml/g. For untreated humus samples the average Kd was 306 ml/g. These results support the role of micro-organism in the retention of technetium on soil, and corroborate the hypothesis that soil micro-organisms induce a decrease in the soil redox potential upon which TcO4

- is reduced to Tc(+IV) and associated with dissolved humic substances and solid organic matter. From the literature, the Kd values for irradiated peat were approximately 7-20 ml/g (Bunzl & Schimmack 1988).

4.6.6 Recommendations of the Kd values for technetium

Figure 65 presents the Tc-99 Kd values as a function of soil depth on humus and mineral soil samples. The data set for the humus samples comprised of the Kd values determined for OL-KK14, OL-KK15, OL-KK16 and OL-K21. The Kd values determined for irradiation sterilized samples were not included in the humus data set because microbial activity is typically ubiquitous in all type of soils soils. The data set for the aerobic mineral soil consisted of the Kd values determined for the aerobic soil samples of OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21. The derivation of the recommended conservative and realistic Kd values for Tc-99 in the anaerobic conditions was based on the data set of the Kd values determined for the anaerobic OL-KK20 and OL-KK21 mineral soil samples (Figure 66). The range of the determined Kd values for humus and mineral soil is given in Table 19, which also presents the recommended conservative and realistic Kd values for technetium in the aerobic and anaerobic conditions.

0,01 0,1 1 10 100

350

300

250

200

150

100

50

0


So

il d

ep

th (

cm

)

Kd (ml/g)

Figure 65. Kd values of Tc-99 on the aerobic soil samples of OL-KK14, OL-KK15, OL-KK16, OL-KK20 and OL-KK21 as a function of soil depth used in the estimation of the recommended conservative and realistic Kd values.

99

0,1 1 10 100 1000

350

300

250

200

150

100

50

0 Soil depth: 30-100 cm 100-200 cm >200 cm

So

il d

epth

(cm

)

Kd (ml/g)

Figure 66. Kd values of Tc-99 on the anaerobic soil samples of OL-KK20 and OL-KK21 as a function of soil depth used in the estimation of the recommended conservative and realistic Kd values.

The conservative Kd values for humus, aerobic and anaerobic mineral soil was 0.0 ml/g. The estimated realistic technetium Kd value was rather high for humus, 21 ml/g, whereas the corresponding value for aerobic mineral soil ranged from 0.4 ml/g to 0.7 ml/g. The realistic Kd values calculated for anaerobic mineral soil were about ten times higher than for the aerobic mineral soil showing no clear dependence on soil depth.

Previously Vidal et al. (2009) estimated that the geometric mean Kd value of technetium on organic soil was 3.1 ml/g, whereas the corresponding values for sand, loam and clay were 0.04 ml/g, 0.07 ml/g, and 0.09 ml/g, respectively. On the other hand, Sheppard et al. (2009b) estimated the overall geometric mean Kd value for technetium on aerated soil to be 3.1 ml/g, whereas the respective value for anaerobic soils was 65 ml/g. The recommended realistic Kd values for aerobic mineral soils presented in this report are within the same range with the values estimated by Vidal et al (2009), but remain smaller than the 3.1 ml/g assessed by Sheppard et al (2009b). The geometric mean Kd value of 21 ml/g for humus is higher than the value (3.1ml/g) proposed by Vidal et al. (2009). Furthermore, the recommended Kd values for technetium in the anaerobic soil conditions are smaller than the value of 65 ml/g estimated by Sheppard et al (2009b).

100

Table 19. The range of the Kd values determined for Tc-99 in the sorption tests and recommended conservative and realistic Kd values.

Aerobic soil Soil depth

(cm) Soil type

Kd range (ml/g)

N Conservative


humus 0.0 - 578 45 0.0 21

5-30 top mineral

soil 0.0 - 2.41 15 0.0 0.4


0.0 - 6.29 62 0.0 0.6


soil 0.0 - 3.01 43 0.0 0.6

>200 deep mineral

soil 0.0 - 8.97 60 0.0 0.7

Anaerobic soil Soil depth

(cm) Soil type

Kd range (ml/g)

N Conservative



0.0 - 93 14 0.0 9.3


soil 0.0 - 145 23 0.0 3.0

>200 deep mineral

soil 0.0 - 3680 41 0.0 7.9

101

5 SUMMARY

Biosphere (dose) assessment programme for the spent nuclear fuel is focused on Cl-36, I-129, C-14, Mo-93, Nb-94, Cs-135, Ni-59, Se-79 and Sr-90 (Hjerpe et al. 2010, p. 36-38; Haapanen et al. 2009, p. 22-23).

The source for I-129 and Tc-99 in the spent nuclear fuel is the fission of U-235 and Pu-239 atoms. Cl-36, Nb-94 and Se-79 are long-lived activation products formed from the stable isotopes present in the nuclear fuel, structural components and construction materials used in the power reactors. Nb-94 and Se-79 are also fission products. Knowledge of the behaviour of these nuclides in the environment is of notable interest because of their long physical half-lives and anionic nature. Cs-135 is also a fission product, which exists solely in the cationic form. The interest in Cs-135 arises from its long physical half-life.

The sorption of radionuclides in soils is affected by parameters specific to the element and to the soil. Chemical form, speciation, is the most important elemental factor affecting the sorption and migration properties of the element. Soil redox potential, pH and complex forming ligands are features that have a great influence on the speciation. Micro-organisms can affect the speciation of radionuclides indirectly by changing the prevailing Eh-pH conditions and also serve as sorbents. Organic matter content and soil mineral composition have a noticeable influence on the retention of radionuclides. Radionuclides having a high tendency of forming anionic species in aqueous solutions, such as chlorine, iodine, selenium and technetium, are poorly sorbed on soil mineral constituents and thus have a high potential mobility in the overburden. The sink for anionic radionuclides, such as iodine and selenium, is most likely the organic matter. Cationic nuclides may be effectively associated with the soil mineral matter, depending on the mineral composition.

Distribution coefficient, Kd, is used to describe the retention of a substance on the solid phase. Kd value is defined as the concentration ratio of the radionuclide between the solid and liquid phases in the equilibrium state. High Kd values indicate the considerable retention of radionuclide on soil or sediment, whereas for low Kd values the retention is small. Kd is widely used as an indicator of the mobility of radionuclides in the environment.

The sorption of caesium, chlorine, iodine, niobium, selenium and technetium was studied on aerobic and anaerobic Olkiluoto soil samples in experimental batch sorption tests to gain site specific knowledge on their retention.

Soil chemical and geotechnical characteristics

The chemical and geotechnical analyses for soil samples of OL-KK20 and OL-KK21 were carried out. The main results are:

The pH and dry matter content increased as a function of depth, whereas organic matter content decreased.

The main minerals in the mineral soil samples were quartz, plagioclase, potassium feldspar and kaolinite in the grain size fractions <0.01 mm and 1.0-2.0 mm. Small amounts of micas, amphibole, hematite and chlorite were found.

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Cation exchange capacity showed significant variation among the sampling depths (due to the different soil types). CECs ranged from 63 mmol/kg to 287 mmol/kg in OL-KK20. The respective variation in OL-KK21 was 64-218 mmol/kg.

The specific surface area was the highest in the smallest grain size fraction <0.063 mm for OL-KK14, ranging from 2.1 m2/g to 9.0 m2/g. The specific surface for the mineral soil layers decreased in the order MS2>MS3>MS4>MS1. In OL-KK20 and OL-KK21 the specific surface area decreased with depth. The range in OL-KK20 was 0.8-6.8 m2/g and 0.9-4.0 m2/g in OL-KK21.

Caesium

Caesium was retained efficiently on the mineral soil samples. For the soil pits OL-KK14, OL-KK15 and OL-KK16 the Kd values decreased as a function of depth corresponding to a decrease in the clay fraction content and CEC. Kd for the topmost mineral soil layer MS1 was smaller than for the underlying MS2 layer in OL-KK14 and OL-KK15. For the mineral soil layers, the Kd values ranged approximately between 3 000 ml/g and 33 000 ml/g. The retention was the smallest in the humus layers, where the Kd value range was 200-12 000 ml/g.

The sorption of caesium on the aerobic untreated soil samples of OL-KK20 and OL-KK21 was smaller than for the dried and fractioned mineral soil samples of OL-KK14, OL-KK15 and OL-KK16. The mineral soil samples of OL-KK21 retained caesium better than the samples of OL-KK20. The Kd values varied between 1 300 ml/g and 7 300 ml/g. Sorption in the anaerobic conditions was reduced probably due to an increase in the NH4

+ concentration and thus pronounced competition of the sorption sites. The Kd values decreased from 1 300-5 500 ml/g in the aerobic conditions to 650-3 300 ml/g in the anaerobic conditions.

Humus was not sorbing caesium efficiently. Sorption decreased with time from 6 600 ml/g on day seven to 1 800 ml/g on day 23. The reason for this could be a possible decrease in the microbe population due to decline in the oxygen and nutrient concentration with increasing equilibrium time.

The effect of competing cations decreased in the order Cs+>NH4+>K+>Ca2+>Na+. In the

concentration of 10 mM stable Cs+, the Kd values decreased to <1 ml/g, whereas in 1000 mM Na+ solutions the Kd values were still >50 ml/g indicating very weak competition between caesium and sodium of the soil sorption sites. The effect of ammonium and potassium ions were similar and resulted in smaller Kd values compared with the soil solution simulant even in 0.1 mM K+ or NH4

+ concentration. The effect of calcium was similar to that of sodium.

Chlorine

Retention of chloride anion on mineral soil samples did not take place; the Kd values were approximately 0 ml/g. Sorption on the untreated humus samples was minor, and the sorption slightly decreased with increasing equilibrium time. The Kd values ranged from 0 ml/g to 1.70 ml/g. Irradiation sterilisation did not have a marked effect on the retention of chlorine on humus. On dried and fractioned humus samples of OL-KK15,

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the sorption increased with time and the highest Kd value 61 ml/g was determined for the smallest grain size fraction <0.063 mm.

Iodine

The Kd values of iodine for dried and fractioned humus samples of OL-KK14 and OL-KK15 were rather high, 604 ml/g and 2670 ml/g, respectively. For OL-KK16 the Kd was slightly lower, approximately 78 ml/g. The retention of iodine decreased in the mineral soil layers as a function of depth with decreasing organic matter content and increasing equilibrium pH.

The sorption of iodine was higher on untreated mineral soil samples of OL-KK20 and OL-KK21 than on dried and fractioned soil samples of OL-KK14, OL-KK15 and OL-KK16. One of the reasons may be the decrease in the microbe population for the fractioned samples caused by sample pretreating (drying). The Kd values for OL-KK20 ranged from 14.2 ml/g to 26.4 ml/g, while the respective variation for OL-KK21 was 5.90-136 ml/g. Iodine did not sorb on the anaerobic mineral soil samples and the Kd values remained practically at 0 ml/g. Probably the speciation of iodine as poorly retaining iodide (I-) or the negative surface charge of the mineral soil induced by the high equilibrium pH resulted in the low sorption.

Untreated humus samples of OL-KK21 retained iodine quite well even though sorption on dried and fractioned humus samples was higher. The Kd values declined as a function of equilibrium time from 475 ml/g on day five to 9.6 ml/g on day 93. It is possible that the decrease in the sorption was due to the decay in the microbe population induced by the consumption of oxygen and nutrients. This is supported by the fact that the sorption of iodine on irradiation sterilised humus samples was comparably smaller; the Kd values were only 2.50-10.40 ml/g on day seven whereas for nonsterilized humus sample the respective value was 244 ml/g. It is also possible that irradiation has induced some changes in the sample composition, for example increase the solubility of the fulvic/humic acids.

Niobium

Among the studied elements, niobium retained on the mineral soil samples the most efficiently. Sorption experiments in the aerobic conditions with untreated mineral soil samples of OL-KK20 was conducted at two different pH value ranges. At the lower pH range of 3.2 to 4.5, the Kd varied randomly with respect to depth between 15 000 ml/g and 127 000 ml/g. At the higher pH values between 6.8 and 7.9, the Kd values were in the range 27 000-154 000 ml/g. The highest values were found for the uppermost sampling depth of 0.70 m. The Kd values determined from the filtered samples were higher than the values for centrifuged samples, which was due to the sorption of Nb-95 on colloidal material removed by filtering. The fraction of direct filter sorption was notably smaller. The Kd values for centrifuged and filtered samples were 5 000-20 000 ml/g and 27 000-154 000 ml/g, respectively.

Sorption of niobium on anaerobic mineral soil samples of OL-KK20 was studied at the pH range of 5.3-6.1. The Kd value in the uppermost soil depth (0.7 m) was approximately at the same level as in the aerobic conditions (48 000 ml/g), but at the

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deeper layers the difference in the Kd values was approximately one order of magnitude lower.

The retention of niobium on untreated humus samples of OL-KK21 was smaller than on the mineral soil samples. The highest Kd value was approximately 1 800 ml/g whilst the typical Kd was below 1 000 ml/g and somewhat increased with time.

Selenium

The Kd values for selenium on the aerobic soil samples of OL-KK20 had no clear dependence on soil depth. The highest Kd for the centrifuged samples was 244 ml/g at the sample depth 0.70 m, whereas the lowest was 9.50 ml/g at 3.00 m depth. In the anaerobic conditions the sorption of selenium decreased with soil depth from 98 ml/g at 0.70 m to approximately 5 ml/g at 2.05 m and 3.00 m depths. In the anaerobic samples the sorption of selenium decreased with time.

Organic matter is the main sorbent for selenium in soils which is indicated by the elevated Kd values for humus compared with the mineral soil samples. Retention increased with time, for example from 67 ml/g on day one to 1 400 ml/g on day 63 for the humus sample 1 of OL-KK21.

Technetium

The retention of technetium is the highest on soil layers enriched with organic matter. In the soil layers of OL-KK14, OL-KK15 and OL-KK16, the Kd values decreased from 15.1 ml/g, 126 ml/g and 1.44 ml/g in the humus layer to <1 ml/g in the mineral soil layers. In the aerobic mineral soil samples of OL-KK20 the Kd values were approximately 1.3 ml/g.

In the anaerobic conditions the sorption of technetium increased with decreasing redox potential due to the reduction of highly mobile pertechnetate anion (TcO4

-) to sparingly soluble Tc(+IV) species. This was seen as relatively high Kd values for mineral soil samples of OL-KK21, up to 570 ml/g. There was no clear dependency with the sorption and soil depth, but typically the Kd values were higher for the soil samples taken from the soil pit’s deepest parts (2.40-3.60 m). For the soil pit OL-KK20, the redox conditions remained quite oxidising and, thus low Kd values, around 2.0 ml/g, were observed.

The retention of technetium increased as a function of time on untreated humus samples of OL-KK21; Kd value on day seven was 3.54 ml/g and 547 ml/g on day 93. The highest value (547 ml/g) is comparable to the highest values in the anaerobic conditions (570 ml/g). Irradiation sterilisation decreased considerably the retention of technetium on humus, for example, the Kd value for the humus sample 3 was only 7.40 ml/g on day 93. Thus, the effect of micro-organisms may be crucial to soil redox conditions and the retention of technetium on soils.

Final conclusions

Caesium and niobium are retained very efficiently on mineral soil regardless of the prevailing redox conditions and thus their migration to overlaying biosphere is greatly hindered.

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Technetium, and possibly also selenium are retained very well and immobilised in the reducing soil regime due to the formation of reduced, sparingly soluble species Tc(+IV) and Se(0,-II). Once oxidising conditions come prevailing, Tc(+IV) and Se(0,-II) are oxidised to more mobile forms TcO4

- and SeO32- and SeO4

2- and migrated upwards within the soil profile by diffusion and with the groundwater flow. In the uppermost parts of the soil profiles technetium and selenium are retained in the soil layers enriched with organic matter.

Iodine is very poorly sorbing in anaerobic, reducing mineral soil conditions where the prevailing species is iodide (I-). Iodine continues the upward migration until a soil layer enriched with organic matter is encountered. In this layer, iodine is retained and its further migration is delayed in some extent.

Chlorine is sorbed poorly on mineral and humus soil layers. Thus, the migration of Cl-36 is not supposed to be inhibited at any time and it is freely distributed into the biosphere.

Based on the results presented in this work, the most important radionuclides in the safety assessment of spent nuclear fuel are Cl-36 and I-129 irrespective the soil conditions. Se-79 and Tc-99 are considered less mobile and important as far as the conditions remain reducing. Upon oxidation, these elements become mobile and their importance in the safety assessment increases. Caesium and niobium are considered the least important of the studied elements due to their effective sorption on mineral soil. Table 20 summarises the processes taking place in soil and their effect on the studied radionuclides.

Table 20. The effect of processes taking place in soil on Cs-135, Cl-36, I-125, Nb-94, Se-75 and Tc-99.

Process Cs-135 Cl-36 I-125 Nb-94 Se-75 Tc-99 Retention on organic matter + + + + + + Retention on clay minerals/

clay soils + ± + ± ± ± Retention on weakly

crystalline Al and Fe oxides ± ± + ± Retention by carbonates ± + + Retention by reducing minerals (e.g. pyrite) ± ± ± Retention by micro-

organisms ± ± + ± ± + Incorporation into organic

substances (by micro-organisms)

± ± ± ± ±

High pH + ± - - - - Formation of precipitates ±

Volatilization ± ± Decrease in the redox potential (aerobic →

anaerobic) - - - +

+

+ process seen in Olkiluoto samples; correlation with retention is positive - process seen in Olkiluoto samples; correlation with retention is negative ± process not seen/verified in Olkiluoto samples positive correlation with retention in the literature negative correlation with retention in the literature indirect negative correlation with retention in the literature process is not important/does not take place (enough) information not available in the literature

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4 91

.8

0.4

94.8

0.

123

9.6

39.0

-

-

3.0

0.4

92.0

0.

4 93

.5

0.4

1.1

87.5

1.

3 89

.7

3.4

0.5

92.0

0.

4 94

.9

1.1

0.7

90.1

0.

9 93

.3

- -

- -

- 1.

75

0.5

91.0

0.

6 92

.9

- -

- -

- 2.

4 0.

4 91

.7

0.5

93.3

- -

- -

- 2.

95

0.4

90.5

0.

5 94

.6

- -

- -

- 3.

6 0.

3 89

.9

0.4

91.6

0.12

1 hum

us s

ampl

e 1

of th

e ex

cava

tor

pit O

L-K

K21

0.

122 h

umus

sam

ple

2 of

the

exca

vato

r pi

t OL

-KK

21

0.12

3 hum

us s

ampl

e 3

of th

e ex

cava

tor

pit O

L-K

K21

-

sam

ple

not t

aken

124

125

AP

PE

ND

IX 3

. OL

-KK

20 a

nd

OL

-KK

21 m

iner

alo

gy

App

endi

x 3.

1. M

iner

alog

y of

the

aero

bic

soil

sam

ples

of

the

exca

vato

r pi

t OL

-KK

20.

Gra

in s

ize

<0.

01 m

m

soil

dep

th (

m)

Min

eral

s Q

uart

z (%

) P

otas

sium

fe

ldsp

ar (

%)

Pla

gioc

lase

(%

) K

aoli

nite

(%

) M

ica

(%)

Am

phib

ole

(%)

Hem

atit

e (%

)

0.7

45

10

30

10

<5

+ (+

) 1.

3**

50

15

25

10

<5

+ (+

)2.

05

45

10

30

10

<5

+ (+

)

3.0

35

25

30

10

<5

+ (+

)3.

4*

50

10

25

10

<5

+ (+

)G

rain

siz

e 1.

0-2.

0 m

m

soil

dep

th (

m)

Min

eral

s Q

uart

z (%

) P

otas

sium

fe

ldsp

ar (

%)

Pla

gioc

lase

(%

) K

aoli

nite

(%

) M

ica

(%)

Am

phib

ole

(%)

Hem

atit

e (%

)

0.7

60

10

15

10

<5

+ (+

)

1.3

60

15

25

<5

<5

‐ ‐

2.05

50

20

25

5

<5

‐ ‐

3.0

45

25

20

10

<5

(+)

(+)

3.4

45

25

20

10

<5

‐ ‐

* sa

mpl

e he

ated

in 5

50 °

C f

or o

ne h

our

** s

ampl

e he

ated

in 5

50 °

C f

or o

ne h

our

and

IR s

pect

rum

was

mea

sure

d

(+)

poss

ibly

125

126

AP

PE

ND

IX 3

. OL

-KK

20 a

nd

OL

-KK

21 m

iner

alo

gy

App

endi

x 3.

2. M

iner

alog

y of

the

aero

bic

soil

sam

ples

of

the

exca

vato

r pi

t OL

-KK

21.

Gra

in s

ize

<0.

01 m

m

soil

dep

th

(m)

Min

eral

s

Qu

artz

(%

)

Pot

assi

um

fe

ldsp

ar

(%)

Pla

gioc

lase

(%

) K

aoli

nite

(%

) M

ica

(%)

Am

ph

ibol

e (%

) H

emat

ite

(%)

Ch

lori

te

(%)

0.4

45

15

30

10

<5

(+)

(+)

- 1.

1*

45

15

30

10

<5

+

(+)

- 1.

75

45

15

25

10

5 +

(+

) -

2.4

50

10

30

5 <

5 +

(+

) -

2.95

**

50

10

30

5 <

5 +

(+

) +

3.

6 50

15

25

5

<5

+

(+)

- G

rain

siz

e <

0.01

mm

soil

dep

th

(m)

Min

eral

s

Qu

artz

(%

)

Pot

assi

um

fe

ldsp

ar

(%)

Pla

gioc

lase

(%

) K

aoli

nite

(%

) M

ica

(%)

Am

ph

ibol

e (%

) H

emat

ite

(%)

Ch

lori

te

(%)

0.4

55

20

20

5 <

5 -

- -

1.1

45

25

20

10

<5

- -

- 1.

75

55

20

25

- -

- -

- 2.

4 55

15

25

5

<5

- -

- 2.

95

55

15

25

5 <

5 -

- -

3.6

65

10

15

5 <

5 -

- -

* sa

mpl

e he

ated

in 5

50°C

for

one

hou

r **

sam

ple

heat

ed in

550°C

for

one

hou

r an

d IR

spe

ctru

m w

as m

easu

red

(+)

poss

ibly

126

127

APPENDIX 4. OL-KK14, OL-KK20 and OL-KK21 specific surface areas Appendix 4.1. Specific surface areas of soil samples of the excavator pit OL-KK14. Soil layer Grain size (mm) Specific surface area (m2/g) MS1 <0.063 2.1 0.063-0.125 1.0 0.125-0.25 1.1 0.25-0.50 0.9 0.50-1.0 0.7 1.0-2.0 0.6 MS2 <0.063 9.0 0.063-0.125 4.6 0.125-0.25 6.9 0.25-0.50 7.2 0.50-1.0 6.8 1.0-2.0 5.6 MS3 <0.063 7.2 0.063-0.125 3.5 0.125-0.25 3.2 0.25-0.50 3.8 0.50-1.0 3.7 1.0-2.0 3.4 MS4 <0.063 5.2 0.063-0.125 2.6 0.125-0.25 2.3 0.25-0.50 2.9 0.50-1.0 3.9 1.0-2.0 4.1

127

128

APPENDIX 4. OL-KK14, OL-KK20 and OL-KK21 specific surface areas

Appendix 4.2. Specific surface areas of soil samples of the excavator pits OL-KK20 and OL-KK21.

OL-KK20 OL-KK21

Soil depth (m) Specific surface

area (m2/g) Soil depth (m)

Specific surface area (m2/g)

0.7 6.8 0.4 3.5 1.3 2.5 1.1 4.0 2.05 1.4 1.75 2.7 3.0 1.0 2.4 1.0 3.4 0.8 2.95 1.0 - - 3.6 0.9

- sample not taken

128

129

AP

PE

ND

IX 5

. Kd v

alu

es o

f ca

esiu

m

App

endi

x 5.

1. K

d va

lues

of

caes

ium

for

the

soil

laye

rs o

f th

e ex

cava

tor

pits

OL

-KK

14, O

L-K

K15

and

OL

-KK

16.

Soi

l la

yer

OL

-KK

14

OL

-KK

15

OL

-KK

16

sam

ple

dept

h (c

m)

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

sam

ple

dept

h (c

m)

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

sam

ple

dept

h (c

m)

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

hum

us

5 11

759

10

355

4.

7 7

198

152

4.5

10

964

835

5.6

MS

1 20

3

041

369

5.2

50

982

153

6.8

30

11 2

28

733

7.1

MS

2 60

32

787

4

650

6.5

80

18 9

49

2162

7.

3 50

8

906

121

7.3

MS

3 10

5 13

969

2

109

6.9

110

7 86

4 09

5 7.

9 11

0 6

942

679

8.0

MS

4 24

0 4

192

575

7.5

160

5 39

6 82

7 8.

0 30

0 3

157

394

8.3

MS

5 -

- -

- 26

0 2

936

437

8.6

- -

- -

- sa

mpl

e no

t tak

en

129

130

APPENDIX 5. Kd values of caesium

Appendix 5.2. Kd values of caesium for the different grain sizes of the soil layers of the excavator pit OL-KK15. Soil layer Grain size

(mm) Kd (ml/g) Standard deviation

(ml/g) pH

humus <0.063 584 450 4.6 0.063-0.125 127 3.5 4.6 0.125-0.25 124 1.6 4.4 0.25-0.50 112 3.8 4.5 0.50-1.0 126 7.4 4.5 1.0-2.0 115 1.3 4.5

MS1 <0.063 6 937 3 098 6.5 0.063-0.125 1 832 199 6.6 0.125-0.25 1 123 161 6.6 0.25-0.50 1 002 51 6.9 0.50-1.0 722 47 7.0 1.0-2.0 610 53 7.1

MS2 <0.063 18 408 528 7.3 0.063-0.125 11 058 391 7.2 0.125-0.25 19 767 53 7.3 0.25-0.50 20 457 1 219 7.3 0.50-1.0 18 823 1 028 7.3 1.0-2.0 17 898 311 7.5

MS3 <0.063 11 264 186 7.9 0.063-0.125 5 955 261 7.8 0.125-0.25 5 695 63 7.9 0.25-0.50 6 343 197 7.9 0.50-1.0 7 853 282 8.0 1.0-2.0 6 412 605 8.0

MS4 <0.063 7 057 195 8.0 0.063-0.125 4 443 38 8.0 0.125-0.25 4 783 277 7.9 0.25-0.50 4 956 181 8.0 0.50-1.0 5 095 265 8.0 1.0-2.0 4 835 259 8.0

MS5 <0.063 4 060 71 8.1 0.063-0.125 2 157 81 8.4 0.125-0.25 1 780 75 8.7 0.25-0.50 1 622 21 8.7 0.50-1.0 2 790 172 8.8 1.0-2.0 6 093 659 8.8

131

AP

PE

ND

IX 5

. Kd v

alu

es o

f ca

esiu

m

App

endi

x 5.

3. K

d va

lues

of

caes

ium

for

the

exca

vato

r pi

ts O

L-K

K20

aer

obic

and

ana

erob

ic a

nd O

L-K

K21

aer

obic

soi

l sam

ples

.

Soi

l de

pth

(m)

OL

-KK

20

OL

-KK

21

aero

bic

anae

robi

c ae

robi

c

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

Eh

(Rm

V)

Soi

l de

pth

(m)

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

0.7

5 45

6 7.

6 7.

6 3

319

3 14

6 8.

6 34

3 0.

4 7

256

7 20

0 7.

3 1.

3 2

661

8.1

8.1

2 09

3 97

8 9.

2 60

1.

1 5

327

5 21

1 7.

6 2.

05

1 68

0 8.

1 8.

1 1

038

016

9.5

22

1.75

3

666

3 66

0 8.

0 3.

0 19

17

8.1

8.1

1 04

7 02

6 9.

6 30

2.

4 3

559

2 62

3 7.

9 3.

4 1

315

8.2

8.2

649

620

9.6

29

2.95

2

400

2 33

0 7.

8 -

- -

- -

- -

- 3.

6 2

742

2 13

2 7.

9 -

sam

ple

not t

aken

131

132

AP

PE

ND

IX 5

. Kd v

alu

es o

f ca

esiu

m

App

endi

x 5.

4. K

d va

lues

of

caes

ium

for

the

exca

vato

r pi

t OL

-KK

20 a

erob

ic a

nd a

naer

obic

soi

l sam

ples

. O

L-K

K20

Aer

obic

Tim

e (d

ays)

1

3 5

7

Soi

l de

pth

(m)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

0.7

3 42

2 39

9 7.

0 4

114

623

8.6

4 88

6 62

4 7.

7 4

184

1 42

9 7.

5 1.

3 1

709

7.5

7.6

2 25

2 28

4 9.

1 2

201

29

7.9

2 11

5 92

4 7.

4 2.

05

902

28

7.7

1 31

7 32

9.

1 1

154

13

7.9

1 52

6 60

0 7.

7 3.

0 81

0 5.

6 7.

6 1

234

29

9.2

1 59

0 20

8 7.

8 72

8 78

8.

1 3.

4 66

3 74

7.

6 92

0 30

9.

1 1

101

45

7.9

580

152

8.3

OL

-KK

20 A

erob

icO

L-K

K20

An

aero

bic

Tim

e (d

ays)

21

42

24

5 1

Soi

l de

pth

(m)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Kd

erro

r 2σ

(m

l/g)

pH

K

d (m

l/g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Eh

(Rm

V)

0.7

6 23

5 31

21

5.8

5 06

6 38

0 5.

6 4

891

187

5.7

2 42

9 1

069

8.9

360

1.3

2 21

4 65

5.

9 2

803

310

6.4

2 74

3 1

730

6.1

1 88

7 23

2 9.

5 21

6 2.

05

1 56

9 36

1 5.

9 1

728

159

6.1

1 85

1 35

2 6.

1 78

1 32

7 9.

7 17

3 3.

0 1

732

368

5.9

1 85

2 57

8 6.

1 3

207

2 38

7 6.

1 82

7 43

9.

8 19

3 3.

4 1

119

239

5.9

1 20

4 18

9 6.

5 2

136

1 65

1 6.

1 35

4 6.

0 9.

8 17

8 O

L-K

K20

Ana

erob

icT

ime

(day

s)

3 5

7 28

Soi

l de

pth

(m)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g))

pH

Eh

(Rm

V)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Eh

(Rm

V)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Eh

(Rm

V)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Eh

(Rm

V)

0.7

2 36

8 97

8.

7 38

7 3

135

555

8.7

289

3 78

1 99

8 8.

5 33

7 4

884

248

8.4

342

1.3

1 19

1 50

3 9.

3 14

9 2

535

754

9.2

-36

2 09

4 23

0 9.

1 -3

1 2

759

435

8.9

1.5

2.05

1

051

79

9.6

52

1 07

5 81

9.

6 -4

9 1

024

52

9.4

-55

1 25

8 4.

4 9.

2 -1

1 3.

0 99

1 33

9.

6 48

95

0 15

8 9.

5 -4

2 1

070

212

9.5

-47

1 39

9 20

1 9.

3 -3

.5

3.4

727

26

9.7

46

604

92

9.7

-39

744

84

9.6

-43

817

210

9.4

3.5

132

133

AP

PE

ND

IX 5

. Kd v

alu

es o

f ca

esiu

m

App

endi

x 5.

5. K

d va

lues

of

caes

ium

for

the

exca

vato

r pi

t OL

-KK

21 a

erob

ic s

oil s

ampl

es.

Tim

e (d

ays)

1

3 5

7

Soi

l de

pth

(m)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

0.4

6392

1

339

7.0

6 95

8 1

183

7.6

8 16

0 90

7.

6 7

166

806

7.25

1.

1 34

13

17

7.3

3 97

9 29

3 8.

3 4

495

448

7.8

5 04

2 36

0 7.

48

1.75

23

90

56

7.7

2 89

4 25

8.

9 3

723

380

8.0

3 89

0 31

7 7.

73

2.4

1160

92

7.

5 1

832

10

8.4

2 85

9 87

9 7.

9 1

902

213

7.69

2.

95

1300

54

7.

5 1

879

210

7.9

1 88

5 22

4 7.

9 2

104

323

7.69

3.

6 73

8 13

7.

5 1

477

60

8.0

2 35

3 53

7 8.

0 1

453

168

7.61

T

ime

(day

s)

21

42

112

AV

ER

AG

E

Soi

l de

pth

(m)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

0.4

6944

97

6 7.

4 6

940

1 12

6 7.

3 8

230

638

6.9

7 25

6 7

200

7.28

1.

1 46

09

778

7.7

5 70

3 24

8 7.

5 10

047

1

111

7.3

5 32

7 5

211

7.63

1.

75

4071

17

7 7.

8 3

679

569

7.8

5 01

4 2

311

7.8

3 66

6 3

660

7.97

2.

4 26

99

153

7.8

3 02

3 1

396

7.8

11 4

33

12 5

02

8.0

3 55

8 2

623

7.87

2.

95

2556

71

7.

8 2

564

654

7.8

4 51

4 47

4 7.

8 2

400

2 33

0 7.

78

3.6

2063

15

4 7.

9 3

075

567

7.9

8 03

7 6

802

8.1

2 74

2 2

132

7.85

133

134

AP

PE

ND

IX 5

. Kd v

alu

es o

f ca

esiu

m

App

endi

x 5.

6. K

d va

lues

of

caes

ium

for

the

exca

vato

r pi

t OL

-KK

20 a

erob

ic s

oil s

ampl

es in

the

pres

ence

of

stab

le C

s, N

a or

K c

ompe

ting

ca

tion

. C

sCl

conc

entr

atio

n (m

M)

0.01

0.

1 1

10

Soi

l sol

utio

n si

mul

ant

Soi

l dep

th

(m)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

0.7

215

0.9

7.2

72.1

6.

8 7.

1 15

.6

7.2

7.1

1.05

0.

4 7.

9 4

184

1 42

9 7.

46

1.3

53.3

0.

9 7.

6 15

.6

3.3

7.6

5.27

0.

6 7.

5 0.

79

0.5

7.5

2 11

5 92

4 7.

41

2.05

46

.7

13

7.6

9.51

1.

2 7.

6 3.

52

0.5

7.6

0.38

0.

3 7.

6 1

526

600

7.71

3.

0 40

.4

7.2

7.8

11.0

0.

7.

7 4.

13

0.5

7.6

0.87

0.

1 7.

7 72

8 78

8.

12

3.4

49.3

9.

2 7.

8 10

.3

0.1

7.7

3.56

0.

2 7.

7 0.

92

0.2

7.6

580

152

8.3

NaC

l co

ncen

trat

ion

(mM

)

0.1

1 10

10

0 10

00

Soi

l dep

th

(m)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

0.7

8 89

2 62

88

7.2

8 31

2 82

5 7.

1 2

744

32

7.2

826

198

7.1

163

29

6.66

1.

3 4

624

228

7.6

4 50

2 53

7.

6 1

851

315

7.7

520

110

7.9

93.9

11

8.

82

2.05

1

349

297

7.7

1 65

5 45

3 7.

7 89

3 43

7.

8 29

9 15

7.

9 64

.1

2.9

8.88

3.

0 1

015

94

7.7

1 89

4 41

2 7.

6 82

4 38

7.

8 28

0 16

7.

9 70

.7

9.0

8.71

3.

4 1

720

81

7.8

822

44

7.8

923

156

7.7

235

9.1

8.0

55.3

6.

1 8.

87

KC

l co

ncen

trat

ion

(mM

)

0.1

1 10

10

0 10

00

Soi

l dep

th

(m)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

0.7

3 03

2 35

9 7.

7 27

524

4519

7.

3 75

2 12

8 7.

0 9.

49

1.0

7.2

-11.

4 0.

1 7.

31

1.3

1 28

6 10

7 7.

4 24

6 13

7.

4 27

.9

0.9

7.2

-7.8

9 0.

7 8.

3 -1

1.9

0.1

8.87

2.

05

601

60

7.5

110

17

7.3

8.60

0.

9 7.

6 -9

.93

0.4

8.3

-12.

1 0.

1 8.

67

3.0

666

51

7.5

130

23

7.5

6.78

0.

2 7.

7 -9

.97

0.4

8.7

-11.

9 0.

1 8.

9 3.

4 52

1 19

7.

6 90

.3

3.3

7.5

6.80

5.

5 7.

9 -1

0.3

0.8

8.6

-943

2.

8 8.

79

134

135

AP

PE

ND

IX 5

. Kd v

alu

es o

f ca

esiu

m

App

endi

x 5.

6. K

d va

lues

of

caes

ium

for

the

exca

vato

r pi

t OL

-KK

20 a

erob

ic s

oil s

ampl

es in

the

pres

ence

of

NH

4 or

Ca

as c

ompe

ting

cat

ion.

N

H4C

l co

ncen

trat

ion

(mM

)

0.1

1 10

10

0 10

00

Soi

l dep

th

(m)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

0.7

1281

20

6.

4 7

143

9 29

7 7.

3 16

4 23

7.

2 5.

83

1.5

7.2

-0.4

2 0.

5 7.

2 1.

3 57

4 89

7.

3 12

0 5.

1 8.

2 12

.4

0.6

8.2

-2.1

6 3.

1 7.

8 -3

.60

3.7

8.2

2.05

40

7 92

6.

7 11

5 82

7.

4 7.

98

0.3

7.4

-0.2

5 0.

1 7.

5 -1

.06

0.6

7.4

3.0

8 80

1 1

165

7.1

292

340

7.9

9.58

1.

1 7.

8 2.

71

2.6

7.7

-0.1

8 0.

1 7.

9 3.

4 39

9 46

7.

4 44

.2

2.9

8.1

6.43

0.

1 8.

2 0.

45

0.1

8.0

-0.0

3 0.

1 8.

0 C

aCl 2

co

ncen

trat

ion

(mM

)

0.1

1 10

10

0 10

00

Soi

l dep

th

(m)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

0.7

5 37

1 1

004

6.3

3 72

2 34

6 6.

6 15

80

13

6.7

707

45

6.7

238

24

6.6

1.3

2 19

3 15

0 7.

2 1

601

350

7.4

1280

54

7.

6 50

6 27

7.

8 13

8 7.

4 7.

6 2.

05

1 11

5 31

2 6.

7 1

000

193

6.9

989

379

7.1

483

2.7

7.1

101

2.4

7.0

3.0

1 41

5 76

6.

8 1

089

20

6.8

921

19

7.2

430

31

7.1

91.4

0.

8 7.

1 3.

4 86

1 0.

6 7.

4 75

0 67

7.

7 53

0 11

6 7.

9 29

6 94

7.

9 61

.4

5.6

6.7

135

136

APPENDIX 6. Kd values of chlorine Appendix 6.1. Kd values of chlorine for the excavator pit OL-KK15 dried and fractioned humus samples.

0.25 months Grain size (mm)

Kd (ml/g) Standard deviation

(ml/g) pH

<0.063 -0.12 0.3 4.5 0.063-0.125 -0.07 0.5 4.4 0.125-0.25 0.00 0.1 4.3 0.25-0.50 0.02 0.1 4.2 0.50-1.0 0.20 0.1 4.2 1.0-2.0 -0.20 0.1 4.2

>2.0 -0.45 0.2 4.2 3 months

Grain size (mm) Kd (ml/g)

Standard deviation (ml/g)

pH

<0.063 60.9 0.5 4.5 0.063-0.125 19.8 0.8 4.4 0.125-0.25 30.1 0.1 4.3 0.25-0.50 30.4 0.6 4.3 0.50-1.0 29.5 1.0 4.4 1.0-2.0 27.3 0.7 4.3

>2.0 27.8 1.3 4.3

137

AP

PE

ND

IX 6

. Kd v

alu

es o

f ch

lori

ne

App

endi

x 6.

2. K

d va

lues

of

chlo

rine

for

the

exca

vato

r pi

t OL

-KK

21 u

ntre

ated

and

irra

diat

ed h

umus

sam

ples

1, 2

and

3.

Unt

reat

ed h

umus

sam

ples

T

ime

(day

s)

Hum

us 1

H

umus

2

Hum

us 3

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)S

tand

ard

devi

atio

n (m

l/g)

pH

K

d (m

l/g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

7 0.

96

0.1

6.7

0.82

0.

4 5.

5 0.

75

0.1

7.1

29

0.68

0.

1 7.

0 0.

49

0.2

6.0

0.73

0.

1 7.

2 93

0.

95

0.1

8.5

1.67

0.

4 5.

9 1.

00

0.3

7.0

186

0.89

0.

1 7.

8 -

- -

- -

- 36

5 0.

31

0.1

7.7

1.42

0.

4 6.

7 0.

46

0.1

7.7

Unt

reat

ed, i

rrad

iati

on s

teri

lize

d hu

mus

sam

ples

T

ime

(day

s)

Hum

us 1

H

umus

2

Hum

us 3

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)S

tand

ard

devi

atio

n (m

l/g)

pH

K

d (m

l/g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

7 1.

55

0.4

5.5

0.88

0.

1 4.

8 1.

24

0.1

6.8

93

-4.1

3 7.

6 5.

7 0.

80

0.4

4.7

1.18

0.

9 6.

3 -

sam

ple

not d

one

137

138

AP

PE

ND

IX 7

. Kd v

alu

es o

f io

din

e

App

endi

x 7.

1. K

d va

lues

of

iodi

ne f

or th

e so

il la

yers

of

the

exca

vato

r pi

ts O

L-K

K14

, OL

-KK

15 a

nd O

L-K

K16

.

Soi

l la

yer

OL

-KK

14

OL

-KK

15

OL

-KK

16

sam

ple

dept

h (c

m)

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

sam

ple

dept

h (c

m)

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

sam

ple

dept

h (c

m)

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

hum

us

5 60

4 53

6 4.

7 7

2 67

0 25

72

4.4

10

78

47

5.5

MS

1 20

55

8.

4 5.

4 50

2.

1 0.

3 6.

7 30

18

2.

9 7.

1

MS

2 60

1.

5 0.

2 6.

9 80

1.

1 0.

1 7.

4 50

7.

7 0.

9 7.

3

MS

3 10

5 0.

5 0.

1 7.

2 11

0 0.

03

0.1

8.2

110

2.7

0.4

8.3

MS

4 24

0 0.

1 0.

1 7.

9 16

0 0.

2 0.

1 8.

1 30

0 1.

1 0.

1 8.

5

MS

5 -

- -

- 26

0 -0

.6

0.1

8.6

- -

- -

- sa

mpl

e no

t tak

en

138

139

AP

PE

ND

IX 7

. Kd v

alu

es o

f io

din

e

App

endi

x 7.

2. K

d va

lues

of

iodi

ne f

or d

ried

and

fra

ctio

ned

hum

us s

ampl

es o

f th

e ex

cava

tor

pits

OL

-KK

14, O

L-K

K15

and

OL

-KK

16.

Gra

in s

ize

(mm

)

OL

-KK

14

OL

-KK

15

OL

-KK

16

Kd

(ml/

g)

Stan

dard

de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Stan

dard

de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Stan

dard

de

viat

ion

(ml/

g)

pH

<0.

063

359

17

4.7

3799

80

5 4.

5 -

- -

0.06

3-0.

125

389

41

4.8

2045

13

3 4.

5 16

3 33

5.

6 0.

125-

0.25

46

0 71

4.

7 2

303

372

4.5

144

64

5.6

0.25

-0.5

0 12

42

78

4.7

1 86

7 21

4 4.

4 47

0.

8 5.

5 0.

50-1

.0

786

154

4.8

2 71

5 43

4 4.

4 24

12

5.

5 1.

0-2.

0 38

7 8.

0 4.

7 3

290

158

4.4

10

0.1

5.4

- sa

mpl

e no

t don

e (n

o m

ater

ial l

eft)

139

140

AP

PE

ND

IX 7

. Kd v

alu

es o

f io

din

e

App

endi

x 7.

3. K

d va

lues

of

iodi

ne f

or th

e ex

cava

tor

pits

OL

-KK

20 a

erob

ic a

nd a

naer

obic

and

OL

-KK

21 a

erob

ic s

oil s

ampl

es.

Soi

l de

pth

(m)

OL

-KK

20

OL

-KK

21

aero

bic

anae

robi

c ae

robi

c

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

Eh

(Rm

V)

Soi

l de

pth

(m)

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

0.7

26

17

7.4

-0.1

0.

5 9.

0 37

9 0.

4 13

6 13

3 7.

4 1.

3 18

12

7.

9 -0

.4

0.2

9.6

261

1.1

22

18

7.7

2.05

26

12

7.

9 -0

.1

0.2

9.7

252

1.75

6.

0 5.

0 7.

9 3.

0 14

9.

8 7.

8 -0

.03

0.2

9.7

247

2.4

7.8

4.6

8.0

3.4

19

11

7.9

-0.4

0.

2 9.

8 24

0 2.

95

5.9

4.6

7.9

- -

- -

- -

- -

3.6

12

8.0

8.0

- sa

mpl

e no

t tak

en

140

141

AP

PE

ND

IX 7

. Kd v

alu

es o

f io

din

e

App

endi

x 7.

4. K

d va

lues

of

iodi

ne f

or th

e ex

cava

tor

pit O

L-K

K21

unt

reat

ed a

nd u

ntre

ated

, irr

adia

ted

hum

us s

ampl

es 1

, 2 a

nd 3

. U

ntre

ated

hum

us s

ampl

es

Tim

e (d

ays)

H

umus

1

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

1 23

3 8.

2 7.

5 3

344

19

7.4

5 47

5 0.

5 7.

3 7

244

288

7.0

21

17

0.6

7.7

93

9.6

1.0

7.6

Unt

reat

ed, i

rrad

iati

on s

teri

lize

d hu

mus

sam

ples

Tim

e (d

ays)

Hum

us 1

H

umus

2

Hum

us 3

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

7 2.

5 0.

2 5.

7 6.

0 0.

5 4.

8 10

11

5.

5 93

63

5.

9 5.

6 48

28

5.

3 59

0.

6 5.

5

141

142

AP

PE

ND

IX 8

. Kd v

alu

es o

f n

iob

ium

A

ppen

dix

8.1.

Kd

valu

es o

f ni

obiu

m f

or th

e ex

cava

tor

pit O

L-K

K20

aer

obic

and

ana

erob

ic s

oil s

ampl

es.

OL

-KK

20 a

erob

ic; t

race

r in

0.5

M H

NO

3 +

0.3

M H

F, f

ilte

red

Tim

e (d

ays)

7

21

42

63

Soi

l de

pth

(m)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

0.7

5 55

4 4

171

3.6

45 7

54

2 41

0 2.

8 41

955

4

212

2.9

5 18

43

2 45

7 3.

2 1.

3 27

040

27

952

5.

6 14

590

85

9 3.

3 23

919

7

972

3.8

54 4

483

642

098

3.9

2.05

14

098

13

354

4.

4 15

940

98

9 2.

9 14

856

1

116

3.3

17 9

33

2 20

4 3.

3 3.

0 16

613

11

927

4.

8 17

274

1

470

2.9

424

143

560

741

4.3

13 7

33

477

3.2

3.4

18 3

75

20 5

06

5.1

20 7

11

5 98

1 2.

7 52

303

38

011

3.

7 12

530

29

98

3.6

OL

-KK

20 a

erob

ic;

trac

er in

0.1

M H

Cl,

filt

ered

T

ime

(day

s)

7 21

42

63

Soi

l de

pth

(m)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

0.7

184

463

32 6

33

5.3

19 7

81

274

7.2

100

533

49 3

08

7.4

309

709

137

498

7.2

1.3

100

720

17 1

36

7.5

13 2

22

2 09

2 8.

0 28

709

3

096

7.9

39 2

84

17 7

53

8.0

2.05

68

654

7

406

7.5

17 2

53

8 11

4 7.

9 29

676

7

748

8.0

36 9

95

16 6

59

8.1

3.0

55 1

41

4 70

2 7.

4 2

788

871

7.9

21 1

41

4 89

5 7.

8 28

757

6

317

8.0

3.4

53 9

29

37 3

64

7.4

3 64

1 1

099

8.0

28 8

74

2 46

4 8.

0 68

185

47

900

8.

0 O

L-K

K20

an

aero

bic;

trac

er in

0.5

M H

NO

3 +

0.3

M H

F, 1

.0x1

0-3 M

CH

ES

, fil

tere

d T

ime

(day

s)

1 3

7 21

Soi

l de

pth

(m)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Eh

(Rm

V)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Eh

(Rm

V)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Eh

(Rm

V)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Eh

(Rm

V)

0.7

6 32

3 13

04

4.5

453

5 73

7 1

610

5.1

462

788

166

1 06

1 22

9 5.

5 33

3 10

2 99

5 58

600

5.

8 45

0 1.

3 7

412

55

5.8

249

2 27

6 31

8 5.

6 30

8 6

982

1 64

0 6.

4 19

4 7

699

1 04

6 7.

0 12

9 2.

05

469

31

5.4

261

1 14

8 47

0 6.

0 22

4 4

669

2 56

7 6.

4 19

1 5

366

1 74

3 6.

5 16

5 3.

0 35

9 27

5.

1 29

5 72

5 26

5.

6 24

6 3

653

81

6.2

203

6 77

2 16

0 6.

3 18

2 3.

4 51

0 11

9 3.

6 42

5 1

069

61

3.9

410

2 04

9 56

5 5.

8 24

0 5

588

1 63

5 6.

1 20

2

142

143

APPENDIX 8. Kd values of niobium

Appendix 8.2. Kd values of niobium for the aerobic and anaerobic soil samples of the excavator pit OL-KK20. Aerobic; 0.5 M HNO3 + 0.3 M HF, filtered

Soil depth (m) Kd (ml/g) Geometric standard

deviation (ml/g)pH Eh (RmV)

0.7 30 132 17 753 3.2 - 1.3 127 414 29 835 4.5 - 2.05 15 385 12 630 3.7 - 3.0 97 675 21 060 4.0 - 3.4 24 459 17 248 4.0 -

Aerobic; tracer 0.1 M HCl, centrifuged, unfiltered 0.7 20 397 19 078 6.9 - 1.3 9 091 7 800 7.6 - 2.05 10 018 8 627 7.7 - 3.0 7 223 5 783 7.7 - 3.4 5 042 4 077 7.6 -

Aerobic; tracer 0.1 M HCl, filtered 0.7 153 621 100 093 6.8 - 1.3 45 484 34 397 7.9 - 2.05 38 145 32 671 7.9 - 3.0 26 957 17 256 7.8 - 3.4 38 657 23 110 7.8 -

Anaerobic; tracer 0.5 M HNO3 + 0.3 M HF, 1.0x10-3 M CHES, filtered 0.7 48 271 29 901 5.3 425 1.3 3 862 3 068 6.0 220 2.05 2 670 1 844 6.1 210 3.0 2 935 1 592 5.7 232 3.4 2 709 1 559 5.3 319

- not measured

144

APPENDIX 8. Kd values of niobium

Appendix 8.3. Kd values of niobium for the untreated humus samples 1, 2 and 3 of excavator pit OL-KK21. Humus 1 Time (days)

Centrifuged, unfiltered Filtered


(ml/g) pH Kd (ml/g)

Standard deviation

(ml/g) pH

1 990 497 7.0 1054 485.6788 7.2 3 859 687 7.1 918 731.2094 7.5 7 969 224 7.1 1014 155.8739 7.2 28 1447 62 7.2 1768 202.8191 7.3

Humus 2 Time (days)



(ml/g) pH Kd (ml/g)

Standard deviation

(ml/g) pH

1 781 477 5.1 811 502 5.1 3 614 114 5.1 664 145 5.4 7 429 286 5.3 455 300 5.4 28 669 590 5.6 756 689 5.6

Humus 3 Time (days)



(ml/g) pH Kd (ml/g)

Standard deviation

(ml/g) pH

1 1077 219 7.1 1098 225 7.3 3 568 315 6.9 589 335 7.3 7 663 70 7.0 699 54 7.1 28 1350 33 6.9 1447 80 7.0

145

APPENDIX 9. Kd values of selenium Appendix 9.1. Kd values of selenium for the aerobic and anaerobic soil samples of the excavator pit OL-KK20. Aerobic; centrifuged, unfiltered

Soil depth (m) Kd (ml/g) Geometric standard

deviation (ml/g)pH Eh (RmV)

0.7 244 237 6.7 - 1.3 51.5 17 7.9 - 2.05 160 76 8.0 - 3.0 9.50 8.2 7.9 - 3.4 10.6 7.0 8.0 -

Aerobic; filtered 0.7 258 250 6.7 - 1.3 54.9 17 8.0 - 2.05 170 78 8.0 - 3.0 9.68 8.3 7.9 - 3.4 11.2 7.3 8.0 -

Anaerobic; filtered 0.7 98.2 97 7.0 202 1.3 53.6 46 8.0 88 2.05 5.00 4.6 8.1 60 3.0 5.23 3.9 8.1 86 3.4 12.8 6.0 7.9 131

- not measured

146

AP

PE

ND

IX 9

. Kd v

alu

es o

f se

len

ium

A

ppen

dix

9.2.

Kd

valu

es o

f se

leni

um f

or th

e ex

cava

tor

pit O

L-K

K20

aer

obic

and

ana

erob

ic s

oil s

ampl

es.

OL

-KK

20 a

erob

ic;

cen

trif

uge

d, u

nfi

lter

ed

Tim

e (d

ays)

7

21

42

63

Soi

l de

pth

(m)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

0.7

195

19

7.1

221

246

25

28

6.8

6.5

232

229

2.49

35

6.

6 6.

8 23

3 39

0 11

89

7.

0 6.

0

1.3

5.4

3.1

7.8

9.1

16.7

3.

2 4.

1 8.

0 7.

6 8.

3 31

4 0.

43

97

8.0

8.1

15

36.8

4.

8 13

8.

1 7.

7

2.05

22

10

7.

9 26

27

2 9.

3 32

1 8.

0 7.

8 58

8 88

.5

0.59

5.

5 8.

0 8.

1 98

17

0 20

16

8.

0 8.

1

3.0

4.2

1.3

7.8

5.8

9.99

0.

3 0.

8 7.

9 7.

8 8.

5 19

.2

0.40

1.

9 8.

0 7.

8 7.

5 15

.9

2.5

0.7

8.0

7.8

3.4

3.7

1.3

7.9

4.4

9.79

0.

1 3.

1 7.

9 7.

8 5.

2 21

.9

0.39

18

8.

0 8.

1 6.

4 29

.4

0.7

31

8.0

8.1

OL

-KK

20 a

erob

ic;

filt

ered

Tim

e (d

ays)

7

21

42

63

Soi

l de

pth

(m)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g))

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

0.7

204

27

7.1

223

251

26

23

6.7

6.5

234

314

18

39

6.6

6.64

23

7 39

5 11

97

6.

8 6.

0

1.3

5.6

3.0

7.9

9.1

16.9

3.

2 3.

8 8.

0 8.

0 8.

6 34

1 0.

6 95

8.

05

8.14

15

37

.3

4.8

12

8.1

8.1

2.05

20

9.

0 7.

9 27

29

2 9.

7 34

6 8.

0 7.

9 62

7 91

.0

38

5.0

7.97

8.

09

98

182

20.1

20

8.

0 8.

1

3.0

4.6

1.5

7.9

5.8

10.2

0.

4 0.

7 8.

0 7.

8 8.

8 19

.7

3.1

1.7

7.97

8.

03

7.5

16.3

2.

6 0.

7 8.

0 8.

0

3.4

3.9

0.7

7.9

4.5

10.0

0.

2 3.

3 8.

0 7.

9 5.

2 23

.4

0.2

20

8.02

8.

09

6.5

31.8

0.

4 34

4 8.

0 8.

2

OL

-KK

20 a

nae

robi

c;fi

lter

ed

Tim

e (d

ays)

7

31

Soi

l de

pth

(m)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Eh

(Rm

V)

Kd

(ml/

g)

Sta

ndar

d de

viat

ion

(ml/

g)

pH

Eh

(Rm

V)

0.7

100

2.24

6.

9 77

95

.9

14

7.1

261

1.3

33.3

1.

13

8.1

47

74.0

0.

3 8.

3 88

2.05

6.

94

0.70

8.

0 67

3.

06

2.8

8.3

84

3.0

8.68

0.

72

7.8

5.5

1.79

3.

6 8.

5 12

0

3.4

24.2

0.

96

7.7

13

1.50

19

8.

6 21

9

146

147

APPENDIX 9. Kd values of selenium

Appendix 9.3. Kd values of selenium for the humus samples 1, 2 and 3 of the excavator pit OL-KK21. Humus 1 Time (days)



(ml/g) pH Kd (ml/g)

Standard deviation

(ml/g) pH

1 67 1.4 6.8 67 2.2 6.8 3 99 6.2 6.7 100 5.6 6.8 7 174 93 6.9 177 95 7.0 21 302 0.9 6.7 306 1.6 6.8 42 475 28 7.1 485 33 7.3 63 1 394 156 7.3 1 519 212 7.3

Humus 2 Time (days)



(ml/g) pH Kd (ml/g)

Standard deviation

(ml/g) pH

1 75 4.8 4.4 75 4.7 4.4 3 194 11 4.4 197 10 4.4 7 312 11 4.4 322 6.2 4.4 21 350 0.5 4.5 360 2.4 4.5 42 422 35 4.7 428 31 4.5 63 433 13 4.9 444 14 4.7

Humus 3 Time (days)



(ml/g) pH Kd (ml/g)

Standard deviation

(ml/g) pH

1 57 1.3 6.3 58 1.2 6.3 3 94 4.3 6.4 96 4.0 6.4 7 151 2.2 6.5 153 3.1 6.7 21 321 27 6.6 326 28 6.7 42 465 121 7.1 474 124 7.3 63 785 65 7.0 821 85 7.2

148

AP

PE

ND

IX 1

0. K

d v

alu

es o

f te

chn

etiu

m

App

endi

x 10

.1. K

d va

lues

of

tech

neti

um f

or th

e so

il la

yers

of

the

exca

vato

r pi

ts O

L-K

K14

, OL

-KK

15 a

nd O

L-K

K16

.

Soi

l la

yer

OL

-KK

14

OL

-KK

15

OL

-KK

16

sam

ple

dept

h (c

m)

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

sam

ple

dept

h (c

m)

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

sam

ple

dept

h (c

m)

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

hum

us

5 15

.1

13

5.0

7 12

6 10

5 4.

6 10

1.

44

1.4

6.1

MS

1 20

0.

20

0.4

5.7

50

-0.0

9 6.

8 7.

0 30

0.

45

0.7

7.4

MS

2 60

0.

63

0.6

6.9

80

1.02

0.

4 7.

5 50

0.

80

0.5

7.6

MS

3 10

5 0.

62

0.5

7.1

110

0.40

0.

3 8.

1 11

0 0.

93

0.3

8.1

MS

4 24

0 0.

54

0.6

7.6

160

0.36

0.

9 8.

1 30

0 1.

80

0.9

8.2

MS

5 -

- -

- 26

0 -1

.06

3.1

8.8

- -

- -

- sa

mpl

e no

t tak

en

148

149

AP

PE

ND

IX 1

0. K

d v

alu

es o

f te

chn

etiu

m

App

endi

x 10

.2. K

d va

lues

of

tech

neti

um f

or th

e ex

cava

tor

pits

OL

-KK

20 a

erob

ic a

nd a

naer

obic

and

OL

-KK

21 a

naer

obic

soi

l sam

ples

. S

oil

dept

h (m

)

OL

-KK

20

OL

-KK

21

aero

bic

anae

robi

c an

aero

bic

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

Eh

(Rm

V)

Soi

l de

pth

(m)

Kd

(ml/

g)

Geo

met

ric

stan

dard

de

viat

ion

(ml/

g)

pH

Eh

(Rm

V)

0.7

1.38

0.

8 7.

7 2.

19

1.4

8.7

367

0.4

42.7

38

7.

2 84

1.

3 1.

27

1.2

8.1

3.04

2.

4 9.

1 26

0 1.

1 1.

05

0.7

8.1

134

2.05

1.

45

1.2

8.1

1.10

1.

0 9.

3 22

6 1.

75

39.8

14

8.

8 10

8 3.

0 1.

27

0.3

8.0

0.86

0.

7 9.

4 20

3 2.

4 55

4 54

8.

6 11

6 3.

4 1.

22

1.1

8.0

2.62

0.

7 9.

2 21

3 2.

95

101

82

8.9

46

- -

- -

- -

-

3.6

568

166

8.8

73

- sa

mpl

e no

t tak

en

149

150

APPENDIX 10. Kd values of technetium

Appendix 10.3. Kd values of technetium for the excavator pit OL-KK21 untreated and untreated, irradiated humus samples 1, 2 and 3. Untreated humus samples Time (days)

Humus 1


(ml/g) pH

7 3.54 0.5 7.0 31 366 38 7.7 93 547 39 7.2

Untreated, irradiation sterilized humus samples

Time (days)

Humus 1 Humus 2 Humus 3

Kd (ml/g)

Standard deviation

(ml/g) pH

Kd (ml/g)

Standard deviation

(ml/g) pH

Kd (ml/g)

Standard deviation

(ml/g) pH

7 1.82 0.3 5.4 4.46 4.7 4.8 2.38 0.1 5.7 93 8.82 0.4 5.4 12.3 0.3 4.7 7.39 1.4 5.5

distribution coefficients of caesium, chlorine, iodine, niobium

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