Behavior of Radiocesium in the Forest

Abstract

This chapter gives the whole picture of how the radiocesium that fell into the forest has been moving in the forest, based on the results of a decade-long survey. The radiocesium entered the forest from the atmosphere as “fallout” immediately after the accident. Most of the radiocesium was initially trapped on the leaves and branches, and then the radiocesium transferred to the forest floor through litterfall and rain within a few years. A small percentage of the total amount of radiocesium in the forest is absorbed from soil by trees, and eventually returns to the ground surface as litterfall; in other words, it is cycling in the forest. Forests have the ability to retain radiocesium. At present, most of the radiocesium transferred to mineral soil, the circulation of radiocesium in the forest is slow, and the behavior of radiocesium in the Fukushima forests is about to enter the phase of quasi-equilibrium and equilibrium. To understand the behavior of the radiocesium and predict the future status, modeling approaches are being carried out in conjunction with field observations.

3.1 Overview of Behavior

The radiocesium in the forest entered the forest from the atmosphere as “fallout” immediately after the accident. After that, radiocesium moves in the forest through the movement of water, fallen leaves, absorption by trees, etc. Such a large movement and cycling of radiocesium in the ecosystem can be said to be a characteristic of forest ecosystems that is not found in agricultural land (Fig. 3.1). The distribution of radiocesium in the forest changed significantly from the time when the fallout occurred immediately after the accident to the early phase (Fig. 3.2), but the amount of transfer gradually decreased in the transition phase of about 10 years after the accident. Soon, within a few decades, a period of stability (equilibrium) will be reached in which radiocesium appears to be almost unchanged on an annual basis (Fig. 3.2). This does not mean that the movement of radiocesium has completely stopped, but the amount of movement is small, and for example the amount absorbed by trees from the soil and the amount released from trees to the soil are balanced, which is called a steady state (equilibrium). Based on our monitoring results, we think that the behavior of radiocesium in forest is currently (around 2021, 10 years since the accident) in a transition phase. The transition phase is sometimes referred to as the quasi-equilibrium state because it is approaching equilibrium (Sect. 3.5). We would like to take a closer look at what exactly was happening during each of these phases.

Fig. 3.1

Major movements of radiocesium in the forest (Arrows show the movement of radiocesium)

Fig. 3.2

Schematic diagram of the behavior of radiocesium in the forest and its change with time

3.2 Introduction: Two Types of Radiocesiums: Cesium-134 and Cesium-137

Before we look at the behavior of radiocesium in the forest, let’s talk about two different radioactive cesiums (radiocesiums): cesium-134 and cesium-137 . As we saw in the previous chapter, the types of radiocesium released by the Fukushima nuclear accident have a short physical half-life of 2 years (cesium-134 ) and a long physical half-life of 30 years (cesium-137 ), and they are considered to have been released in an approximate 1:1 ratio during the accident. Cesium-134 has a short half-life; the radiation dose is accordingly greatly reduced in the early phase (see Chaps. 2 and 6 for details). In general, the term “radiocesium” is sometimes used to refer to these two radionuclides together, and sometimes to cesium-137 alone. Since cesium-134 has a short half-life, the behavior of radiocesium in the forest on an annual basis is focused on cesium-137 . This is because cesium-137 remains in the forest in the long term and should be watched more closely, and the mechanism of movement of cesium-134 and cesium-137 in the forest is exactly the same. In this chapter we will refer to cesium-137 as “radiocesium” and look at its behavior in the forest to discuss its dynamics in the forest on an annual basis.

3.3 Large Changes in the Distribution of Radiocesium in the Early Post-Accident Phase

The radiocesium that falls on forests moves the most in the years following the accident.

3.3.1 Most of the Radiocesium that Fell on the Forest Was Initially Trapped onto the Leaves and Branches

The trees in a forest reach a height of 10–30 m and have leaves and branches that extend out from the trunk of the tree to form a layer (canopy ) that catches the light and rain. The leaves spread out and cover the ground surface to efficiently capture the energy of sunlight. In particular, cedar and cypress trees, which are the dominant tree species in artificial forests in Japan, spread their branches and leaves in layers, which makes it easy for them to trap radiocesium falling from the sky. For example, Kato et al. [13] reported that even 5 months after the accident, over 60% of the radiocesium in coniferous cedar and cypress forests was still attached to the canopy of the trees. In addition, Gonze and Calmon [14] collected the results of studies on cedar and cypress forests and estimated that about 90% of the fallen radiocesium was attached to the trees in the period from a few days to a few weeks after the accident. Thus, it can be considered that 60–90% of the radiocesium that fell on cedar and cypress forests was first trapped by leaves and branches.

However, the rate of radiocesium trapped by trees is considered to vary depending on the tree species. Since the accident at the Fukushima Daiichi Nuclear Power Plant occurred in March, the leaves of deciduous broadleaf trees such as konara oak had not yet opened when radiocesium fell out. Therefore, the trap rate of deciduous broadleaf trees by the canopy is considered to be lower than that of the evergreen coniferous trees that had leaves. In addition, although pine is an evergreen conifer , the surface area of its leaves is smaller than that of cedar and cypress, so it is thought that its trap rate was lower. According to observations made by the Forestry and Forest Products Research Institute in the summer of 2011, the ratio of radiocesium retained in the canopy of cedar forests as a percentage of the total forest ranged from 22% to 44%, while the ratio was around 18% in mixed forests of red pine and konara oak [15].

Radioactive materials released into the environment as a result of a nuclear power plant accident become very small liquid or solid particles (aerosols) and are transported through the atmosphere according to the wind at that time. When it rains, a lot of radioactive materials fall out and are deposited on the ground. This is how the highly contaminated area stretching northwest from the Fukushima Daiichi Nuclear Power Plant was created. In the case of forests, in addition to weather conditions, different types of forests, such as different leafiness and density of the trees at the time of the accident, affected the behavior of radiocesium as described above.

3.3.2 Then Radiocesium Transferred to the Forest Floor Through Litterfall and Rain

Radiocesium trapped in the canopy is transported to the forest floor through litterfall, throughfall (rainfall from/through the canopy ), and stem flow (flow of intercepted water down the trunk). In the early phase, the contribution of each pathway to the transport of radiocesium varied significantly with time. The radiocesium concentration in rainfall in the forest was high in the first months after the accident, but it became lower after the summer of 2011 [16,17,18]. On the other hand, the radiocesium concentration in litterfall and stem flow decreased gradually over a period of years, and their contribution increased after autumn 2011. However, it has become clear from observations that the contribution rate varies depending on the forest [17, 19].

The radiocesium concentration decreased with time due to the leaching effect, shedding of contaminated leaves and branches by litterfall, and development of new leaves and branches with low radiocesium concentration. Since the trend of decrease in concentration was generally exponential from the beginning to the transition phase, the concept of half-life (ecological half-life) can be applied (Fig. 3.3). The exponential change is expressed linearly when the vertical axis is logarithmic, as shown in Fig. 3.3. According to monitoring by the Forestry and Forest Products Research Institute, the radiocesium concentration in leaves and branches of cedar forests decreased in the 5 years up to 2015 with an ecological half-life as fast as 6 months to one and a half years [20]. The radiocesium concentration in bark decreased to less than half of its original concentration over the same period, but much more slowly than leaves and branches due to its longer lifespan and less radiocesium fall-off due to its shape and texture. As a result, immediately after the accident, the radiocesium concentration in the bark was lower than in the leaves and branches, but a few years later, the radiocesium concentration in the bark was higher than in the leaves and branches. In addition, the radiocesium concentration in the leaves of konara oak, which had not opened its leaves at the time of the accident, was lower than that of cedar and red pine in 2012, but there was no noticeable change in the concentration after that.

Fig. 3.3

Temporal changes in activity concentration of cesium-137 in (a) leaves, (b) branches, and (c) bark of Japanese cedar and konara oak (example from Kawauchi Village). Note that the vertical axis is the common logarithm. The cesium-137 concentrations in leaves and branches decreased exponentially, dropping to about one-hundredth in 5 years, while the cesium-137 concentrations in bark decreased only about one-third in Japanese cedar and hardly at all in konara oak (Source: Data from a survey by the Forestry Agency and the Forestry and Forest Products Research Institute, “Surveys on Radioactive Cesium Distribution in Forests” [21])

Some studies that have monitored the transfer of radiocesium from trees to soil at high frequency over time since the immediate aftermath of the accident suggest that migration may be better described by a superposition of two exponential functions (double exponential function) with a fast and a slow transfer component than by a single exponential function [16, 17]. This phenomenon is also observed in rivers and oceans, and is thought to be due to changes in the processes driving radiocesium transfer over time.

3.3.3 Not Remain Long in the Soil Surface Organic Layer

The radiocesium that transferred from trees to the ground surface during the initial period of rapid change did not remain in the organic layer (litter layer/O horizon) on the ground surface for long in many forests, but quickly transferred to the mineral soil underneath (Fig. 3.4a). The radiocesium concentration in the surface organic layer decreased despite the fact that radiocesium was supplied from trees by fallen leaves and rain. On the other hand, the radiocesium concentration in mineral soil showed an increasing trend after the accident (Fig. 3.4b).

Fig. 3.4

Temporal changes in activity concentration of cesium-137 in (a) organic layer and (b) surface mineral soil (0–5 cm depth) (example from Kawauchi Village). Note that the range of the vertical axis is different in (a) and (b) to make temporal changes easier to understand. Data for konara oak in Otama Village is also included. (Source: Data from a survey by the Forestry Agency and the Forestry and Forest Products Research Institute, “Surveys on Radioactive Cesium Distribution in Forests” [21])

One reason is that the organic layer of Japanese forests is thin, only a few centimeters thick. In European countries, such as Ukraine, Belarus and Fennoscandian countries, which were contaminated by the Chernobyl nuclear accident, the decomposition of organic matter is slower and the organic layer is thicker (e.g. more than 10 cm) due to cooler weather and less precipitation. As a result, radiocesium was retained in the thick organic layer in Europe. In general, in mineral soils, radiocesium is strongly adsorbed and fixed by clay minerals, making it difficult to move (see next section). However, in the organic layer, which does not contain clay minerals, radiocesium can move easily and be absorbed by plants. Future research will clarify how this difference in the retention of radiocesium in the organic layer affects the long-term movement of radiocesium in the forest.

3.4 Radiocesium in Soil

Mineral soils have a mechanism to hold radiocesium in the surface layer.

3.4.1 Most of the Radiocesium Remains in the Surface Layer of Mineral Soil

Radiocesium transferred to mineral soil mainly stays in the shallow part of the soil . Figure 3.5 shows the results observed in 2011 and 2017 in the soil of a cedar forest in Kawauchi Village, Fukushima Prefecture. Some radiocesium had already reached a depth of 20 cm even in the summer of 2011. By 2017, the radiocesium in the organic layer had decreased significantly due to its transfer to the mineral soil, and most of it was distributed and remained at the shallowest depth of 0–5 cm.

Fig. 3.5

Change in depth profile of cesium-137 in soil sampled in the summer of (a) 2011 and (b) 2017 (Japanese cedar forest, Kawauchi Village) (Source: Data from a survey by the Forestry Agency and the Forestry and Forest Products Research Institute, “Surveys on Radioactive Cesium Distribution in Forests” [21])

3.4.2 Why Does Radiocesium Remain in the Surface Layer?

The retention of radiocesium in shallow layers of the mineral soil is due to the strong adsorption and fixation of cesium by clay minerals in the soil . There are several mechanisms for adsorption and fixation of cesium in the soil , and the ability to adsorb/fix cesium varies depending on the type. The weakest immobilization occurs when cesium is adsorbed by the negative charge at the surface of soil organic matter and clay minerals. Since this negative charge has a high affinity for other cations, the adsorbed cations are easily exchanged and cesium is released [22, 23].

The ability of clay minerals to fix cesium differs depending on the type of mineral. Broadly speaking, there are two types of minerals: (1) those with high selectivity for cesium but reversible adsorption (both adsorbing and releasing cesium) (Fig. 3.6a), and (2) those with high selectivity for cesium and high fixation power (having frayed edge sites that are not easily releasable once adsorbed and fixed) (Fig. 3.6b). Some clay minerals, such as illite (a type of mica) and vermiculite (used as a soil conditioner), have a frayed edge site structure, and it is said that once ionic cesium enters the frayed edge site, it is difficult to be released again. An index called radiocesium interception potential¹ (RIP) is used to evaluate the immobilization capacity of radiocesium by frayed edge sites. There is also a study that shows that the absorption of radiocesium by plants (grasses) is suppressed in soils with high RIP [24]. This study also showed that the penetration of radiocesium into the soil depths was faster in soils with low RIP. It can be said that soil properties greatly affect the behavior of radiocesium in the forest.

Fig. 3.6

Types of cesium (ionized) fixation by clay minerals. (a) Adsorbed, but not fixed (ex. montmorillonite), (b) strongly fixed (ex. illite or vermiculite) (Source: Adapted from Ministry of the Environment, BOOKLET to Provide Basic Information Regarding Health Effects of Radiation, “Chap. 4: Concept of Radiological Protection, 4.4 Long-term Effects, Behavior of Radioactive Cesium in the Environment: Adsorption and Fixation by Clay Minerals” [1])

There have been several studies in Fukushima on the fixation of radiocesium in the surface organic layer and mineral soil. In Fukushima, where not many years have passed since the accident, retention of radiocesium by organic matter is also important. In a study by Toriyama et al. [25] on the amount of radiocesium in the soil particles classified by the specific gravity, it was found that light particles, which are considered to be organic matter, have a radiocesium concentration that is about eight times higher than that of heavier particles of mineral matter, and that organic matter particles retain 40% of the radiocesium contained in soil at a depth of 0–5 cm, despite the fact that they are only 10% by weight. In contrast to the previous results, which were mainly studied in agricultural lands with less organic matter than in forests, the study in forests with more organic matter suggests that the function of radiocesium fixation by organic matter as well as clay minerals cannot be ignored, at least in forest soils within a few years after the accident.

Some studies have examined the retention of radiocesium in the surface organic layer and mineral soil in detail. Manaka et al. [26] conducted extraction experiments from soil surface organic layer and mineral soil, and clarified the changes in the ratio of exchangeable radiocesium to the total amount of radiocesium in the organic layer and mineral soil for about 7 years after the accident. Their results showed that exchangeable radiocesium, which accounted for 10% and 6% of the total amount of radiocesium in the organic layer and mineral soil, respectively, in the first 5 months after the accident, decreased to about 2–4% after 2–4 years, and then remained stable and largely unchanged. This suggests that even after a sufficient amount of time has passed since the transfer of radiocesium into the soil , not all of the radiocesium adsorbed and fixed on organic matter and clay minerals will be strongly fixed, but some will be repeatedly adsorbed and released, and may exist in the soil in a form that is easily transferred and absorbed by plants. How fast the fixation of radiocesium proceeds in the soil and how easily it can move and reach an equilibrium state is an unresolved issue that needs to be watched very closely because it will affect the absorption by trees in the future.

Another important factor of the soil in the absorption of radiocesium by plants is potassium, one of the major essential elements for plants. Cesium and potassium are both alkali metal elements, and in the soil they are dissolved in water, ionized and exchanged, adsorbed by clay minerals and organic matter, and some of them are absorbed by plants as ions. Cesium is a larger element than potassium, but when ionized, its radius is close to that of potassium, and it behaves similarly in adsorption on clay minerals and absorption by plants. Potassium and cesium are absorbed into the plant through the roots via the same pathway, but the more potassium, an essential element, is present in the soil , the less cesium is absorbed by the plant. Using this property, potassium fertilization has become a condition for resuming farming in areas contaminated by radioactivity to reduce the absorption of radiocesium into crops. In forests as well, it was known that potassium in the soil was effective in reducing the absorption of radiocesium by trees, but there were not many research examples. After the Fukushima nuclear accident, the effect of potassium has been confirmed through potassium fertilization experiments in hinoki cypress forests [27] and tests using konara oak trees (Quercus serrata ) in the laboratory [28]. In addition, although not through fertilization, studies on konara oak coppice forests (forests in which new branches have grown out from stumps to become mature trees again) have revealed that exchangeable potassium in forest soil strongly affects the absorption of radiocesium by trees (Sect. 6.5).

Cesium is one trace element not necessary for tree growth, but the stable isotope (non-radioactive) cesium-133 (¹³³Cs) is originally found in the earth’s crust and soil in small amounts (a few μg in 1 g of soil [29]). Cesium is chemically similar to potassium, which is an essential element for plants, so it behaves similarly to potassium in soil and plant bodies, but not exactly the same. It should also be remembered that the radiocesium released from the accident fell on the forest in very small amounts, from 1/1000,000 to 1/1000 of the amount of the naturally occurring stable isotope cesium-133 (Chap. 4).

3.4.3 Migration of Radiocesium by Soil Animals and Fungi

As we have seen, radiocesium in the forest transfers from the aboveground to the ground surface by water movement (e.g. rainfall) and tree defoliation. Once it reaches the mineral soil, it moves very slowly within the soil layer, but over a long period of time, it gradually moves from top to bottom following water and gravity. On the other hand, the movement of radiocesium in the soil by soil animals and fungi cannot be ignored. For example, earthworms feed on the soil and excrete it as feces on the surface. Such disturbance by organisms is called bioturbation. It is not clear to what extent they contribute to the movement of radiocesium in the forest, as they have not been fully quantified in the Fukushima forests, but it is likely that some radiocesium is moved from top to bottom and vice versa by the activities of earthworms and other soil animals.

The function of fungi has also become clear. The high radiocesium concentration in mushrooms (the fruiting bodies of fungi) is widely known to be caused by the collection of radiocesium by mycelium. Many fungi have the ability to selectively absorb potassium. It is thought that cesium, which has chemically similar properties, is also absorbed along with potassium. Depending on the species, fungi that spread their mycelium widely in the organic layer and shallow part of the mineral soil are also considered to be involved in the transfer of radiocesium to some extent [30, 31]. Using this function, a method called “mycoextraction” has been devised in which wood chips are laid on top of contaminated soil to collect radiocesium from contaminated mineral soil using the power of natural fungi, and the wood chips are taken out of the forest ecosystem for disposal. The recovery rate is a few percent [30].

3.5 Transfer of Radiocesium into the Tree

Uptake into and movement within the tree vary depending on the species and environment.

Some of the radiocesium that fell directly onto the tree canopy or ground surface immediately after the accident, or that transferred from the tree canopy to the ground surface, is taken up by the tree. There are two pathways of uptake: radiocesium attached to the leaves and bark is directly absorbed into the tree, and radiocesium in the soil is absorbed through the roots. It has been reported that 25% of the total deposited amount of radiocesium was absorbed into the body of the tree from the surface of the tree other than the roots in an experiment in which radiocesium was sprayed directly onto saplings [32]. Imamura et al. showed that about half of the radiocesium in the tree could have been absorbed by the roots based on the data from the Fukushima study [33]. Immediately after the accident, radiocesium can be absorbed by both surface and root absorption pathways, but since most of the radiocesium will transfer to the soil within 2–3 years, absorption through roots is expected to account for most of the absorption thereafter.

The radiocesium that is taken into the tree moves through the tree with the flow of water and nutrients (translocation) [32, 34]. Trees, whether deciduous or evergreen, also shed their leaves. Before the leaves fall, some of the cesium is drawn back from the leaves into the tree just like any other elements, but the cesium contained in the fallen leaves transfers to the ground surface. Immediately after the accident, the distribution of radiocesium in the tree was uneven due to the multiple absorption pathways and the high concentration in the leaves and branches among the various parts of the tree, but the radiocesium concentration in each part of the tree changed over time due to the balance of absorption from the roots, nutrient and water transfer in the tree, and release by litterfall. And the transfer, absorption, and release between parts of the tree will then approach a balanced state during the transition phase.

3.5.1 Movement of Radiocesium in a Tree

The main part in a tree is the wood (xylem), which is used as lumber. The wood here refers to the trunk of the tree without the bark (Fig. 3.7). Since wood is used for various purposes such as construction, paper, mushroom cultivation (with bark), etc., special attention was paid to it after the accident and continuous monitoring was conducted. As a result, it was found that the trend of changes in the radiocesium concentration in wood varies greatly depending on the species. For example, the radiocesium concentration of Japanese cedar increased slightly or did not change much, while the concentration of konara oak increased significantly after the accident. In addition, when comparing the results of surveys conducted at several sites, there were large differences in the radiocesium concentration in wood of different sites, even for the same tree species. It has been reported that the potassium status in the soil affects the radiocesium concentration in trees (Sect. 6.5). In addition, studies conducted in Europe after the Chernobyl nuclear accident reported that the radiocesium concentration in trees growing in hydromorphic soil (soil that becomes reduced due to stagnation of water, such as in valley sides) and soil with thick deposits of humus tends to increase [35, 36]. Therefore, the differences in radiocesium concentration among the same tree species in the studied forest can be considered to be greatly influenced by the nature of the growing soil , in addition to the differences in radiocesium deposition in the forests (detailed figures are given in the next section).

Fig. 3.7

The horizontal structure of wood (Photo: Courtesy of Shinta Ohashi, the Forestry and Forest Products Research Institute)

The wood is further divided into sapwood (outer part) and heartwood (inner part) (Fig. 3.7). Sapwood contains living cells that transport water and provide mechanical support for the entire tree. Heartwood, on the other hand, is made up of dead cells and has only a mechanical support function and is generally distinguishable from sapwood by its discoloration. The distribution of radiocesium in the sapwood and heartwood was also found to vary greatly depending on the species (Fig. 3.8). For 1–2 years after the accident, the radiocesium concentration in sapwood was higher than that in heartwood, regardless of the species. Thereafter, the radiocesium concentration in the heartwood of Japanese cedar gradually increased, and in one study site the radiocesium concentration in the heartwood was about twice as high as in the sapwood [37]. On the other hand, the concentration in heartwood remained low in konara oak. In cedar trees, potassium concentration in heartwood is known to be higher than in sapwood [38]. Similar to potassium, cedar trees seem to have a mechanism to move radiocesium to the inner part of the wood. The mechanisms and causes of this are currently being studied.

Fig. 3.8

Temporal changes in activity concentration of cesium-137 in the (a) sapwood and (b) heartwood of Japanese cedar and konara oak (example from Kawauchi Village) (Source: Data from a survey by the Forestry Agency and the Forestry and Forest Products Research Institute, “Surveys on Radioactive Cesium Distribution in Forests” [21])

It is thought that the amount of radiocesium of each part of the tree will gradually become balanced between each part of the forest while cycling between trees and soils during the transition phase, approaching the steady phase where radiocesium concentrations appeared to be unchanged. According to the report of the International Atomic Energy Agency (IAEA , Sect. 5.3) after the Chernobyl nuclear accident, radiocesium in forests is considered to be in an early phase in which large changes in distribution occur for about 5 years after the accident, and after that it is considered to shift to a quasi-equilibrium state in which changes gradually become small [39, 40]. In Japan, 10 years have passed since the Fukushima nuclear accident, and the changes in the distribution in the forest have become smaller, so it can be said that we are approaching the equilibrium state after the early and transition phases.

3.5.2 Transfer Factor: Different Species Have Different Radiocesium Concentrations in Wood

The transfer factor (TF) is an indicator to estimate the degree of contamination of agricultural crops to protect people from internal exposure through food in areas contaminated by radioactivity. This is because the concentration of radioactive materials in contaminated crops varies depending on the type of crop and the type of soil in which it grows. In the case of crops, the TF is calculated as follows (Fig. 3.9a).

$$ TF=\frac{\mathrm{the}\ \mathrm{activity}\ \mathrm{concentration}\ \mathrm{of}\ \mathrm{radionuclide}\ \mathrm{in}\ \mathrm{plants}}{\mathrm{the}\ \mathrm{activity}\ \mathrm{concentration}\ \mathrm{of}\ \mathrm{radionuclide}\ \mathrm{in}\ \mathrm{soil}}\left(\frac{Bq/ kg}{Bq/ kg}\right) $$

The transfer factor concept has since been applied to the absorption and transfer of radioactive materials by various organisms in the environment. The aggregated transfer factor (T_ag) , which is different from that for agricultural crops, is used to evaluate the degree of transfer for forest trees or other products, and is calculated as follows (Fig. 3.9b).

Fig. 3.9

Different definitions of (a) transfer factor (TF) and (b) aggregated transfer factor (T_ag) . Both coefficients are ratios, but in the case of transfer factor, the units are canceled out because the units of the numerator and denominator are the same

$$ {T}_{\mathrm{ag}}=\frac{\mathrm{activity}\ \mathrm{concentration}\ \mathrm{of}\ \mathrm{radionuclides}\ \mathrm{in}\ \mathrm{tree}\ \mathrm{compartments}\ \mathrm{or}\ \mathrm{forest}\ \mathrm{products}}{\mathrm{the}\ \mathrm{to}\mathrm{tal}\ \mathrm{deposition}\ \mathrm{to}\ \mathrm{forest}\ \mathrm{floor}\ \mathrm{per}\ \mathrm{unit}\ \mathrm{area}}\left(\frac{Bq/ kg}{Bq/{m}^2}\right) $$

While the radiocesium concentration in the soil is approximately uniform in the depth direction due to plowing and mixing in the agricultural land, the distribution of the radiocesium concentration in the soil differs greatly in the depth direction in forest soil , which is not plowed (Fig. 3.5). In addition, there is an organic layer on the forest surface. Therefore, if we collect soil samples in forests and calculate the transfer factor in the same way as for agricultural land, the factor varies greatly depending on the depth from which the soil is collected. Therefore, it is expressed as a T_ag with the integrated amount of radiocesium in soil per unit area (inventory ) as the denominator and the radiocesium concentration in trees as the numerator. These TFs and T_ags are not directly comparable because they have different definitions and units.

Another caveat is that the T_ag obtained in forests where the time since the accident has not been long enough, such as in the case of the Fukushima nuclear accident, do not necessarily reflect the absorption of radiocesium from the soil by trees. This is due not only to the absorption, but also to the large influence of radiocesium deposited directly on the tree surface. In the case of forest trees, T_ag can vary by a factor of 10 or more from the early to transition phase.

Observations in Europe affected by the Chernobyl nuclear accident are summarized in the International Atomic Energy Agency report [40] and in Calmon et al. [35]. According to this report, the T_ag for wood of coniferous trees is 1.5 × 10⁻³ m²/kg (geometric mean, n = 31 surveys), for wood for deciduous trees 3.5 × 10⁻⁴ m²/kg (geometric mean, n = 12 surveys), and the combined average is 1.0 × 10⁻³ m²/kg (geometric mean). Although the number of studies is not necessarily large, it has been reported that the values differed by tree species and soil (e.g. very high in peat soils), and were higher in wet conditions.

There have been several reports after the Fukushima nuclear accident. Ohashi et al. compiled data from published papers and reports in a project of the International Atomic Energy Agency. The data for 2015, when multiple data were taken, showed an increase from about 10⁻⁴ to 10⁻³ m²/kg for Japanese cedar and cypress, and from about 10⁻⁴ to 10⁻³ m²/kg for konara oak (Fig. 3.10) [9]. When the T_ags were compared for pine and oak trees with European studies after the Chernobyl nuclear accident, the values obtained from the Fukushima studies were lower for pine, but the number of data is not sufficient, so further verification is necessary. For conifers other than pine and oak, the values were similar [9].

Fig. 3.10

Temporal changes in aggregated transfer factor (T_ag) of wood reported in Japan after the Fukushima nuclear accident. (a) Japanese cedar, (b) Japanese cypress, (c) Japanese red pine, (d) konara oak (Source: Adapted from IAEA , TECDOC-1927 [9])

In the study on the transfer of forest products from the Fukushima nuclear accident, a new index has been proposed to substitute for the T_ag. The T_ag takes the surface deposition as the denominator, whereas the denominator of the new index is the initial total deposition of the entire ecosystem (Fig. 3.11). The term “normalized concentration” is used in this document. After the accident at the Fukushima Daiichi Nuclear Power Plant, monitoring surveys of radiation dose and radioactivity levels from the sky using aircraft were quickly and repeatedly carried out. The observation results were published as maps and widely used by researchers. To calculate the T_ag, the accumulation of radiocesium in soil has to be measured. However, it is not easy to conduct ground surveys and collect soil samples in a wide area, especially in forested areas. Therefore, normalized concentration was devised as an index to evaluate the absorption characteristics of radiocesium by trees or forest products by using data from airborne monitoring and taking the initial total deposition of the entire forest as the denominator to eliminate the effects of different contamination levels between measurement points. The unit is the same as the T_ag. After enough time has passed (e.g. after early phase) and most of the radiocesium in the forest ecosystem has transferred to the soil , the normalized concentration and the T_ag will become close to each other. Although there is a limitation that the resolution of the currently available airborne monitoring survey is quite large (250 m mesh), it is an effective method for utilizing and comparing radiocesium data in forest products in places where soil radioactivity is not known.

Fig. 3.11

Difference in denominators when determining the aggregated transfer factor and normalized concentration (Over time, most of the radiocesium in the forest ecosystem will transfer to the soil , then there will be little difference)

The T_ag makes it possible to compare and compile radiocesium concentration data of trees measured at various locations without the influence of the different contamination level of the location. In addition, it can also be used to calculate the approximate radiocesium concentration of trees at a location by multiplying the T_ag (e.g., 1 × 10⁻³ m²/kg) by the accumulated amount at the location (e.g., 100 kBq/m²).

3.6 Migration of Radiocesium out of the Forest

Forests have the ability to retain radiocesium.

3.6.1 Radiocesium Rarely Leaves the Forest

Stream water flowing through forests contains small amounts of fragmented fallen leaves and various substances dissolved in the water, which flow out of the forest system with the water, especially during large runoff events. According to research conducted after the Chernobyl nuclear accident, forests are said to retain radiocesium in the forest ecosystem and prevent it from flowing out of the system. However, since the topography of forests in Japan is much steeper than that of the areas surrounding Chernobyl, and since a large amount of rain often falls at once due to typhoons and storms, there was a concern that radiocesium could flow out of forests into rivers and farmland. Therefore, even after the Fukushima nuclear accident, a lot of monitoring has been conducted in mountain streams and downstream rivers. As a result, radiocesium was detected in streams and rivers, and the radiocesium concentrations were high immediately after the accident and decreased exponentially. It was also found that the rate of decrease was rapid in the early phase and slowed down after about a year had passed [41, 42]. The discharge characteristics of radiocesium are modeled by two exponential functions with different slopes.

Figure 3.12 shows the flow rate of stream water and the radiocesium concentration in the water observed in a stream in Koriyama City, Fukushima Prefecture, the year after the accident. On the second day of the observation, there was rainfall, which increased the amount of water in the stream, and the amount of water decreased as the rainfall stopped. As the rainfall and stream water increased, the radiocesium concentration in the stream water increased. During normal times, before the increase in stream water, the radiocesium concentration in the stream water was below the detection limit, and after rainfall when the flow rate decreased, the concentration dropped below the detection limit again. This observation indicates that radiocesium is released from the stream when the stream water volume increases due to relatively heavy rainfall. This increase in the radiocesium concentration was not caused by radiocesium in the form of ions dissolved in the stream water, but by radiocesium attached to relatively large particles of soil and organic matter (suspended matter) in the turbid water. The fact that the peak in the radiocesium concentration appeared before the peak in the flow rate in this observation case can also be considered to be due to the fact that the particles with higher concentration flowed through first [43,44,45,46].

Fig. 3.12

Activity concentration of cesium-137 in stream water and water discharge observed in a stream in Fukushima Prefecture (Source: Adapted from Shinomiya et al. 2013 [45])

It is known that the discharge of radiocesium outside the forest system is also affected by forest disturbance. Nishikiori et al. conducted similar long-term observations and found that deforestation (clear-cutting) caused a slight increase in the discharge of radiocesium [46]. The discharge of radiocesium into stream water after major forest modification, such as deforestation or forest decontamination, can be considered to be affected by the distance between the stream and the forest area, as well as the point at which the water is collected.

When monitoring stream water in these contaminated areas, radiocesium is detected in the stream water when the water rises and causes turbidity, although the amount is much smaller than immediately after the accident. However, it is also important to look at runoff outside the forest as a percentage of the total forest accumulation. Comparing the amount of radiocesium discharged from the forest with the amount of radiocesium accumulated in the entire stream basin, it has been found that the annual discharge rate is less than 1% of the amount of radiocesium accumulated in the entire basin, even during the relatively early phases of the accident when radiocesium moves easily (e.g., Ministry of the Environment, BOOKLET to Provide Basic Information Regarding Health Effects of Radiation, “Chap. 4: Concept of Radiological Protection, 4.4 Long-term Effects, Behavior of Radioactive Cesium in the Environment: Outflow from Forest Soil ” [1]). It can be said that the amount of radiocesium flowing out of forests is very small compared to the total amount of radiocesium deposited in forests, even in Japan, where slopes are steep and rainfall is frequent and often intense. However, in recent years, large-scale floods caused by record-breaking torrential rains have been occurring frequently. It will be necessary to continue monitoring closely to see how much the runoff increases in such cases.

3.6.2 Little Radiocesium Re-Scattered by Forest Fires in Fukushima

As we have seen, forests have the property of retaining radiocesium while circulating a small part of it in the forest ecosystem. Forest fires have been attracting attention as a potential source of re-scattering (resuspension/redistribution) of radiocesium in the forest. When forest fires occur, soil surface organic layers and trees on the ground surface burn and generate airborne dust, and radiocesium, which has a low boiling point (671 °C), becomes a gas that can be released into the atmosphere and spread to other areas. In addition, burning of combustible materials in the forest will change the distribution of radioactive materials within the forest.

Studies conducted after the Chernobyl nuclear accident, both in artificially generated fire experiments and during actual forest fires, have confirmed the possibility that fires can re-scatter forest radiocesium into the air [47,48,49,50]. For example, two large fires in the Chernobyl exclusion zone in 2015 were estimated to have released 10.9 TBq of cesium-137 into the atmosphere [48]. In 2016, fires also broke out in the Red Forest, where the trees had died of high radiation during the 1986 accident [51].

In Fukushima after the accident, forest fires have occurred several times (Fig. 3.13). Kaneko et al. [52] investigated the burned areas of forest fires that occurred between April 29 and May 10, 2017 at Jumanyama mountain (in Namie and Futaba Towns and located approximately 10 km west of Fukushima Daiichi Nuclear Power Plant) in Fukushima Prefecture, and found that the radiocesium concentration (sum of cesium-134 and -137) in the organic layer on the ground surface was higher in the burned areas than in the unburned areas, although the number of surveyed points was limited. This can be considered to be due to the fact that the organic layer on the ground surface, which is made of dry organic matter, is easily burned, and the radiocesium concentration is higher due to the decrease in volume caused by combustion. In addition, there was concern that the loss of the organic layer, which has the function of preventing runoff and erosion of the surface soil , could result in the release of radioactive cesium from the forest. However, no significant movement or outflow of radiocesium within the forest has been observed [53]. It was also reported that no significant changes were observed in the air dose rate records at nearby monitoring posts [54].

Fig. 3.13

The site of a forest fire in Minamisoma City on April 3, 2016. Taken on April 11, 2016 (Courtesy of Shinji Kaneko, the Forestry and Forest Products Research Institute)

Forest fires can change the cycling, distribution, and fixation of radiocesium in forests, and increase the risk of re-scattering and redistribution, but the actual risk can be considered to vary depending on conditions such as the size (area) of the fire, fire intensity, distance from rivers, and distance from living areas. In addition, when forest fires occurred in Fukushima, some residents expressed concern about the spread of radiocesium. Therefore, it is important to steadily conduct monitoring at the sites of forest fires that have already occurred to clarify the actual situation and dispel the concerns of residents.

Globally, most forest fires are caused by natural causes, such as lightning strikes, mainly in arid regions, while forest fires in humid Japan are mostly caused by human activities. The risk of forest fires and their re-scatter may not be so high in the difficult-to-return areas in Fukushima because the general public is not allowed to enter these areas. On the other hand, if the amount of combustible materials such as fallen leaves and dead trees on the forest surface increases due to the loss of forest management, the risk of fire spread and intensity may increase. It has also been pointed out that climate change may cause the death of trees and inhibition of decomposition of organic matter on the ground surface, resulting in an increase in combustible materials on the ground surface [47]. It is important to prevent forest fires from occurring in the first place, but it is also essential to be prepared to extinguish fires in the early phases and prevent them from spreading if they do occur. In addition, in efforts to prevent the occurrence and spread of forest fires in contaminated areas, it is essential to carefully assess the risk of re-scattering in the event of a forest fire.

3.7 Predicting the Future Distribution of Radiocesium in Forests

Forecasting can be a powerful tool if used with an understanding of uncertainty.

3.7.1 Reproduction and Prediction by Computer Simulation

As we have seen, radiocesium is moving in the forest, and its distribution in the forest changes with time. Models are often used to capture the behavior of radiocesium in the forest. A model is a mathematical expression that describes and represents a natural phenomenon by dividing it into a number of elements and processes. Analysis using models is called modeling or simulation . After the Fukushima nuclear accident, research was also conducted to analyze and predict the contamination of forests through modeling [55,56,57,58,59,−60]. Modeling of radiocesium in forests is a dynamic representation of the behavior of radiocesium in the forest by connecting the accumulated amount of radiocesium in each part of trees and soil with the amount of movement (flux) connecting these parts. For example, the radionuclide cycle model in forests (RIFE1 model) [57, 64], which was developed after the Chernobyl nuclear accident [61,62,63], was used in Fukushima studies. The model represents a forest by dividing it into leaves, branches, and bark that have undergone direct deposition , and wood that has not undergone direct deposition , the organic layer on the soil surface, and mineral soil, and the behavior of radiocesium in the forest is reproduced (Fig. 3.14).

Fig. 3.14

Example of the structure of a radionuclide cycling model in forest (modified RIFE1 model) (Source: Adapted from Hashimoto, et al. [64])

What controls the behavior of a model is the structure of the model and the parameters that specify the amount of movement and accumulation. In general, a lot of observational data is used in building models to control the behavior of the model and to verify the model results. On the other hand, it is difficult to comprehensively study radiocesium in the forest through field observations but modeling can compensate for these difficulties and evaluate the movement of radiocesium, which allows us to understand the behavior of radiocesium in a more integrated way. Future predictions can also be made through simulations. Hashimoto et al. simulated the behavior of radiocesium in Japanese cedar and konara oak forests by adjusting the parameters of the RIFE1 model described above using data observed in Fukushima [57, 64]. As a result, it was predicted that most of the radiocesium in the forests would transfer to the soil after the second year after the accident, and that this state would continue in the future (Fig. 3.15). As a result of predicting the radiocesium concentration in the wood of Japanese cedar and konara oak for 20 years after the accident, it was predicted that there would be no change to a slight decrease in the radiocesium concentration in Japanese cedar, and that the increasing trend in the radiocesium concentration seen after the accident would cease in konara oak (Fig. 3.16).

Fig. 3.15

Predicted distribution of radiocesium in a Japanese cedar forest (Source: Adapted from Hashimoto et al. [64])

Fig. 3.16

Future prediction of cesium-137 concentration in wood of (a) Japanese cedar and (b) konara oak (Source: Adapted from Hashimoto et al. [64])

To quantify the transfer of radiocesium in the forest, the amount of absorption by trees and transfer from trees to soil were analyzed from the output results of the model (Fig. 3.17). As a result, it was suggested that the transfer of radiocesium from trees to soil (emission) and the absorption of radiocesium from soil by trees would be balanced in 5–10 years after the accident, and that the cycling of radiocesium in the forest would move toward a steady state (equilibrium). On the other hand, it was also suggested that even after the steady state was reached, less than 1% of the total deposited amount of radiocesium in the forest immediately after the accident could continue to circulate in the forest.

Fig. 3.17

Absorption and emission of radiocesium in the forest calculated by the model. Movement is shown as a percentage of the total amount in the forest immediately after the accident (Source: Adapted from the Forestry and Forest Products Research Institute, Collection of Research Results 2020 fiscal year, “Predicting the Movement of Radiocesium in Forests” [66])

In this way, modeling can clarify the important processes and the range of parameters involved in understanding the phenomena. By reflecting this information in monitoring, it will be possible to further enhance and develop observations. Several modeling studies have been conducted in the forests of Fukushima [55,56,57,58,59, 64]. It should be noted that there are differences in the prediction results depending on the structure of the model. Hashimoto et al. compared the prediction of radiocesium concentration in wood in Japanese cedar forests using the FoRothCs model developed by Nishina et al. [56, 58] with the RIFE1 model, and it was shown that the FoRothCs model resulted in a larger decrease than the RIFE1 model, although the two models were generally comparable. Such a comparison using multiple models was also made during the Chernobyl nuclear accident [65]. We also believe that it is important to constantly update future predictions by adding new data to verify that the predictions are correct. The close linkage between modeling and monitoring (observation data) is a useful way to deepen our understanding of the phenomena and improve the accuracy of future predictions.

3.7.2 Future Predictions of Air Dose Rates

As will be explained in detail in Chap. 6, air dose rates in forests are generally consistent with the predictions of decreases due to radioactive decay. Therefore, Fukushima Prefecture has been predicting how air dose rates in forests will change 15 and 25 years after the accident using data from multipoint surveys of air dose rates that have been conducted every year since the accident (Fig. 3.18) [67]. As a result, looking at the average air dose rate of 362 locations where surveys have been conducted continuously since 2011, the survey results for August 2011 show an average of 0.91 μSv/h, and those for March 2020 (9 years after the accident) show 0.20 μSv/h. Furthermore, the prediction based on the data for the first 9 years shows that the air dose rate will be 0.15 μSv/h in 2026 (15 years after the accident) and 0.12 μSv/h in 2036 (25 years after the accident). It should be noted that cesium-134 (half-life: 2 years), which was emitted along with cesium-137 during the accident, rapidly decays in the first few years after the accident, resulting in a significant decrease in air dose rates during that period, but after that the decrease will be slower due to the fact that only cesium-137 (half-life: 30 years) will remain (see Chap. 6). However, air dose rates have been steadily decreasing over time. Such a steady decrease in air dose rates can be seen in highly contaminated areas as well. Figure 3.19 shows the results of future predictions of air dose rates not only for forests but also for the region including the areas with high air dose rates extending northwest around the Fukushima Daiichi Nuclear Power Plant [68]. It can be seen that the contaminated areas are shrinking and the air dose rates are decreasing even in the areas with high air dose rates. On the other hand, it also shows that air dose rates in this northwestern extension of the region will still be high 30 years after the accident.

Fig. 3.18

Future predictions of spatial distribution of air dose rate in the forest in Fukushima (Source: Data from Fukushima Prefecture’s “Results of Radioactive Materials Surveys in Forests in 2019 fiscal year” [67], with a decrease due to radioactive decay applied based on the survey results for 2019 fiscal year)

Fig. 3.19

Future predictions of air dose rate distributions within 80 km of the Fukushima Daiichi Nuclear Power Plant (Source: Data from the observation of 1st, 7th 12th Airborne Monitoring Survey [3] and the prediction of Kinase, et al. [68]; Data of prediction : Courtesy of Sakae Kinase, Japan Atomic Energy Agency)

3.7.3 How Should We Deal with the Predictions?

Both the dynamics of radiocesium in forests and the prediction of air dose rates seen in this chapter are based on actual measurement data after the accident, and some kind of model is used to predict the future state. Such predictions are highly dependent on actual measurement data, and in the absence of sufficient measurement data, we may make predictions that deviate greatly from reality. Therefore, it is necessary to continue to conduct surveys at least over the next few decades to verify and monitor whether the predictions made by the models are correct, and to update the predictions as needed. There is also the problem that wide-area assessments do not sufficiently represent heterogeneity among locations, for example, they sometimes do not reflect minor differences in topography, forest types, and deposition . Furthermore, it is well known that the radiocesium concentration in various forest products varies widely even within the same study area. Nevertheless, understanding and predicting radiocesium concentrations and air dose rates over such a wide area is very important in terms of providing an overall picture and outlook of the problem of radioactive contamination over such a wide area as forests. The prospect of future contamination levels of the forest environment and forest products is of utmost concern to the residents of the affected areas and to those who are trying to restart and rebuild the forest industry. Information on the uncertainty of the forecast should be also provided, and it is important to take this into account when using the forecast information.

3.8 To Summarize the Behavior of Radiocesium in the Forest

In this chapter, we have looked at the distribution and transfer of radiocesium in the forest by part of the forest and by process. The overall picture can be summarized as follows.

As mentioned in the beginning of this chapter, in forests, the trees first trap radiocesium. Within a few months to a few years, the radiocesium transfers to the ground surface and then to the mineral soil. Depending on the forest, most of the radiocesium transfer to the soil and accumulate within a few years. A small percentage of the total amount of radiocesium in the forest is absorbed by the trees, and eventually returns to the ground surface as litterfall. Most of the radiocesium remains in the surface layer of the soil . It is also clear that the amount of radiocesium leaving the forest through stream water is very limited compared to the total amount accumulated in the watershed. Figure 3.20 shows the distribution of radiocesium observed in a cedar forest in 2011, 2012, and 2017. It shows the distribution of radiocesium inventory per unit area in the forest, and shows that radiocesium, which was trapped in large amounts in leaves and branches in 2011, decreased significantly after 1 year, and by 2017 most of it had moved to mineral soil. The behavior of radiocesium in the forests of Fukushima is about to enter the phase of quasi-equilibrium and equilibrium from the early phase of contamination, when the radiocesium moved significantly within the forests.

Fig. 3.20

Changes in radiocesium distribution in a cedar forest (example of a Japanese cedar forest). (a) 2011, (b) 2012, (c) 2017 (Source: Data from a survey by the Forestry Agency and the Forestry and Forest Products Research Institute, “Surveys on Radioactive Cesium Distribution in Forests” [21])