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Availability and immobilization of 137Cs in subtropical high mountain forest and grassland soils
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Journal of Environmental Radioactivity 99 (2008) 882e889 www.elsevier.com/locate/jenvrad
Availability and immobilization of 137Cs in subtropical high mountain forest and grassland soils C.-Y. Chiu a,*, C.-J. Wang b, C.-C. Huang b a
Research Center for Biodiversity, Academia Sinica, Taipei, Taiwan b Taiwan Radiation Monitoring Center, Kaoshiung, Taiwan Accepted 9 November 2007 Available online 27 December 2007
Abstract To understand the behavior of 137Cs in undisturbed soils after nuclear fallout deposition between the 1940s and 1980s, we investigated the speciation of 137Cs in soils in forest and its adjacent grassland from a volcano and subalpine area in Taiwan. We performed sequential extraction of 137Cs (i.e., fractions readily exchangeable, bound to microbial biomass, bound to FeeMn oxides, bound to organic matter, persistently bound and residual). For both the forest and grassland soils, 137Cs was mainly present in the persistently bound (31e41%) and residual (22e62%) fractions. The proportions of 137Cs labile fractions e bound to exchangeable sites, microbial biomass, MneFe oxides, and organic matter e were lower than those of the recalcitrant fractions. The labile fractions in the forest soils were also higher than those in the grassland soils, especially in the volcanic soil. The results suggest that the labile form of 137Cs was mostly transferred to the persistently bound and resistant fractions after long-term deposition of fallout. The readily exchangeable 137Cs fraction was higher in soils with higher organic matter content or minor amounts of 2:1 silicate clay minerals. Ó 2007 Elsevier Ltd. All rights reserved. Keywords:
137
Cs; Microbial biomass; Organic matter; Soil
1. Introduction Organic matter plays a critical role in the bioavailability of 137Cs in the soil, because organic matter reduces adsorption and subsequent fixation of 137Cs on clay minerals (Barber, 1964; Valcke and Cremers, 1994), maintaining 137 Cs in a labile, easily exchangeable form. Transfer of Cs from soil to edible plant part increases with increasing organic matter content (van Bergeijk et al., 1992), which potentially threatens human health through the food chain. However, the interaction between the soil mineral and organic phase with 137Cs availability is complicated. Some studies suggest that 137Cs is controlled only by the mineral phase, even in soils with high organic matter content (Shand et al., 1994; Rigol et al., 1999). In addition, many studies suggest that microflora contain a critical pool of 137 Cs in organically rich soils (Dighton et al., 1991; Simkiss et al., 1993; Sanchez et al., 2000). Microflora actively transport 137Cs from organic horizons to fresh litter, a key function in the recycling and persistence of 137Cs in forest soils (Bru¨ckmann and Wolters, 1994; Rafferty et al., 1997; Vinichuk et al., 2005). Fukuyama and Takenaka (2004) * Corresponding author. E-mail address: [email protected] (C.-Y. Chiu). 0265-931X/$ - see front matter Ó 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.jenvrad.2007.11.015
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found that microbial activity plays a critical role in upward migration of 137Cs from soil to litter and might induce the removal of 137Cs with litter from the forest through runoff. Speciation of 137Cs is used to characterize the chemical binding of 137Cs to the different constituents of soil. Most studies (Shand et al., 1994; Baumann et al., 1996; Rigol et al., 1999; Sanchez et al., 2000) investigated the fate of 134Cs or 137Cs in different soil phases only in the short-term, or a few years after the Chernobyl deposition. Some other studies looked at local 137Cs contamination, such as the release of 137Cs due to accidents at nuclear facilities (Wasserman et al., 2002). By comparison, relatively few studies have investigated soil stabilized for a long time. Despite the general recognition that the amount of nuclear fallout from weapon testings in the 1960s is lower in areas of lower latitude (Hardy, 1968), large amounts of 137Cs deposited into a subtropical mountain forest ecosystem in Taiwan still remain in the surface soils (Chiu et al., 1999a,b). However, the deposition due to the reactor accident at Chernobyl in 1986 was negligible in these areas (Lin and Huang, 1988) and is appropriate for observing characteristics of 137Cs in soils after long-term deposition. Our previous study of organic soils from an undisturbed Chamaecyparis forest in northeastern Taiwan showed that the labile form of 137Cs is mostly transferred to the persistently bound and resistant fractions after long-term deposition of fallout (Chiu et al., 2002). In this new study, we selected two new sites with soils of different pedogenetic origin to compare the bioavailability and behavior of 137Cs in soils between forest and grassland, where radiocaesium has been stabilized for 40 years. 2. Material and methods 2.1. Study sites One of the two study sites is located at a volcano, Yamming Mountain (25 90 N, 121 340 E), in the northern region of Taiwan at approximately 800 m above sea level (Fig. 1). The average annual temperature is approximately 16.7 C and the evenly distributed annual precipitation approximately 4900 mm. The vegetation is basically evergreen broadleaf forest dominated by broad-leaved Lauraceae and Fagaceae trees. Adjacent grassland, dominated by silver grass (Miscanthus floridulus) and dwarf bamboo (Yushania usawai), stretches wide because of infrequent fires. The soils are Andosols formed on andesitic pyroclastic rocks. They exhibit andic properties with very high aluminum saturation. Kaolinite, halloysite, gibbsite, and quartz dominate the clay mineral assemblage, with minor amounts of 2:1 silicate clays. Aluminum toxicity and phosphorus deficiency are believed to contribute to humus accumulation (Chen et al., 1999). The soils of the forest and grassland have distinct litter layers. The slope is 15%. The soil contains about 8e12 cm of a humified Ah horizon underneath the litter layer. The second site is in a subalpine forest reserve at Tatajia (23 280 N, 120 540 E), in the saddle of the Jade Mountain, in central Taiwan, at 2700 m above sea level. The climate is mountainous temperate, with a mean annual precipitation and temperature of approximately 3170 mm and 11.1 C, respectively. The mountainous forests have a tall coniferous canopy of hemlock (Tsuga chinensis), a broad-leaved tree (Trochodendron aralioides) understory, and a ground layer of dwarf bamboo (Yushania niitakayamensis). The dominant species in the grassland are Miscanthus transmorrisonesis and Y. niitakayamensis. A Yushania-dominated zone adjacent to the forestland forms a transition zone or ecotone between the grassland and forest. The origin of the grassland is still unknown, although it might have come from natural succession of the vegetation with elevation or from infrequent fires. Nevertheless, the vegetation has not changed for several decades. The soils of the investigated system are partially podozolized Spodozols (loamy Umbric), with a distinct litter layer, strongly acidic and a steep slope of 25%. Besides the litter layer (Oi horizon), the forest soil contains approximately 13e20 cm of humified organic matter (Oe and Oa horizon) above the mineral horizons. Kaolinite and illite dominate the clay mineral assemblage, with minor amounts of vermiculite. The surface humified mineral layer (Ah horizon) in the grassland soil is deeper than that in forest soil. 2.2. Soil After carefully removing litter, soil samples were taken from the O/A horizons in June 1999. Soil samples were 10 cm deep collected with use of a soil auger 8 cm in diameter. Six samples were combined for each soil type. The samples of fresh soil were passed through a 2-mm sieve and stored in polyethylene bags at 4 C in moist state before use. The soil samples had high organic matter content and cation exchange capacity (CEC) values. The base saturation values were low, contributing to low pH values. The radioactivity level of 137Cs was much higher in volcanic soils than subalpine soils (Table 1). The activity of 137Cs was associated with geographic distribution, because we found in early studies (Chiu et al., 1999a) and elsewhere that the northern mountain areas received more precipitation from mainland monsoons in winter than other seasons and accumulated more nuclear fallout in the soils than other areas.
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Fig. 1. Map of sampling sites. Symbols and indicate the locations of collecting samples for grassland and forest, respectively. The altitude for the two locations is shown using shading and a legend.
Table 1 Chemical properties of soil samples used for sequential extraction of 137Cs Properties pHa CEC (cmol kg1) Base saturation (%) Total organic C (%) Radioactivity of 137Cs (Bq kg1)
Subalpine
Volcanic
Forest
Grassland
Forest
Grassland
3.6 0.1 74.8 1.9 5.1 0.3 39.6 2.6 16.7 1.3
4.1 0.1 34.3 0.4 2.6 0.1 13.7 0.3 37.0 1.7
3.6 0.1 37.7 2.0 1.6 0.3 9.3 0.3 121.5 4.5
3.5 0.1 40.3 2.1 1.1 0.2 8.2 0.2 141.3 6.2
Average of three samples standard deviation of the mean. a Water:soil ¼ 2:1.
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2.3. Sequential extraction of
137
885
Cs
Sequential extraction involved 100 g of soil. The procedure basically followed that of Tessier et al. (1979), but the extraction for the carbonate-bound fraction was omitted because of the low pH of the studied soils. The selective sequential extraction procedure was modified according to Baumann et al. (1996). Fraction I is called ‘‘the exchangeable fraction’’, which was determined by extraction with K2SO4 solution from unfumigated soil. Fraction II is called ‘‘the microbial biomass fraction’’, which was determined as the difference in extraction between chloroform-fumigated and unfumigated treatment. Soils were fumigated for 24 h with ethanol-free CHCl3. After the fumigant was removed, the soils were then extracted with 0.5 M K2SO4. The non-fumigated control soils were extracted under the same conditions at the time fumigation commenced. Microbial biomass Cs (Csmic) was calculated as follows: Csmic ¼ [(Cs extracted from fumigated soil) (Cs extracted from non-fumigated soil)]/KCs. The KCs factor (the extractable part of microbial biomass Cs) was 0.5 as proposed by Bru¨ckmann and Wolters (1994). The soil samples left after the fumigation extraction were used for further steps of sequential extraction. Fraction III is called ‘‘the FeeMn oxide bound fraction’’, which was determined by extraction with 1000 ml of 0.1 M NH2OH$HCl in 0.01 M HNO3 for 5 h at 90 C. Fraction IV is called ‘‘the organic matter-bound fraction’’, which was determined from the residue of fraction III. The residue was resuspended with 100 ml of water and heated at 80 C. An amount of 1000 ml of 30% H2O2 (pH 2.0, adjusted with HNO3) was gradually added while stirring. After cooling, 250 ml of 3.2 M NH4AC in 20% HNO3 was added and agitated for 30 min. The liquid phase contains the desired fraction. Fraction V is called ‘‘the persistently bound fraction’’, which was determined by treating the residue of fraction IV with 1000 ml of 7 M HNO3 and stirring for 6 h at 90 C. The liquid phase contains the desired fraction. Fraction VI is called ‘‘the residual solid fraction’’ resulting from fraction V neutralized with NaOH and dried at 100 C. For each of these extraction steps, the supernatant was removed by centrifugation for 10 min at 10,000 g. The residue of each extraction was washed with 50 ml of distilled water followed by centrifugation and combined with the original supernatant. Centrifuge bottles (250 ml) were used to facilitate centrifuge washing of the samples after each extraction to minimize loss of material. The supernatant of each fraction was filtered and evaporated by heating either to complete dryness or to a minimum volume (<20 ml). The concentration of 137Cs was measured by direct gamma-spectrometry by use of a high-purity Ge detector coupled to a computerized data acquisition system (Canberra Series 95, 4096 channel pulse height analyzer), with Canberra Genie PC software used for spectrum analysis. The efficiency calibration involved use of a standard multi-gamma-ray source mixed in agar. The counting time varied from 80,000 to 320,000 s. The minimum detectable activity of 137Cs was approximately 0.2 Bq. 2.4. Separation and determination of organic matter The water soluble organic carbon (WSOC) fraction was determined by extracting 1 g soil with use of Millipore Milli-Q H2O at a 1:10 ratio of soil to water, followed by shaking at 200 rpm for 16 h at room temperature. The extract was decanted, then centrifuged at 10,000 g for 30 min, and the supernatant was filtered through a Whatman No. 42 filter paper. Soil microbial biomass carbon was measured by fumigation extraction of fresh soil samples. Microbial biomass C (Cmic) was calculated as follows: Cmic ¼ [(organic C extracted from fumigated soil) (organic C extracted from non-fumigated soil)]/KC. The KC factor (the extractable part of microbial biomass C) was 0.45 (Wu et al., 1990). The process for extraction of soil humus basically followed that of Tan (1985). One gram of air-dried soil was extracted first with use of 0.1 M HCl to remove non-humic substances, then the residue was extracted with use of 0.1 M NaOH (with a soil-to-extractant ratio of 1:20) under N2 at room temperature for 16 h. The alkaline supernatant was separated from the residue by centrifugation at 15,000 g for 20 min, then acidified with 6 M HCl to pH 1.0 and allowed to stand at room temperature for 24 h. The supernatant (fulvic acids) was separated from the coagulated material (humic acids) by centrifugation at 8000 g for 15 min. The humic acid fraction was then purified by use of a HCleHF mixture, then the residue was washed thoroughly with distilled water and redissolved with 0.1 M NaOH. Each extracted fraction was analyzed for total organic carbon (O. I. Analytical 1010) by the heatepersulfate oxidation method. Soil organic matter was measured with use of a Fisons NA1500 elemental analyzer.
3. Results and discussion The composition of labile forms of C and extractable humic substances differed greatly among studied soils in vegetation and pedogenic characteristics (Table 2). Microbial biomass C content was higher in forest than in grassland soils. Microbial biomass C, fulvic and humic acids were associated with pedogenic characteristics, and their content
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Table 2 Ratios of each fraction of C to total organic C in studied soils (%) Fraction
Subalpine
Water soluble C Microbial biomass C Fulvic acids Humic acids
Volcanic
Forest
Grassland
Forest
Grassland
0.4 0.0 2.4 0.1 3.8 0.3 7.4 0.5
1.7 0.1 1.3 0.1 7.0 0.4 10.6 0.7
1.4 0.1 0.7 0.0 12.2 0.8 27.5 2.5
0.2 0.0 0.4 0.0 9.9 0.6 22.4 1.4
Average of three samples standard deviation of the mean.
was higher in volcanic than in subalpine soils. Microbial biomass C and WSOC contributed a small portion of the total organic C in all soils. Most of the 137Cs was bound to persistent and residual fractions. For both the forest and grassland soils, 137Cs was mainly present in fraction V, the persistently bound fraction (31e41%), and fraction VI, the residue fraction (22e 62%). Only trace amounts of 137Cs existed in the readily exchangeable fraction (Fig. 2). The proportions of the labile fractions, such as that bound to exchangeable sites, microbial biomass, MneFe oxides, and organic matter (fractions IeIV), were much lower than those of the recalcitrant Vand VI fractions. The labile fractions were higher in forest than in grassland soils. Subalpine forest soil showed a different pattern than the other sites; the persistent and residual fractions contained less 137Cs, which could be attributed to the enormous amount of organic matter (Table 1). Nevertheless, only minor proportions of 137Cs were found in the organic matter fraction of these soils; only 15% was found even in subalpine forest soil, the other sites having even smaller values (1e4%).
80 Forest
a
Grassland
(%)
60
40
20
0
Exch.
Microb.
Oxides
OM
Perst.
80 Forest
Resd.
b
Grassland
(%)
60
40
20
0 137
Exch.
Microb.
Oxides 137
OM
Perst.
Resd.
Fig. 2. Fraction of fallout Cs in studied soils. Recovery of Cs activities from the summation of sequential extraction fractions to bulk soils were within 15% deviations. Vertical bars indicate standard deviation of two replicate measurements. (a) Subalpine soil (b) Volcanic soil.
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In organic soils, the fraction of Cs absorbed in organic sites constitutes a labile, exchangeable pool (Valcke and Cremers, 1994). Radiocaesium is much less efficiently bound and is correspondingly much more mobile in organic soils than in more clay-rich soils. Tegen et al. (1991) suggested that the newly deposited 137Cs is initially bound to organic matter in the early stages of decomposition and is more labile than Cs bound to organic matter produced during the decomposition process. The latter solid species are more resistant in the soil. Compared to results of relatively short-term deposition (Baumann et al., 1996; Rigol et al., 1999), our soils showing long-term deposition had lower amounts of 137Cs bound to exchangeable, microbial biomass or organic matter fractions but higher amounts as persistent and residual fractions, which might be due to the well-known ‘‘aging effect’’ (Absalom et al., 1996; Spezzano, 2005). Recent study of soils affected by the Chernobyl accident indicate that most of the 137Cs is in the residual fraction (Hou et al., 2003), which supports this aging phenomenon. The role of soil microbial biomass in 137Cs retention may be important in highly organic soils (Sanchez et al., 2000). Several studies have reported the upward movement of 137 Cs in forest soil, where fungal activities play key role in translocation of 137Cs (Bru¨ckmann and Wolters, 1994; Rafferty et al., 1997). Being aware of the importance of the availability of 137Cs retained by fungi, some investigators tried to determine the fraction of nuclear fallout 137Cs in fungal hyphae in soil (Bru¨ckmann and Wolters, 1994; Baumann et al., 1996). Recent study indicates that a critical part of fungal mycelia and fruit bodies of mycorrhizal fungi are water soluble (Vinichuk et al., 2005), which contributes to the mobility of 137Cs in the forest soils. Although the fraction of fungal biomass-bound 137Cs is considered a labile form in soil, some studies (Baumann et al., 1996) found that the conventional technique for measuring microbial biomass might underestimate fungal biomass 137Cs, because 137Cs released from microorganisms after chloroform fumigation could be easily fixed by clay minerals, even in organic soil. This observation could explain in part why our volcanic soils contained a relatively higher proportion of 137Cs bound to microbial biomass than the subalpine soils (Fig. 2). Volcanic soils in this area have minor amounts of 2:1 silicate clays (Chen et al., 1999), the mineral associated with permanent 137Cs fixation, thus enabling higher potential for nutrient recycling. In addition, microbial biomass C content was lower in volcanic soils than in subalpine soils (Table 2), which coincided with the proportion of 137Cs and suggested higher microbial activities in volcanic soils. Microbial biomass-bound 137Cs contributes only a minor portion of the total amount of 137Cs, which is consistent with the results of microbial biomass C in these study sites (Table 2) and those of high mountain soils in Taiwan (Chen and Chiu, 2000; Imberger and Chiu, 2001). A great many studies have investigated the diversity of soil microbial biomass in different ecosystems (Wardle, 1998). Although the form of land use is expected to affect the results of conversion of measured data into microbial biomass in different microbial community structures, Joergensen (1996) confirmed that the KC values (0.45) of Vance et al. (1987) and Wu et al., (1990) remained valid, and this value is similar to the KCs value (0.5). Bru¨ckmann and Wolters (1994) proposed to determine microbial biomass Cs. Regardless of the possible underestimation of the microbial biomass fraction of 137Cs (Baumann et al., 1996), the apparently negligible amounts in labile forms supports the general concept that the microbial biomass fraction may act as a pinhole, which supports the recycling of this element within the soil and in the plantesoil system. Organic substances, particularly fulvic acids, play a critical role in the migration of some metal ions in the soil (Saar and Weber, 1982). Our other studies showed that fulvic acids migrate downward in the soil horizons under perhumid forest conditions (Chen and Chiu, 2000), but organic substances do not seem to enhance the leaching of 137Cs to the deeper mineral horizons (Chiu et al., 1999a). This phenomenon is consistent with the observation that interaction of Csþ with organic substances is weak (van Bergeijk et al., 1992). Rigol et al. (1996) also suggest no relation between humic acid content and 134Cs desorption. Shand et al. (1994) showed that oxidizable organic matter, corresponding to humin, strongly retained 137Cs. The real humic substances fraction can barely be defined with the sequential extraction procedure in this study because of the different separation systems. If we consider that humic and fulvic acids, the extractable humic substances, are in moderately labile form, they could be partially associated with the fraction bound to organic matter, and humin is the persistent fraction. The results in Fig. 2 suggest that fulvic and humic acids contribute only in a minor way to 137Cs recycling. This suggestion is also true for the volcanic soils, because the organic matter-bound fraction is low, although volcanic soils contain high humic and fulvic acid content. In this study site, the noncrystalline forms of Fe and Al are associated largely with humus, which is related to the little or no formation of allophane in A horizons (Chen et al., 1999). Such metalehumus complexes are highly stable and resistant to microbial attack and one of the mechanisms used to retain organic matter in volcanic soils (Shoji et al., 1993). We believe that metalehumus complexes could relate to the persistent fraction or residual fraction of this study. Tree biomass might be an important pool
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of mobile 137Cs stored in the forest ecosystem (Strebl et al., 1999). Through litterfall and decomposition of litter, the 137 Cs returns to soil organic matter. Soil with a large amount of organic matter is expected to fix less 137Cs. Most 137Cs passing labile phases is taken up again by microbial biomass and vegetation, particularly in forest soils, which are covered by a thick litter layer. Such phenomenon explains a certain portion of 137Cs remaining in the labile form in the subalpine forest soil. The proportions of 137Cs we found were relatively higher than those in our previous study, conducted in an undisturbed Chamaecyparis forest (Chiu et al., 2002). Parent material and weathering conditions contribute to differing content of 2:1 clay minerals and consequently affect the fixation of 137Cs. For instance, our recent study showed that vermiculite and illite were dominant in the poorly drained soil of the Chamaecyparis forest (Pai et al., 2007), where 137Cs could be transferred easily to a persistently bound and resistant fraction. Rapid up-lifting and erosive conditions resulted in a relatively young geological age for mountainous Taiwan. In addition, the soil with high clay content is barely drained under humid weather conditions, which retards the weathering of phyllosilicates and retains relatively high 2:1 clay minerals (Pai et al., 2004, 2007). A parallel study in the nearby subalpine area indicated that illite and kaolinite are the major clay minerals in the surface horizon of the forest soil (Pai et al., 2004). Decomposition products of coniferous litter increased the forest floor acidity of the surface horizons and promoted the rapid breakdown of illite. Abundant 2:1 clay minerals (illite and vermiculite) exist in grassland soil (unpublished data), which indicates the lower weathering intensity in grassland than in forest soil. The permanent fixation of 137Cs with 2:1 clay minerals in grassland soil explains why only a small proportion of 137Cs exists in labile phases in grassland soil. By contrast, the difference in 137Cs activity between grassland and forest in volcanic soils was relatively small and might be attributed to the characteristics of soil materials per se, since volcanic soils contain a minor amount of 2:1 clays (Chen et al., 1999). 4. Conclusion The grassland and forest soils of our study sites in Taiwan showed features of a possible transformation of 137Cs from labile pools to persistently bound and resistant fractions after long-term deposition of fallout, although forest soils retained a relatively higher proportion in the labile fractions than did grassland soils. Different pedogenetic origins play a crucial role in the differing bioavailability and behavior of 137Cs between forest and grassland soil. The difference in labile 137Cs fractions between forest and grassland soils seems to be associated with soil organic matter content and seems to be higher in soils with abundant 2:1 clay minerals and less so in volcanic soils. Acknowledgements The study was supported by Academia Sinica and National Science Council grants (NSC 92-2623-7-001-002-NU). References Absalom, J.P., Crout, N.M.J., Young, S.D., 1996. Modeling radiocesium fixation in upland organic soils of Northwest England. Environmental Science and Technology 30, 2735e2741. Barber, D.A., 1964. Influence of soil organic matter on the entry of caesium-137 into plants. Nature 204, 1326e1327. Baumann, A., Schimmack, W., Steindl, H., Bunzl, K., 1996. Association of fallout radiocesium with soil constituents: effect of sterilization of forest soils by fumigation with chloroform. Radiation and Environmental Biophysics 35, 229e233. van Bergeijk, K.E., Noordijk, H., Lembrechts, J., Frissel, M.J., 1992. Influence of pH, soil type and soil organic matter content on soil-to-plant transfer of radiocesium and strontium as analyzed by a nonparametric method. Journal of Environmental Radioactivity 15, 265e276. Bru¨ckmann, A., Wolters, V., 1994. Microbial immobilization and recycling of 137Cs in the organic layers of forest ecosystems: relationship to environmental conditions, humification and invertebrate activity. The Science of the Total Environment 157, 249e256. Chen, J.S., Chiu, C.Y., 2000. Effect of topography on the composition of soil organic substances in a perhumid sub-tropical montane forest ecosystem in Taiwan. Geoderma 96, 19e30. Chen, Z.S., Asio, V.B., Yi, D.F., 1999. Characteristics and genesis of volcanic soils along a toposequence under a subtropical climate in Taiwan. Soil Science 164, 510e525. Chiu, C.Y., Lai, S.Y., Lin, Y.M., Chiang, H.C., 1999a. Distribution of the radionuclide 137Cs in the soils of a wet mountainous forest in Taiwan. Applied Radiation and Isotopes 50, 1097e1103. Chiu, C.Y., Lai, S.Y., Wang, C.J., Lin, Y.M., 1999b. Transfer of 137Cs from soil to plants in a wet montane forest in subtropical Taiwan. Journal of Radioanalytical and Nuclear Chemistry 239, 511e515.
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Journal of Environmental Radioactivity 99 (2008) 882e889 www.elsevier.com/locate/jenvrad
Availability and immobilization of 137Cs in subtropical high mountain forest and grassland soils C.-Y. Chiu a,*, C.-J. Wang b, C.-C. Huang b a
Research Center for Biodiversity, Academia Sinica, Taipei, Taiwan b Taiwan Radiation Monitoring Center, Kaoshiung, Taiwan Accepted 9 November 2007 Available online 27 December 2007
Abstract To understand the behavior of 137Cs in undisturbed soils after nuclear fallout deposition between the 1940s and 1980s, we investigated the speciation of 137Cs in soils in forest and its adjacent grassland from a volcano and subalpine area in Taiwan. We performed sequential extraction of 137Cs (i.e., fractions readily exchangeable, bound to microbial biomass, bound to FeeMn oxides, bound to organic matter, persistently bound and residual). For both the forest and grassland soils, 137Cs was mainly present in the persistently bound (31e41%) and residual (22e62%) fractions. The proportions of 137Cs labile fractions e bound to exchangeable sites, microbial biomass, MneFe oxides, and organic matter e were lower than those of the recalcitrant fractions. The labile fractions in the forest soils were also higher than those in the grassland soils, especially in the volcanic soil. The results suggest that the labile form of 137Cs was mostly transferred to the persistently bound and resistant fractions after long-term deposition of fallout. The readily exchangeable 137Cs fraction was higher in soils with higher organic matter content or minor amounts of 2:1 silicate clay minerals. Ó 2007 Elsevier Ltd. All rights reserved. Keywords:
137
Cs; Microbial biomass; Organic matter; Soil
1. Introduction Organic matter plays a critical role in the bioavailability of 137Cs in the soil, because organic matter reduces adsorption and subsequent fixation of 137Cs on clay minerals (Barber, 1964; Valcke and Cremers, 1994), maintaining 137 Cs in a labile, easily exchangeable form. Transfer of Cs from soil to edible plant part increases with increasing organic matter content (van Bergeijk et al., 1992), which potentially threatens human health through the food chain. However, the interaction between the soil mineral and organic phase with 137Cs availability is complicated. Some studies suggest that 137Cs is controlled only by the mineral phase, even in soils with high organic matter content (Shand et al., 1994; Rigol et al., 1999). In addition, many studies suggest that microflora contain a critical pool of 137 Cs in organically rich soils (Dighton et al., 1991; Simkiss et al., 1993; Sanchez et al., 2000). Microflora actively transport 137Cs from organic horizons to fresh litter, a key function in the recycling and persistence of 137Cs in forest soils (Bru¨ckmann and Wolters, 1994; Rafferty et al., 1997; Vinichuk et al., 2005). Fukuyama and Takenaka (2004) * Corresponding author. E-mail address: [email protected] (C.-Y. Chiu). 0265-931X/$ - see front matter Ó 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.jenvrad.2007.11.015
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found that microbial activity plays a critical role in upward migration of 137Cs from soil to litter and might induce the removal of 137Cs with litter from the forest through runoff. Speciation of 137Cs is used to characterize the chemical binding of 137Cs to the different constituents of soil. Most studies (Shand et al., 1994; Baumann et al., 1996; Rigol et al., 1999; Sanchez et al., 2000) investigated the fate of 134Cs or 137Cs in different soil phases only in the short-term, or a few years after the Chernobyl deposition. Some other studies looked at local 137Cs contamination, such as the release of 137Cs due to accidents at nuclear facilities (Wasserman et al., 2002). By comparison, relatively few studies have investigated soil stabilized for a long time. Despite the general recognition that the amount of nuclear fallout from weapon testings in the 1960s is lower in areas of lower latitude (Hardy, 1968), large amounts of 137Cs deposited into a subtropical mountain forest ecosystem in Taiwan still remain in the surface soils (Chiu et al., 1999a,b). However, the deposition due to the reactor accident at Chernobyl in 1986 was negligible in these areas (Lin and Huang, 1988) and is appropriate for observing characteristics of 137Cs in soils after long-term deposition. Our previous study of organic soils from an undisturbed Chamaecyparis forest in northeastern Taiwan showed that the labile form of 137Cs is mostly transferred to the persistently bound and resistant fractions after long-term deposition of fallout (Chiu et al., 2002). In this new study, we selected two new sites with soils of different pedogenetic origin to compare the bioavailability and behavior of 137Cs in soils between forest and grassland, where radiocaesium has been stabilized for 40 years. 2. Material and methods 2.1. Study sites One of the two study sites is located at a volcano, Yamming Mountain (25 90 N, 121 340 E), in the northern region of Taiwan at approximately 800 m above sea level (Fig. 1). The average annual temperature is approximately 16.7 C and the evenly distributed annual precipitation approximately 4900 mm. The vegetation is basically evergreen broadleaf forest dominated by broad-leaved Lauraceae and Fagaceae trees. Adjacent grassland, dominated by silver grass (Miscanthus floridulus) and dwarf bamboo (Yushania usawai), stretches wide because of infrequent fires. The soils are Andosols formed on andesitic pyroclastic rocks. They exhibit andic properties with very high aluminum saturation. Kaolinite, halloysite, gibbsite, and quartz dominate the clay mineral assemblage, with minor amounts of 2:1 silicate clays. Aluminum toxicity and phosphorus deficiency are believed to contribute to humus accumulation (Chen et al., 1999). The soils of the forest and grassland have distinct litter layers. The slope is 15%. The soil contains about 8e12 cm of a humified Ah horizon underneath the litter layer. The second site is in a subalpine forest reserve at Tatajia (23 280 N, 120 540 E), in the saddle of the Jade Mountain, in central Taiwan, at 2700 m above sea level. The climate is mountainous temperate, with a mean annual precipitation and temperature of approximately 3170 mm and 11.1 C, respectively. The mountainous forests have a tall coniferous canopy of hemlock (Tsuga chinensis), a broad-leaved tree (Trochodendron aralioides) understory, and a ground layer of dwarf bamboo (Yushania niitakayamensis). The dominant species in the grassland are Miscanthus transmorrisonesis and Y. niitakayamensis. A Yushania-dominated zone adjacent to the forestland forms a transition zone or ecotone between the grassland and forest. The origin of the grassland is still unknown, although it might have come from natural succession of the vegetation with elevation or from infrequent fires. Nevertheless, the vegetation has not changed for several decades. The soils of the investigated system are partially podozolized Spodozols (loamy Umbric), with a distinct litter layer, strongly acidic and a steep slope of 25%. Besides the litter layer (Oi horizon), the forest soil contains approximately 13e20 cm of humified organic matter (Oe and Oa horizon) above the mineral horizons. Kaolinite and illite dominate the clay mineral assemblage, with minor amounts of vermiculite. The surface humified mineral layer (Ah horizon) in the grassland soil is deeper than that in forest soil. 2.2. Soil After carefully removing litter, soil samples were taken from the O/A horizons in June 1999. Soil samples were 10 cm deep collected with use of a soil auger 8 cm in diameter. Six samples were combined for each soil type. The samples of fresh soil were passed through a 2-mm sieve and stored in polyethylene bags at 4 C in moist state before use. The soil samples had high organic matter content and cation exchange capacity (CEC) values. The base saturation values were low, contributing to low pH values. The radioactivity level of 137Cs was much higher in volcanic soils than subalpine soils (Table 1). The activity of 137Cs was associated with geographic distribution, because we found in early studies (Chiu et al., 1999a) and elsewhere that the northern mountain areas received more precipitation from mainland monsoons in winter than other seasons and accumulated more nuclear fallout in the soils than other areas.
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Fig. 1. Map of sampling sites. Symbols and indicate the locations of collecting samples for grassland and forest, respectively. The altitude for the two locations is shown using shading and a legend.
Table 1 Chemical properties of soil samples used for sequential extraction of 137Cs Properties pHa CEC (cmol kg1) Base saturation (%) Total organic C (%) Radioactivity of 137Cs (Bq kg1)
Subalpine
Volcanic
Forest
Grassland
Forest
Grassland
3.6 0.1 74.8 1.9 5.1 0.3 39.6 2.6 16.7 1.3
4.1 0.1 34.3 0.4 2.6 0.1 13.7 0.3 37.0 1.7
3.6 0.1 37.7 2.0 1.6 0.3 9.3 0.3 121.5 4.5
3.5 0.1 40.3 2.1 1.1 0.2 8.2 0.2 141.3 6.2
Average of three samples standard deviation of the mean. a Water:soil ¼ 2:1.
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2.3. Sequential extraction of
137
885
Cs
Sequential extraction involved 100 g of soil. The procedure basically followed that of Tessier et al. (1979), but the extraction for the carbonate-bound fraction was omitted because of the low pH of the studied soils. The selective sequential extraction procedure was modified according to Baumann et al. (1996). Fraction I is called ‘‘the exchangeable fraction’’, which was determined by extraction with K2SO4 solution from unfumigated soil. Fraction II is called ‘‘the microbial biomass fraction’’, which was determined as the difference in extraction between chloroform-fumigated and unfumigated treatment. Soils were fumigated for 24 h with ethanol-free CHCl3. After the fumigant was removed, the soils were then extracted with 0.5 M K2SO4. The non-fumigated control soils were extracted under the same conditions at the time fumigation commenced. Microbial biomass Cs (Csmic) was calculated as follows: Csmic ¼ [(Cs extracted from fumigated soil) (Cs extracted from non-fumigated soil)]/KCs. The KCs factor (the extractable part of microbial biomass Cs) was 0.5 as proposed by Bru¨ckmann and Wolters (1994). The soil samples left after the fumigation extraction were used for further steps of sequential extraction. Fraction III is called ‘‘the FeeMn oxide bound fraction’’, which was determined by extraction with 1000 ml of 0.1 M NH2OH$HCl in 0.01 M HNO3 for 5 h at 90 C. Fraction IV is called ‘‘the organic matter-bound fraction’’, which was determined from the residue of fraction III. The residue was resuspended with 100 ml of water and heated at 80 C. An amount of 1000 ml of 30% H2O2 (pH 2.0, adjusted with HNO3) was gradually added while stirring. After cooling, 250 ml of 3.2 M NH4AC in 20% HNO3 was added and agitated for 30 min. The liquid phase contains the desired fraction. Fraction V is called ‘‘the persistently bound fraction’’, which was determined by treating the residue of fraction IV with 1000 ml of 7 M HNO3 and stirring for 6 h at 90 C. The liquid phase contains the desired fraction. Fraction VI is called ‘‘the residual solid fraction’’ resulting from fraction V neutralized with NaOH and dried at 100 C. For each of these extraction steps, the supernatant was removed by centrifugation for 10 min at 10,000 g. The residue of each extraction was washed with 50 ml of distilled water followed by centrifugation and combined with the original supernatant. Centrifuge bottles (250 ml) were used to facilitate centrifuge washing of the samples after each extraction to minimize loss of material. The supernatant of each fraction was filtered and evaporated by heating either to complete dryness or to a minimum volume (<20 ml). The concentration of 137Cs was measured by direct gamma-spectrometry by use of a high-purity Ge detector coupled to a computerized data acquisition system (Canberra Series 95, 4096 channel pulse height analyzer), with Canberra Genie PC software used for spectrum analysis. The efficiency calibration involved use of a standard multi-gamma-ray source mixed in agar. The counting time varied from 80,000 to 320,000 s. The minimum detectable activity of 137Cs was approximately 0.2 Bq. 2.4. Separation and determination of organic matter The water soluble organic carbon (WSOC) fraction was determined by extracting 1 g soil with use of Millipore Milli-Q H2O at a 1:10 ratio of soil to water, followed by shaking at 200 rpm for 16 h at room temperature. The extract was decanted, then centrifuged at 10,000 g for 30 min, and the supernatant was filtered through a Whatman No. 42 filter paper. Soil microbial biomass carbon was measured by fumigation extraction of fresh soil samples. Microbial biomass C (Cmic) was calculated as follows: Cmic ¼ [(organic C extracted from fumigated soil) (organic C extracted from non-fumigated soil)]/KC. The KC factor (the extractable part of microbial biomass C) was 0.45 (Wu et al., 1990). The process for extraction of soil humus basically followed that of Tan (1985). One gram of air-dried soil was extracted first with use of 0.1 M HCl to remove non-humic substances, then the residue was extracted with use of 0.1 M NaOH (with a soil-to-extractant ratio of 1:20) under N2 at room temperature for 16 h. The alkaline supernatant was separated from the residue by centrifugation at 15,000 g for 20 min, then acidified with 6 M HCl to pH 1.0 and allowed to stand at room temperature for 24 h. The supernatant (fulvic acids) was separated from the coagulated material (humic acids) by centrifugation at 8000 g for 15 min. The humic acid fraction was then purified by use of a HCleHF mixture, then the residue was washed thoroughly with distilled water and redissolved with 0.1 M NaOH. Each extracted fraction was analyzed for total organic carbon (O. I. Analytical 1010) by the heatepersulfate oxidation method. Soil organic matter was measured with use of a Fisons NA1500 elemental analyzer.
3. Results and discussion The composition of labile forms of C and extractable humic substances differed greatly among studied soils in vegetation and pedogenic characteristics (Table 2). Microbial biomass C content was higher in forest than in grassland soils. Microbial biomass C, fulvic and humic acids were associated with pedogenic characteristics, and their content
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Table 2 Ratios of each fraction of C to total organic C in studied soils (%) Fraction
Subalpine
Water soluble C Microbial biomass C Fulvic acids Humic acids
Volcanic
Forest
Grassland
Forest
Grassland
0.4 0.0 2.4 0.1 3.8 0.3 7.4 0.5
1.7 0.1 1.3 0.1 7.0 0.4 10.6 0.7
1.4 0.1 0.7 0.0 12.2 0.8 27.5 2.5
0.2 0.0 0.4 0.0 9.9 0.6 22.4 1.4
Average of three samples standard deviation of the mean.
was higher in volcanic than in subalpine soils. Microbial biomass C and WSOC contributed a small portion of the total organic C in all soils. Most of the 137Cs was bound to persistent and residual fractions. For both the forest and grassland soils, 137Cs was mainly present in fraction V, the persistently bound fraction (31e41%), and fraction VI, the residue fraction (22e 62%). Only trace amounts of 137Cs existed in the readily exchangeable fraction (Fig. 2). The proportions of the labile fractions, such as that bound to exchangeable sites, microbial biomass, MneFe oxides, and organic matter (fractions IeIV), were much lower than those of the recalcitrant Vand VI fractions. The labile fractions were higher in forest than in grassland soils. Subalpine forest soil showed a different pattern than the other sites; the persistent and residual fractions contained less 137Cs, which could be attributed to the enormous amount of organic matter (Table 1). Nevertheless, only minor proportions of 137Cs were found in the organic matter fraction of these soils; only 15% was found even in subalpine forest soil, the other sites having even smaller values (1e4%).
80 Forest
a
Grassland
(%)
60
40
20
0
Exch.
Microb.
Oxides
OM
Perst.
80 Forest
Resd.
b
Grassland
(%)
60
40
20
0 137
Exch.
Microb.
Oxides 137
OM
Perst.
Resd.
Fig. 2. Fraction of fallout Cs in studied soils. Recovery of Cs activities from the summation of sequential extraction fractions to bulk soils were within 15% deviations. Vertical bars indicate standard deviation of two replicate measurements. (a) Subalpine soil (b) Volcanic soil.
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In organic soils, the fraction of Cs absorbed in organic sites constitutes a labile, exchangeable pool (Valcke and Cremers, 1994). Radiocaesium is much less efficiently bound and is correspondingly much more mobile in organic soils than in more clay-rich soils. Tegen et al. (1991) suggested that the newly deposited 137Cs is initially bound to organic matter in the early stages of decomposition and is more labile than Cs bound to organic matter produced during the decomposition process. The latter solid species are more resistant in the soil. Compared to results of relatively short-term deposition (Baumann et al., 1996; Rigol et al., 1999), our soils showing long-term deposition had lower amounts of 137Cs bound to exchangeable, microbial biomass or organic matter fractions but higher amounts as persistent and residual fractions, which might be due to the well-known ‘‘aging effect’’ (Absalom et al., 1996; Spezzano, 2005). Recent study of soils affected by the Chernobyl accident indicate that most of the 137Cs is in the residual fraction (Hou et al., 2003), which supports this aging phenomenon. The role of soil microbial biomass in 137Cs retention may be important in highly organic soils (Sanchez et al., 2000). Several studies have reported the upward movement of 137 Cs in forest soil, where fungal activities play key role in translocation of 137Cs (Bru¨ckmann and Wolters, 1994; Rafferty et al., 1997). Being aware of the importance of the availability of 137Cs retained by fungi, some investigators tried to determine the fraction of nuclear fallout 137Cs in fungal hyphae in soil (Bru¨ckmann and Wolters, 1994; Baumann et al., 1996). Recent study indicates that a critical part of fungal mycelia and fruit bodies of mycorrhizal fungi are water soluble (Vinichuk et al., 2005), which contributes to the mobility of 137Cs in the forest soils. Although the fraction of fungal biomass-bound 137Cs is considered a labile form in soil, some studies (Baumann et al., 1996) found that the conventional technique for measuring microbial biomass might underestimate fungal biomass 137Cs, because 137Cs released from microorganisms after chloroform fumigation could be easily fixed by clay minerals, even in organic soil. This observation could explain in part why our volcanic soils contained a relatively higher proportion of 137Cs bound to microbial biomass than the subalpine soils (Fig. 2). Volcanic soils in this area have minor amounts of 2:1 silicate clays (Chen et al., 1999), the mineral associated with permanent 137Cs fixation, thus enabling higher potential for nutrient recycling. In addition, microbial biomass C content was lower in volcanic soils than in subalpine soils (Table 2), which coincided with the proportion of 137Cs and suggested higher microbial activities in volcanic soils. Microbial biomass-bound 137Cs contributes only a minor portion of the total amount of 137Cs, which is consistent with the results of microbial biomass C in these study sites (Table 2) and those of high mountain soils in Taiwan (Chen and Chiu, 2000; Imberger and Chiu, 2001). A great many studies have investigated the diversity of soil microbial biomass in different ecosystems (Wardle, 1998). Although the form of land use is expected to affect the results of conversion of measured data into microbial biomass in different microbial community structures, Joergensen (1996) confirmed that the KC values (0.45) of Vance et al. (1987) and Wu et al., (1990) remained valid, and this value is similar to the KCs value (0.5). Bru¨ckmann and Wolters (1994) proposed to determine microbial biomass Cs. Regardless of the possible underestimation of the microbial biomass fraction of 137Cs (Baumann et al., 1996), the apparently negligible amounts in labile forms supports the general concept that the microbial biomass fraction may act as a pinhole, which supports the recycling of this element within the soil and in the plantesoil system. Organic substances, particularly fulvic acids, play a critical role in the migration of some metal ions in the soil (Saar and Weber, 1982). Our other studies showed that fulvic acids migrate downward in the soil horizons under perhumid forest conditions (Chen and Chiu, 2000), but organic substances do not seem to enhance the leaching of 137Cs to the deeper mineral horizons (Chiu et al., 1999a). This phenomenon is consistent with the observation that interaction of Csþ with organic substances is weak (van Bergeijk et al., 1992). Rigol et al. (1996) also suggest no relation between humic acid content and 134Cs desorption. Shand et al. (1994) showed that oxidizable organic matter, corresponding to humin, strongly retained 137Cs. The real humic substances fraction can barely be defined with the sequential extraction procedure in this study because of the different separation systems. If we consider that humic and fulvic acids, the extractable humic substances, are in moderately labile form, they could be partially associated with the fraction bound to organic matter, and humin is the persistent fraction. The results in Fig. 2 suggest that fulvic and humic acids contribute only in a minor way to 137Cs recycling. This suggestion is also true for the volcanic soils, because the organic matter-bound fraction is low, although volcanic soils contain high humic and fulvic acid content. In this study site, the noncrystalline forms of Fe and Al are associated largely with humus, which is related to the little or no formation of allophane in A horizons (Chen et al., 1999). Such metalehumus complexes are highly stable and resistant to microbial attack and one of the mechanisms used to retain organic matter in volcanic soils (Shoji et al., 1993). We believe that metalehumus complexes could relate to the persistent fraction or residual fraction of this study. Tree biomass might be an important pool
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of mobile 137Cs stored in the forest ecosystem (Strebl et al., 1999). Through litterfall and decomposition of litter, the 137 Cs returns to soil organic matter. Soil with a large amount of organic matter is expected to fix less 137Cs. Most 137Cs passing labile phases is taken up again by microbial biomass and vegetation, particularly in forest soils, which are covered by a thick litter layer. Such phenomenon explains a certain portion of 137Cs remaining in the labile form in the subalpine forest soil. The proportions of 137Cs we found were relatively higher than those in our previous study, conducted in an undisturbed Chamaecyparis forest (Chiu et al., 2002). Parent material and weathering conditions contribute to differing content of 2:1 clay minerals and consequently affect the fixation of 137Cs. For instance, our recent study showed that vermiculite and illite were dominant in the poorly drained soil of the Chamaecyparis forest (Pai et al., 2007), where 137Cs could be transferred easily to a persistently bound and resistant fraction. Rapid up-lifting and erosive conditions resulted in a relatively young geological age for mountainous Taiwan. In addition, the soil with high clay content is barely drained under humid weather conditions, which retards the weathering of phyllosilicates and retains relatively high 2:1 clay minerals (Pai et al., 2004, 2007). A parallel study in the nearby subalpine area indicated that illite and kaolinite are the major clay minerals in the surface horizon of the forest soil (Pai et al., 2004). Decomposition products of coniferous litter increased the forest floor acidity of the surface horizons and promoted the rapid breakdown of illite. Abundant 2:1 clay minerals (illite and vermiculite) exist in grassland soil (unpublished data), which indicates the lower weathering intensity in grassland than in forest soil. The permanent fixation of 137Cs with 2:1 clay minerals in grassland soil explains why only a small proportion of 137Cs exists in labile phases in grassland soil. By contrast, the difference in 137Cs activity between grassland and forest in volcanic soils was relatively small and might be attributed to the characteristics of soil materials per se, since volcanic soils contain a minor amount of 2:1 clays (Chen et al., 1999). 4. Conclusion The grassland and forest soils of our study sites in Taiwan showed features of a possible transformation of 137Cs from labile pools to persistently bound and resistant fractions after long-term deposition of fallout, although forest soils retained a relatively higher proportion in the labile fractions than did grassland soils. Different pedogenetic origins play a crucial role in the differing bioavailability and behavior of 137Cs between forest and grassland soil. The difference in labile 137Cs fractions between forest and grassland soils seems to be associated with soil organic matter content and seems to be higher in soils with abundant 2:1 clay minerals and less so in volcanic soils. Acknowledgements The study was supported by Academia Sinica and National Science Council grants (NSC 92-2623-7-001-002-NU). References Absalom, J.P., Crout, N.M.J., Young, S.D., 1996. Modeling radiocesium fixation in upland organic soils of Northwest England. Environmental Science and Technology 30, 2735e2741. Barber, D.A., 1964. Influence of soil organic matter on the entry of caesium-137 into plants. Nature 204, 1326e1327. Baumann, A., Schimmack, W., Steindl, H., Bunzl, K., 1996. Association of fallout radiocesium with soil constituents: effect of sterilization of forest soils by fumigation with chloroform. Radiation and Environmental Biophysics 35, 229e233. van Bergeijk, K.E., Noordijk, H., Lembrechts, J., Frissel, M.J., 1992. Influence of pH, soil type and soil organic matter content on soil-to-plant transfer of radiocesium and strontium as analyzed by a nonparametric method. Journal of Environmental Radioactivity 15, 265e276. Bru¨ckmann, A., Wolters, V., 1994. Microbial immobilization and recycling of 137Cs in the organic layers of forest ecosystems: relationship to environmental conditions, humification and invertebrate activity. The Science of the Total Environment 157, 249e256. Chen, J.S., Chiu, C.Y., 2000. Effect of topography on the composition of soil organic substances in a perhumid sub-tropical montane forest ecosystem in Taiwan. Geoderma 96, 19e30. Chen, Z.S., Asio, V.B., Yi, D.F., 1999. Characteristics and genesis of volcanic soils along a toposequence under a subtropical climate in Taiwan. Soil Science 164, 510e525. Chiu, C.Y., Lai, S.Y., Lin, Y.M., Chiang, H.C., 1999a. Distribution of the radionuclide 137Cs in the soils of a wet mountainous forest in Taiwan. Applied Radiation and Isotopes 50, 1097e1103. Chiu, C.Y., Lai, S.Y., Wang, C.J., Lin, Y.M., 1999b. Transfer of 137Cs from soil to plants in a wet montane forest in subtropical Taiwan. Journal of Radioanalytical and Nuclear Chemistry 239, 511e515.
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