Journal of Hazardous Materials 280 (2014) 79–88
Contents lists available at ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
On the possible role of macrofungi in the biogeochemical fate of uranium in polluted forest soils b ˇ Jaroslava Kubrová a,b , Anna Zˇ igová c , Zdenˇek Randa , Jan Rohovec c , Milan Gryndler d , b e f Ivana Krausová , Colin E. Dunn , Pavel Kotrba , Jan Boroviˇcka b,c,∗ a
Institute of Geochemistry, Mineralogy and Mineral Resources, Faculty of Science, Charles University, Albertov 6, CZ-12843 Prague 2, Czech Republic ˇ z 130, CZ-25068 Reˇ ˇ z near Prague, Czech Republic Nuclear Physics Institute, v.v.i., Academy of Sciences of the Czech Republic, Husinec–Reˇ c Institute of Geology, v.v.i., Academy of Sciences of the Czech Republic, Rozvojová 269, CZ-16500 Prague 6, Czech Republic d Institute of Microbiology, v.v.i., Academy of Sciences of the Czech Republic, Vídenská 1083, CZ-142 20 Prague 4, Czech Republic ˇ e 8756 Pender Park Drive, Sidney, BC, V8L 3Z5 Canada f Department of Biochemistry and Microbiology, Institute of Chemical Technology, Prague, Technická 3, CZ-166 28 Prague 6, Czech Republic b
h i g h l i g h t s • Uranium is not accumulated in ectomycorrhizal and saprotrophic macrofungi. • Uranium is not accumulated in ectomycorrhizal roots of Picea abies. • Accumulation of metals in macrofungi does not depend on their fractionation in soils.
a r t i c l e
i n f o
Article history: Received 19 February 2014 Received in revised form 9 July 2014 Accepted 24 July 2014 Available online 2 August 2014 Keywords: Uranium Thorium Silver Lead Soil Ectomycorrhizae Fungi Bioaccumulation
a b s t r a c t Interactions of macrofungi with U, Th, Pb and Ag were investigated in the former ore mining district of Pˇríbram, Czech Republic. Samples of saprotrophic (34 samples, 24 species) and ectomycorrhizal (38 samples, 26 species) macrofungi were collected from a U-polluted Norway spruce plantation and tailings and analyzed for metal content. In contrast to Ag, which was highly accumulated in fruit-bodies, concentrations of U generally did not exceed 3 mg/kg which indicates a very low uptake rate and efficient exclusion of U from macrofungi. In ectomycorrhizal tips (mostly determined to species level by DNA sequencing), U contents were practically identical with those of the non-mycorrhizal fine spruce roots. These findings suggest a very limited role of macrofungi in uptake and biotransformation of U in polluted forest soils. Furthermore, accumulation of U, Th, Pb and Ag in macrofungal fruit-bodies apparently does not depend on total content and chemical fractionation of these metals in soils (tested by the BCR sequential extraction in this study). © 2014 Elsevier B.V. All rights reserved.
1. Introduction Uranium (U) is a radioactive toxic heavy metal [1] with typical background concentrations in rocks and soils of only a few mg/kg [2]. Environmental contamination by U may be caused by various processes including mining and processing operations, coal combustion, testing and use of weapons with depleted uranium (DU). Artificially elevated concentrations of U in topsoils usually reach
∗ Corresponding author at: Nuclear Physics Institute, v.v.i., Academy of Sciences ˇ z 130, CZ-25068 Reˇ ˇ z near Prague, Czech Republic. of the Czech Republic, Husinec–Reˇ Tel.: +420 777 008 658. E-mail address: [email protected] (J. Boroviˇcka). http://dx.doi.org/10.1016/j.jhazmat.2014.07.050 0304-3894/© 2014 Elsevier B.V. All rights reserved.
tens or a few hundred mg/kg [3], but natural accumulation in soils can on rare occasions reach 4000 mg/kg [4]. Under the conditions found in the natural environment, the two important oxidation states are U(IV) and U(VI); the chemistry of U(VI) is dominant in most soils [3]. Furthermore, organic compounds influence the mobility and chemistry of U in the environment [5,6]. Geochemical behavior of many heavy metals in the environment is, at least to some extent, affected by microbial [7,8] and fungal activity [9–11] and U is no exception. In vitro experiments have shown that both free living (saprotrophic) and symbiotic (mycorrhizal) fungi exhibit a high U oxide tolerance and possess the ability to solubilize UO3 and U3 O8 and accumulate U within the mycelium to over 80 mg/g (dry weight); some fungi even cause the biomineralization of well-crystallized uranyl phosphate
80
J. Kubrová et al. / Journal of Hazardous Materials 280 (2014) 79–88
minerals [12]. Furthermore, fungi are capable of dissolving DU and transforming the released mobile U complexes into minerals of the meta-autunite group [13]. Apparently, the fungal activities, summarized by Gadd and Fomina [14], may have relevance to contaminated terrestrial habitats. The dependence of almost all land plants on symbiotic fungi, and the fact that symbiotic fungi are capable of U transformations, may make fungal activity of importance in the biogeochemical fate of U in polluted environments and possibly also in phytoremediation or phytostabilization strategies. Arbuscular mycorrhizal (AM) fungi have been demonstrated to enhance U concentrations in roots of host plants and reduce root-to-shoot translocation of U [15]. However, most of the available knowledge on interactions between fungi and U is based on in vitro experiments. Both saprotrophic (SAP) and symbiotic macrofungi are known as effective accumulators or even hyperaccumulators of toxic heavy metals, noble metals, metalloids and radionuclides [16,17]. Despite the astonishing ability of macrofungi to accumulate some metals, concentrations of U in fruit-bodies from pristine environments are very low, with bioaccumulation factors well below 1 [18]; however, there is a lack of data on interactions between macrofungi and U in polluted environments. Similarly to AM fungi, ectomycorrhizal (EM) fungi may contribute to the retention of U in soils via its accumulation in ectomycorrhizae and extraradical mycelia. Furthermore, accumulation of U in mycelia and fruit-bodies of fungal saprotrophs may contribute to retention and redistribution of U in organic soil horizons. We have therefore conducted a field study where fungal fruit-bodies and ectomycorrhizae (EM tips) were used as indicators to assess the possible role of macrofungi in biogeochemical cycling of U in polluted forest soils. To compare the accumulation ability, Th and Pb were analyzed as elements known to be excluded in macrofungi [18,19] and Ag as an effectively accumulated element [20]. Furthermore, the accumulation ability of macrofungi was investigated in relation to the chemical fractionation (mobility, bioavailability) of the investigated metals in soil. 2. Experimental 2.1. Selected sites The Bohemian Massif in the Czech Republic is well-known for various U deposits. For the purpose of this study, we chose the former U mining district of Pˇríbram in Central Bohemia (Fig. 1a) which is a region with perigranitic monometallic veins [21]. In this area, biogeochemical monitoring has revealed a very distinct localized anomaly of U concentrations in mosses and humus, apparently caused by long-term mining and processing operations [22]. We selected a spruce forest plantation (mixed with birch, pine, poplar and lime trees on its margins) on granitic bedrock near the village of Bytíz, near the abandoned shaft 11a (closed in 1991) and tailings. Macrofungi, EM tips, fine spruce roots and soils were sampled throughout the selected part of the forest (Fig. 1b). In addition, some macrofungi and soils were collected directly at the tailings (Fig. 1c). For the purpose of comparison, EM tips and fine spruce roots were also collected in forest plantations at two sites with background U concentrations in soils in the Czech Republic: Kladská village (West Bohemia) and Chmelná village near Nová Cerekev (the Highland) above granitic and gneissic bedrocks, respectively (Fig. 1a). 2.2. Macrofungi, EM tips and fine roots Samples of macrofungi were collected, identified and processed as described elsewhere [18]. EM tips and non-mycorrhizal fine roots of Norway spruce (Picea abies) were sampled from the Oe
horizon, transported to the laboratory, stored in a fridge at 4 ◦ C and processed within 24 h. Briefly, the roots were washed in a stream of tap water. Then, under a stereo-microscope EM tips and fine roots were separated and cleaned thoroughly using tweezers and dissecting needles in distilled water and dried on filter paper. In order to identify the fungal species forming the EM tips we carried out DNA sequencing. DNA was extracted from a miniscule piece of frozen EM tip using the NucleoSpin Plant II DNA isolation kit (Macherey-Nagel, Germany). To amplify fungal DNA, we performed Polymerase Chain Reaction (PCR) with the primer pair ITS1F-ITS4 as described elsewhere [23]. In the case of negative results (low PCR product yield, sequencing failure), the more specific primer pair ITS1F-ITS4B [24] or semi-nested PCR were used [25]. The obtained fragments were purified by isopropanol precipitation and sequenced (Macrogen Inc., Korea). All sequences were identified to species/genus/family/order level by querying the GenBank database, using the nucleotide–nucleotide (blastn) BLAST search option, available through the National Center for Biotechnology Information (http://blast.ncbi.nlm.nih.gov) and the UNITE online database ([26]; UDB accession numbers). Quality sequences were accessioned into the EMBL-Bank database. 2.3. Soils The selection of a representative soil profile pit was on the basis of a thorough preliminary soil survey involving the examination of 16 profiles in the area of circa 9000 m2 in the spruce forest plantation where fungal samples were collected (N49 40.708, E14 04.246). Morphological description and horizon designation followed the scheme of Jahn et al. [27]. Individual soil profiles were classified according to the World Reference Base for Soil Resources [28]. Soil samples from the soil profile (individual soil horizons) were analyzed at the Research Institute for Soil and Water Conservation (Czech Republic, Prague). Analytical procedures for soil characterization followed standard methods. The pH values were measured in distilled water and 1 M KCl with a SenTix21 electrode using the soil:solution ratio of 1:2.5 for all horizons. The following analyses were done for particular horizons: (i) particle size distribution using the pipette method; (ii) cation exchange capacity and base saturation using the method of Mehlich; (iii) organic carbon (Cox ) by wet combustion with a mixture of potassium dichromate and sulphuric acid; (iv) total nitrogen (Nt ) using the Kjeldahl method [29]. In order to assess the extent of metal pollution, 23 samples of Oa horizon were collected throughout the forest plantation and 6 rather heterogeneous organomineral topsoils were also sampled at the tailings in places where macrofungi were collected (Fig. 1b and c). Soils were air-dried and sieved through a 2 mm stainless steel mesh. A representative part of each sample was milled in an agate mill and used for chemical analyses. 2.4. Chemical analyses Concentrations of U, Th, Pb and Ag in macrofungi were determined using the Element2 (Thermo Scientific, Germany) sector field mass spectrometer with inductively coupled plasma (ICP-SF-MS) as described previously [18]. Concentrations of U in EM tips and fine roots from the polluted site were determined by instrumental neutron activation analysis (INAA) using the radioisotope 239 U (74.7 ˇ keV, t1/2 = 23.45 min) according to Randa et al. [30]. In EM tips and fine roots from pristine sites with lower U concentrations, a more sensitive variant with epithermal neutrons and the radioisotope 239 Np (106.13 keV, t 1/2 = 2.36 d) was used. Total concentrations of U, Th and Ag in soils (sample weight ∼250 mg) were determined non-destructively by INAA using the radioisotopes 239 Np for determination of U, 233 Pa (311.9 keV,
J. Kubrová et al. / Journal of Hazardous Materials 280 (2014) 79–88
81
Fig. 1. (A) Sampling sites in the Czech Republic. (B) Sampling area in forest plantation at Bytíz (Pˇríbram district). (C) Sampling sites within the tailings.
t1/2 = 27.4 d) for determination of Th and 110m Ag (657.75 keV, t1/2 = 249.88 d) for determination of Ag. Lead concentrations in soils, which cannot be determined by INAA, were analyzed by instrumental photon activation analysis (IPAA) using the radioisotope 203 Pb ˇ (279.19 keV, t1/2 = 51.873 h) according to Randa et al. [31]. The quality of analytical procedures was repeatedly verified using certified reference materials processed similarly to our biological and soil samples: NIST 1566b (Oyster Tissue), BCR 670 (Duck Weed), NIST 2711 (Montana Soil) and BCR 667 (Estuarine Sediment). In order to determine fractionation of the investigated metals in individual soil horizons, BCR sequential extraction (fraction < 2 mm, unmilled) was carried out according to Rauret et al. [32] in 3 analytical replicates. Concentrations of metals in soil extracts were analyzed by ICP-SF-MS (samples from steps 2 and 3 after dilution). The results are presented as mean ± standard deviation in the (I) exchangeable, (II) reducible, (III) oxidizable, and (IV) residual fractions. The BCR fractions, however, are more correctly operationally defined as the metal extracted by (I) 0.11 M acetic acid, (II) 0.5 M hydroxylammonium chloride, (III) hydrogen peroxide + 1.0 M ammonium acetate, and (IV) sum of fractions I–III subtracted from the total metal content. The quality of the BCR sequential extraction procedure was verified by the use of the certified reference material BCR 483 (Sewage Sludge Amended Soil) in 3 analytical replicates. The Pb values obtained for the steps I–III (0.32 ± 0.02 mg/kg, 360 ± 20.5 mg/kg, and 57.5 ± 5.12 mg/kg, respectively) fell within the indicative concentration ranges: (I): 0.756 ± 0.7 mg/kg, (II): 379 ± 21 mg/kg, (III): 66.5 ± 22 mg/kg [33].
3. Results and discussion 3.1. Soil The investigated forest plantation in Bytíz is covered by Cambic Leptosols. The studied soil profile (Fig. 2) is developed on granite and is relatively shallow (60 cm deep). As documented in Table 1, the soil is extremely acidic, with pH values decreasing up the soil profile. Cation exchange capacity is higher in the upper part of the soil profile and decreases with depth. The values of base saturation
Fig. 2. Soil profile of Cambic Leptosol at Bytíz.
correspond to the character of pedogenesis. The contents of Cox and Nt are the highest in the upper part of the profile and the C/N ratio values showed some enrichment of soil organic matter by Nt in the Oi, Oe, Oa, and Ah horizons. As expected on the granite bedrock, the analysis of the particle size showed the predominance of sand fraction.
82
J. Kubrová et al. / Journal of Hazardous Materials 280 (2014) 79–88
Table 1 Chemical properties, soil organic matter and particle size distribution of the soil profile in forest plantation at Bytíz. Depth (cm)
Horizon
pHH2 O
pHKCl
BS (%)
CEC (mmol+ /100 g)
Cox (%)
Nt (%)
C/N
Clay (%)
Silt (%)
Sand (%)
Texture class
4-3 3-2 2-0 0-5 5-22 22-60
Oi Oe Oa Ah Bw Cr
4.80 4.89 4.44 3.67 3.54 3.91
4.14 4.27 3.71 2.86 2.98 2.89
8 34 62 46 34 15
64.4 88.6 80.7 28.7 8.21 7.58
40.2 32.1 20.3 4.62 0.38 0.12
1.04 1.52 1.43 0.31 0.05 0.05
38.6 21.1 14.2 14.9 7.60 2.40
25.3 21.8 18.2 13.1 10.1 6.50
36.8 61.8 65.4 36.2 37.6 35.5
37.9 16.4 16.4 50.7 52.3 58.0
Loam silt Loam Silt loam Loam Sandy loam Sandy loam
BS: base saturation. CEC: cation exchange capacity. Cox : organic carbon. Nt : total nitrogen. Table 2 Statistical summary of the total metal concentrations (related to dry matter) in soil samples from forest plantation (Oa horizon, 23 samples) and tailings (organomineral topsoil, 6 samples) at Bytíz. See also Supplementary Table s1. Forest
Concentration (mg/kg) Minimum
Median
Mean
Maximum
U Th Pb Ag
6.08 2.52 66.8 0.11
13.7 5.57 615 0.77
19.5 5.90 565 1.30
74.5 9.87 1251 12.3
Tailings
Concentration (mg/kg) Median
Mean
Maximum
20.4 9.54 880 0.29
21.2 10.2 957 0.56
34.8 15.2 2114 1.37
Minimum U Th Pb Ag
11.5 6.21 135 0.21
Statistical uncertainty of INAA/IPAA determinations of U, Th, Pb and Ag was mostly below 1%, 1%, 3% and 7%, respectively.
3.2. Distribution and mobility of metals in soils The investigated forest plantation can be considered U-polluted, with concentrations of U in the range of 6–75 mg/kg in the Oa horizon; relatively high U levels (11–35 mg/kg) were also determined in the substrates from the tailings (Table 2, Supplementary Table S1). The U pollution, however, did not influence the mineral horizons Bw and Cr, where U concentrations correspond with natural levels [2,3]. Furthermore, the area is polluted by Pb, which is the result of long-term smelting operations in the vicinity [34]. Silver is another pollutant derived from these operations [35]. Its concentrations in O horizons are elevated, but lower than those reported from the close vicinity of the smelter [20]; the median of 0.77 mg/kg in the Oa horizon is quite close to values that would be expected in the background [36]. Concentrations of Th are similar to those in unpolluted environments [2]. In the soil profile (Table 3), the highest concentrations of U, Pb and Ag were encountered in the O horizons (Oa > Oe > Oi) and decreased rapidly with depth (Ah > Bw > Cr). This enrichment in O horizons can be explained by sorption of metals to organic matter (O and Ah horizons), resulting in a slow flux of these pollutants to the underlying Bw and Cr horizons. The only element with the Table 3 Total metal concentrations (related to dry matter) in the soil profile (Cambic Leptosol) in forest plantation at Bytíz. Horizon
Oi Oe Oa Ah Bw Cr Granite
Concentration (mg/kg) U
Th
Pb
Ag
0.43 14.2 23.2 7.16 4.09 5.41 6.59
0.18 3.11 5.78 12.2 15.3 21.8 21.4
28.5 382 789 388 38.4 42.5
Contents lists available at ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
On the possible role of macrofungi in the biogeochemical fate of uranium in polluted forest soils b ˇ Jaroslava Kubrová a,b , Anna Zˇ igová c , Zdenˇek Randa , Jan Rohovec c , Milan Gryndler d , b e f Ivana Krausová , Colin E. Dunn , Pavel Kotrba , Jan Boroviˇcka b,c,∗ a
Institute of Geochemistry, Mineralogy and Mineral Resources, Faculty of Science, Charles University, Albertov 6, CZ-12843 Prague 2, Czech Republic ˇ z 130, CZ-25068 Reˇ ˇ z near Prague, Czech Republic Nuclear Physics Institute, v.v.i., Academy of Sciences of the Czech Republic, Husinec–Reˇ c Institute of Geology, v.v.i., Academy of Sciences of the Czech Republic, Rozvojová 269, CZ-16500 Prague 6, Czech Republic d Institute of Microbiology, v.v.i., Academy of Sciences of the Czech Republic, Vídenská 1083, CZ-142 20 Prague 4, Czech Republic ˇ e 8756 Pender Park Drive, Sidney, BC, V8L 3Z5 Canada f Department of Biochemistry and Microbiology, Institute of Chemical Technology, Prague, Technická 3, CZ-166 28 Prague 6, Czech Republic b
h i g h l i g h t s • Uranium is not accumulated in ectomycorrhizal and saprotrophic macrofungi. • Uranium is not accumulated in ectomycorrhizal roots of Picea abies. • Accumulation of metals in macrofungi does not depend on their fractionation in soils.
a r t i c l e
i n f o
Article history: Received 19 February 2014 Received in revised form 9 July 2014 Accepted 24 July 2014 Available online 2 August 2014 Keywords: Uranium Thorium Silver Lead Soil Ectomycorrhizae Fungi Bioaccumulation
a b s t r a c t Interactions of macrofungi with U, Th, Pb and Ag were investigated in the former ore mining district of Pˇríbram, Czech Republic. Samples of saprotrophic (34 samples, 24 species) and ectomycorrhizal (38 samples, 26 species) macrofungi were collected from a U-polluted Norway spruce plantation and tailings and analyzed for metal content. In contrast to Ag, which was highly accumulated in fruit-bodies, concentrations of U generally did not exceed 3 mg/kg which indicates a very low uptake rate and efficient exclusion of U from macrofungi. In ectomycorrhizal tips (mostly determined to species level by DNA sequencing), U contents were practically identical with those of the non-mycorrhizal fine spruce roots. These findings suggest a very limited role of macrofungi in uptake and biotransformation of U in polluted forest soils. Furthermore, accumulation of U, Th, Pb and Ag in macrofungal fruit-bodies apparently does not depend on total content and chemical fractionation of these metals in soils (tested by the BCR sequential extraction in this study). © 2014 Elsevier B.V. All rights reserved.
1. Introduction Uranium (U) is a radioactive toxic heavy metal [1] with typical background concentrations in rocks and soils of only a few mg/kg [2]. Environmental contamination by U may be caused by various processes including mining and processing operations, coal combustion, testing and use of weapons with depleted uranium (DU). Artificially elevated concentrations of U in topsoils usually reach
∗ Corresponding author at: Nuclear Physics Institute, v.v.i., Academy of Sciences ˇ z 130, CZ-25068 Reˇ ˇ z near Prague, Czech Republic. of the Czech Republic, Husinec–Reˇ Tel.: +420 777 008 658. E-mail address: [email protected] (J. Boroviˇcka). http://dx.doi.org/10.1016/j.jhazmat.2014.07.050 0304-3894/© 2014 Elsevier B.V. All rights reserved.
tens or a few hundred mg/kg [3], but natural accumulation in soils can on rare occasions reach 4000 mg/kg [4]. Under the conditions found in the natural environment, the two important oxidation states are U(IV) and U(VI); the chemistry of U(VI) is dominant in most soils [3]. Furthermore, organic compounds influence the mobility and chemistry of U in the environment [5,6]. Geochemical behavior of many heavy metals in the environment is, at least to some extent, affected by microbial [7,8] and fungal activity [9–11] and U is no exception. In vitro experiments have shown that both free living (saprotrophic) and symbiotic (mycorrhizal) fungi exhibit a high U oxide tolerance and possess the ability to solubilize UO3 and U3 O8 and accumulate U within the mycelium to over 80 mg/g (dry weight); some fungi even cause the biomineralization of well-crystallized uranyl phosphate
80
J. Kubrová et al. / Journal of Hazardous Materials 280 (2014) 79–88
minerals [12]. Furthermore, fungi are capable of dissolving DU and transforming the released mobile U complexes into minerals of the meta-autunite group [13]. Apparently, the fungal activities, summarized by Gadd and Fomina [14], may have relevance to contaminated terrestrial habitats. The dependence of almost all land plants on symbiotic fungi, and the fact that symbiotic fungi are capable of U transformations, may make fungal activity of importance in the biogeochemical fate of U in polluted environments and possibly also in phytoremediation or phytostabilization strategies. Arbuscular mycorrhizal (AM) fungi have been demonstrated to enhance U concentrations in roots of host plants and reduce root-to-shoot translocation of U [15]. However, most of the available knowledge on interactions between fungi and U is based on in vitro experiments. Both saprotrophic (SAP) and symbiotic macrofungi are known as effective accumulators or even hyperaccumulators of toxic heavy metals, noble metals, metalloids and radionuclides [16,17]. Despite the astonishing ability of macrofungi to accumulate some metals, concentrations of U in fruit-bodies from pristine environments are very low, with bioaccumulation factors well below 1 [18]; however, there is a lack of data on interactions between macrofungi and U in polluted environments. Similarly to AM fungi, ectomycorrhizal (EM) fungi may contribute to the retention of U in soils via its accumulation in ectomycorrhizae and extraradical mycelia. Furthermore, accumulation of U in mycelia and fruit-bodies of fungal saprotrophs may contribute to retention and redistribution of U in organic soil horizons. We have therefore conducted a field study where fungal fruit-bodies and ectomycorrhizae (EM tips) were used as indicators to assess the possible role of macrofungi in biogeochemical cycling of U in polluted forest soils. To compare the accumulation ability, Th and Pb were analyzed as elements known to be excluded in macrofungi [18,19] and Ag as an effectively accumulated element [20]. Furthermore, the accumulation ability of macrofungi was investigated in relation to the chemical fractionation (mobility, bioavailability) of the investigated metals in soil. 2. Experimental 2.1. Selected sites The Bohemian Massif in the Czech Republic is well-known for various U deposits. For the purpose of this study, we chose the former U mining district of Pˇríbram in Central Bohemia (Fig. 1a) which is a region with perigranitic monometallic veins [21]. In this area, biogeochemical monitoring has revealed a very distinct localized anomaly of U concentrations in mosses and humus, apparently caused by long-term mining and processing operations [22]. We selected a spruce forest plantation (mixed with birch, pine, poplar and lime trees on its margins) on granitic bedrock near the village of Bytíz, near the abandoned shaft 11a (closed in 1991) and tailings. Macrofungi, EM tips, fine spruce roots and soils were sampled throughout the selected part of the forest (Fig. 1b). In addition, some macrofungi and soils were collected directly at the tailings (Fig. 1c). For the purpose of comparison, EM tips and fine spruce roots were also collected in forest plantations at two sites with background U concentrations in soils in the Czech Republic: Kladská village (West Bohemia) and Chmelná village near Nová Cerekev (the Highland) above granitic and gneissic bedrocks, respectively (Fig. 1a). 2.2. Macrofungi, EM tips and fine roots Samples of macrofungi were collected, identified and processed as described elsewhere [18]. EM tips and non-mycorrhizal fine roots of Norway spruce (Picea abies) were sampled from the Oe
horizon, transported to the laboratory, stored in a fridge at 4 ◦ C and processed within 24 h. Briefly, the roots were washed in a stream of tap water. Then, under a stereo-microscope EM tips and fine roots were separated and cleaned thoroughly using tweezers and dissecting needles in distilled water and dried on filter paper. In order to identify the fungal species forming the EM tips we carried out DNA sequencing. DNA was extracted from a miniscule piece of frozen EM tip using the NucleoSpin Plant II DNA isolation kit (Macherey-Nagel, Germany). To amplify fungal DNA, we performed Polymerase Chain Reaction (PCR) with the primer pair ITS1F-ITS4 as described elsewhere [23]. In the case of negative results (low PCR product yield, sequencing failure), the more specific primer pair ITS1F-ITS4B [24] or semi-nested PCR were used [25]. The obtained fragments were purified by isopropanol precipitation and sequenced (Macrogen Inc., Korea). All sequences were identified to species/genus/family/order level by querying the GenBank database, using the nucleotide–nucleotide (blastn) BLAST search option, available through the National Center for Biotechnology Information (http://blast.ncbi.nlm.nih.gov) and the UNITE online database ([26]; UDB accession numbers). Quality sequences were accessioned into the EMBL-Bank database. 2.3. Soils The selection of a representative soil profile pit was on the basis of a thorough preliminary soil survey involving the examination of 16 profiles in the area of circa 9000 m2 in the spruce forest plantation where fungal samples were collected (N49 40.708, E14 04.246). Morphological description and horizon designation followed the scheme of Jahn et al. [27]. Individual soil profiles were classified according to the World Reference Base for Soil Resources [28]. Soil samples from the soil profile (individual soil horizons) were analyzed at the Research Institute for Soil and Water Conservation (Czech Republic, Prague). Analytical procedures for soil characterization followed standard methods. The pH values were measured in distilled water and 1 M KCl with a SenTix21 electrode using the soil:solution ratio of 1:2.5 for all horizons. The following analyses were done for particular horizons: (i) particle size distribution using the pipette method; (ii) cation exchange capacity and base saturation using the method of Mehlich; (iii) organic carbon (Cox ) by wet combustion with a mixture of potassium dichromate and sulphuric acid; (iv) total nitrogen (Nt ) using the Kjeldahl method [29]. In order to assess the extent of metal pollution, 23 samples of Oa horizon were collected throughout the forest plantation and 6 rather heterogeneous organomineral topsoils were also sampled at the tailings in places where macrofungi were collected (Fig. 1b and c). Soils were air-dried and sieved through a 2 mm stainless steel mesh. A representative part of each sample was milled in an agate mill and used for chemical analyses. 2.4. Chemical analyses Concentrations of U, Th, Pb and Ag in macrofungi were determined using the Element2 (Thermo Scientific, Germany) sector field mass spectrometer with inductively coupled plasma (ICP-SF-MS) as described previously [18]. Concentrations of U in EM tips and fine roots from the polluted site were determined by instrumental neutron activation analysis (INAA) using the radioisotope 239 U (74.7 ˇ keV, t1/2 = 23.45 min) according to Randa et al. [30]. In EM tips and fine roots from pristine sites with lower U concentrations, a more sensitive variant with epithermal neutrons and the radioisotope 239 Np (106.13 keV, t 1/2 = 2.36 d) was used. Total concentrations of U, Th and Ag in soils (sample weight ∼250 mg) were determined non-destructively by INAA using the radioisotopes 239 Np for determination of U, 233 Pa (311.9 keV,
J. Kubrová et al. / Journal of Hazardous Materials 280 (2014) 79–88
81
Fig. 1. (A) Sampling sites in the Czech Republic. (B) Sampling area in forest plantation at Bytíz (Pˇríbram district). (C) Sampling sites within the tailings.
t1/2 = 27.4 d) for determination of Th and 110m Ag (657.75 keV, t1/2 = 249.88 d) for determination of Ag. Lead concentrations in soils, which cannot be determined by INAA, were analyzed by instrumental photon activation analysis (IPAA) using the radioisotope 203 Pb ˇ (279.19 keV, t1/2 = 51.873 h) according to Randa et al. [31]. The quality of analytical procedures was repeatedly verified using certified reference materials processed similarly to our biological and soil samples: NIST 1566b (Oyster Tissue), BCR 670 (Duck Weed), NIST 2711 (Montana Soil) and BCR 667 (Estuarine Sediment). In order to determine fractionation of the investigated metals in individual soil horizons, BCR sequential extraction (fraction < 2 mm, unmilled) was carried out according to Rauret et al. [32] in 3 analytical replicates. Concentrations of metals in soil extracts were analyzed by ICP-SF-MS (samples from steps 2 and 3 after dilution). The results are presented as mean ± standard deviation in the (I) exchangeable, (II) reducible, (III) oxidizable, and (IV) residual fractions. The BCR fractions, however, are more correctly operationally defined as the metal extracted by (I) 0.11 M acetic acid, (II) 0.5 M hydroxylammonium chloride, (III) hydrogen peroxide + 1.0 M ammonium acetate, and (IV) sum of fractions I–III subtracted from the total metal content. The quality of the BCR sequential extraction procedure was verified by the use of the certified reference material BCR 483 (Sewage Sludge Amended Soil) in 3 analytical replicates. The Pb values obtained for the steps I–III (0.32 ± 0.02 mg/kg, 360 ± 20.5 mg/kg, and 57.5 ± 5.12 mg/kg, respectively) fell within the indicative concentration ranges: (I): 0.756 ± 0.7 mg/kg, (II): 379 ± 21 mg/kg, (III): 66.5 ± 22 mg/kg [33].
3. Results and discussion 3.1. Soil The investigated forest plantation in Bytíz is covered by Cambic Leptosols. The studied soil profile (Fig. 2) is developed on granite and is relatively shallow (60 cm deep). As documented in Table 1, the soil is extremely acidic, with pH values decreasing up the soil profile. Cation exchange capacity is higher in the upper part of the soil profile and decreases with depth. The values of base saturation
Fig. 2. Soil profile of Cambic Leptosol at Bytíz.
correspond to the character of pedogenesis. The contents of Cox and Nt are the highest in the upper part of the profile and the C/N ratio values showed some enrichment of soil organic matter by Nt in the Oi, Oe, Oa, and Ah horizons. As expected on the granite bedrock, the analysis of the particle size showed the predominance of sand fraction.
82
J. Kubrová et al. / Journal of Hazardous Materials 280 (2014) 79–88
Table 1 Chemical properties, soil organic matter and particle size distribution of the soil profile in forest plantation at Bytíz. Depth (cm)
Horizon
pHH2 O
pHKCl
BS (%)
CEC (mmol+ /100 g)
Cox (%)
Nt (%)
C/N
Clay (%)
Silt (%)
Sand (%)
Texture class
4-3 3-2 2-0 0-5 5-22 22-60
Oi Oe Oa Ah Bw Cr
4.80 4.89 4.44 3.67 3.54 3.91
4.14 4.27 3.71 2.86 2.98 2.89
8 34 62 46 34 15
64.4 88.6 80.7 28.7 8.21 7.58
40.2 32.1 20.3 4.62 0.38 0.12
1.04 1.52 1.43 0.31 0.05 0.05
38.6 21.1 14.2 14.9 7.60 2.40
25.3 21.8 18.2 13.1 10.1 6.50
36.8 61.8 65.4 36.2 37.6 35.5
37.9 16.4 16.4 50.7 52.3 58.0
Loam silt Loam Silt loam Loam Sandy loam Sandy loam
BS: base saturation. CEC: cation exchange capacity. Cox : organic carbon. Nt : total nitrogen. Table 2 Statistical summary of the total metal concentrations (related to dry matter) in soil samples from forest plantation (Oa horizon, 23 samples) and tailings (organomineral topsoil, 6 samples) at Bytíz. See also Supplementary Table s1. Forest
Concentration (mg/kg) Minimum
Median
Mean
Maximum
U Th Pb Ag
6.08 2.52 66.8 0.11
13.7 5.57 615 0.77
19.5 5.90 565 1.30
74.5 9.87 1251 12.3
Tailings
Concentration (mg/kg) Median
Mean
Maximum
20.4 9.54 880 0.29
21.2 10.2 957 0.56
34.8 15.2 2114 1.37
Minimum U Th Pb Ag
11.5 6.21 135 0.21
Statistical uncertainty of INAA/IPAA determinations of U, Th, Pb and Ag was mostly below 1%, 1%, 3% and 7%, respectively.
3.2. Distribution and mobility of metals in soils The investigated forest plantation can be considered U-polluted, with concentrations of U in the range of 6–75 mg/kg in the Oa horizon; relatively high U levels (11–35 mg/kg) were also determined in the substrates from the tailings (Table 2, Supplementary Table S1). The U pollution, however, did not influence the mineral horizons Bw and Cr, where U concentrations correspond with natural levels [2,3]. Furthermore, the area is polluted by Pb, which is the result of long-term smelting operations in the vicinity [34]. Silver is another pollutant derived from these operations [35]. Its concentrations in O horizons are elevated, but lower than those reported from the close vicinity of the smelter [20]; the median of 0.77 mg/kg in the Oa horizon is quite close to values that would be expected in the background [36]. Concentrations of Th are similar to those in unpolluted environments [2]. In the soil profile (Table 3), the highest concentrations of U, Pb and Ag were encountered in the O horizons (Oa > Oe > Oi) and decreased rapidly with depth (Ah > Bw > Cr). This enrichment in O horizons can be explained by sorption of metals to organic matter (O and Ah horizons), resulting in a slow flux of these pollutants to the underlying Bw and Cr horizons. The only element with the Table 3 Total metal concentrations (related to dry matter) in the soil profile (Cambic Leptosol) in forest plantation at Bytíz. Horizon
Oi Oe Oa Ah Bw Cr Granite
Concentration (mg/kg) U
Th
Pb
Ag
0.43 14.2 23.2 7.16 4.09 5.41 6.59
0.18 3.11 5.78 12.2 15.3 21.8 21.4
28.5 382 789 388 38.4 42.5
