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Element levels in birch and spruce wood ashes — green energy?
S CIE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 3 ( 2 00 8 ) 1 9 1–1 97
a v a i l a b l e a t w w w. s c i e n c e d i r e c t . c o m
w w w. e l s e v i e r. c o m / l o c a t e / s c i t o t e n v
Element levels in birch and spruce wood ashes — green energy? Clemens Reimanna,⁎, Rolf Tore Ottesena , Malin Anderssona , Arnold Arnoldussenb , Friedrich Kollerc , Peter Englmaierd a
Geological Survey of Norway (NGU), N-7491 Trondheim, Norway Norwegian Forest and Landscape Institute, Raveien 9, N-1431 Ås, Norway c Centre for Earth Sciences, Dept. of Lithospheric Research, University of Vienna, Althanstr. 14, A-1090 Vienna, Austria d Department of Freshwater Ecology, Faculty of Life Science, University of Vienna, Althanstr. 14, A-1090 Vienna, Austria b
AR TIC LE I N FO
ABS TR ACT
Article history:
Production of wood ash has increased strongly in the last ten years due to the increasing
Received 30 October 2007
popularity of renewable and CO2-neutral heat and energy production via wood burning.
Received in revised form
Wood ashes are rich in many essential plant nutrients. In addition they are alkaline. The
8 January 2008
idea of using the waste ash as fertiliser in forests is appealing. However, wood is also known
Accepted 8 January 2008
for its ability to strongly enrich certain heavy metals from the underlying soils, e.g. Cd,
Available online 11 February 2008
without any anthropogenic input. Concentrations of 26 chemical elements (Ag, As, Au, B, Ba, Ca, Cd, Co, Cr, Cu, Fe, Hg, K, La, Mg, Mn, Mo, Na, Ni, P, Pb, S, Sb, Sr, Ti, and Zn) in 40 samples
Keywords:
each of birch and spruce wood ashes collected along a 120 km long transect in southern
Wood ashes
Norway are reported. The observed maximum concentrations are 1.3 wt.% Pb, 4.4 wt.% Zn
Birch
and 203 mg/kg Cd in birch wood ashes. Wood ashes can thus contain very high heavy metal
Spruce
concentrations. Spreading wood ashes in a forest is a major anthropogenic interference with
Heavy metals
the natural biogeochemical cycles. As with the use of sewage sludge in agriculture the use of
Fertiliser
wood ashes in forests clearly needs regulation.
Action levels
1.
Introduction
Wood burning is an easily available domestic source of heat and energy. Because it is carbon dioxide neutral and renewable it belongs to the group of energy sources for which the term “green energy”: (see, e.g.,: http:// wikipedia.org/wiki/ Green_energy) has been coined. Due to political decisions to reduce greenhouse gas emissions, energy and heat production systems are at present undergoing a rapid transformation in several western European countries. Thermal and electrical power plants based on biofuel (e.g., wood or wood pellets) are becoming more and more popular. For example, in Sweden the production of wood pellets increased in just 4 years from 10,000 t in 1992 to 500,000 t in 1996 (Hillring and Vinterbäck, 1998) and has reached 1,000,000 t in 2004 (http://www.
© 2008 Elsevier B.V. All rights reserved.
pelletcentre.info/CMS/site.asp?p=2316). In Germany pellet production saw a more than ten-fold increase from less than 100,000 t in 2003 to a planned production of more than 1,200,000 t in 2007 (http://www.pelletcentre.info/CMS/site. asp?p=5418). Wood ashes are a waste product of wood burning. Production of wood ashes is already high in many forested areas and it will increase further in the near future due to the massive promotion of biomass fuels. Wood ashes are rich in many essential plant nutrients, in addition they are alkaline and can thus be utilised as a soil fertiliser and liming material. They have been used to improve the nutritional status of boreal forests (e.g., Moilanen et. al., 2002; Hytönen, 2003; Moilanen et al., 2005, Saarsalmi et al., 2005). Large amounts of nutrients are exported from the forests during logging activities. The
⁎ Corresponding author. E-mail address: [email protected] (C. Reimann). 0048-9697/$ – see front matter © 2008 Elsevier B.V. All rights reserved. doi:10.1016/j.scitotenv.2008.01.015
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idea to return the nutrients to the forests and to decrease the acidic condition of the forest soils at the same time using wood ashes as a fertiliser is thus appealing at first glance. However, not all elements that occur in wood ashes are nutrients. In environmental sciences it is often assumed that heavy metals are not taken up in plants and that their levels are thus very low in wood ashes (e.g., Omil et al., 2007). In contrast, in exploration geochemistry it has long been documented (e.g., Brooks, 1972; Kovalevsky, 1979, 1987; Brooks et al., 1995; Dunn, 2007) that plants take up all metals from the soils and can enrich some of these elements, especially the so called “heavy metals” to surprisingly high levels, without any need for anthropogenic contamination (Reimann et al., 2001b, 2007a,b,c). An example of an element that is strongly enriched in organic materials is lead (Pb) (Reimann et al., 2001a, 2007b). Despite the toxic effects of Pb, it occurs naturally in all plants, and in small traces Pb may even be an essential element (Broyer et al., 1972). It is taken up mainly by root hairs and stored as a pyrophosphate in cell walls (Dunn, 2007). Dunn et al. (1992) report a median value of 311 mg/kg Pb in spruce bark ashes (with local concentrations reaching almost 0.7 wt.%) from samples collected on Cape Breton Island, Nova Scotia, Canada, far from any likely anthropogenic contamination source. The highest Pb concentrations occurred near the Java lead deposit (Dunn et al., 1992). High values of environmentally relevant heavy metals in the ashes of biomass fuels have been reported previously (e.g., Obernberger et al., 1997). Other authors (e.g., Ribbing, 2007) discuss some metal values in wood ashes in light of the new EU waste regulations and come to the “astounding result that even pure-wood ashes can be regarded as hazardous waste” when taking these regulations seriously. The latter comment (“astounding”) highlights a widespread misconception in environmental sciences, namely that Mother Nature is always benign. Toxic metal concentrations are toxic metal concentrations, independent of source. Many plants are since long known as accumulators of certain chemical elements (Brooks, 1972; Brooks et al., 1995). Even for potentially toxic metals plant species can be found that accumulate or even hyperaccumulate them (Brooks et al., 1995). The degree of uptake is species dependent, some plants are well protected against certain metals, others are not. For example the very common willow is known to be a hyperaccumulator of Cd and is often considered for phytoremediation purposes (e.g., Mertens et al., 2006). Pine and larch are two wide-spread known accumulators of Pb (Kovalevsky, 1987; Dunn, 2007). There is thus every reason to be concerned about metal concentrations in wood ashes when the ash is returned to the forest. Some authors discuss the combustion process and suggest that it may be possible to fractionate the heavy metals into certain ash fractions (Obernberger et al., 1997; Vervaeke et al., 2006). Others discuss the importance of binding of metals in the ash when utilising ash as fertiliser in agriculture or forestry (Kuokkanen et al., 2006). When using wood ashes as a fertiliser there exists in any case a substantial risk of introducing high levels of toxic trace elements to the forest surface soils at the time of fertilising. The long-term fate of these elements will largely depend on binding, retention and uptake of these elements into the plants. Adding an alkaline fertiliser to (or changing the pH of)
forest soils will severely disturb the kinetics of the natural biogeochemical cycles of many elements (Reimann et al., 2007b). The development of acidic soils in northern forests is a natural process, due to the slow decay of organic material in a wet and cold climate. Spruce needles especially decay very slowly, an important reason why strongly acidic soils can develop in spruce monocultures. When liming these soils, the more alkaline conditions will result in a change in the composition of the organic fraction. The humic substances will react less acidically and their sorption capacity will decline. In contrast to common belief liming will here not result in a better binding but rather in a more rapid release of many toxic trace elements that are bound in the soil organic compartment to the receiving waters (Reimann et al., 2007b). Birch (Betula pubescens Ehrh.) and spruce (Picea abies (L.) Karsten) wood samples were collected at 3 km intervals along a 120 km long transect, cutting the city of Oslo, Norway. To obtain an overview of the potential environmental hazard when using the ashes of the most usual firewood in S. Norway as a fertiliser, element concentrations and variation for 26 chemical elements in the wood ashes are reported here.
1.1.
Location and land use
The study area is located in southern Norway, in the surroundings of, and including, the city of Oslo (for a map of the sample locations see Reimann et al., 2006 or 2007a,b,c). Oslo has about 600,000 inhabitants and is Norway's largest city. Another 1.4 million people, of Norway's total population of 4.5 million, live close to the city. In general the region to the south of Oslo is one of the most intensely utilised and richest agricultural areas of Norway with many interspersed suburban developments. Directly north of Oslo the terrain rises abruptly to an elevation of 400–700 m above sea level (a.s.l.) and land use changes completely. This area, known as “Nordmarka”, is dominated by forests and many lakes and is extensively utilised for nonmotorised recreation purposes by the Oslo population. Forestry is another important activity in the area. Further north, in the Randsfjord area, agriculture is again the dominant land use. In contrast to the Ås area, animal husbandry is widespread and the area is too far inland from Oslo to feel the pressure of spreading urbanisation. The wood samples are thus collected from an area that can be considered as typical for the firewood-production that is burned in the greater Oslo area. The annual precipitation in the investigated area is in the range of 700–1000 mm. The yearly average temperature is 5.7 °C in Oslo. Due to the higher elevation and higher precipitation, the climate in Nordmarka is markedly colder and wetter than along the rest of the transect. For detailed maps on precipitation, temperature and vegetation see Moen (1998). The dominant wind direction is towards the north, i.e. from Oslo towards Nordmarka and Randsfjord.
1.2.
Geology
The transect was selected so that it crossed different lithologies, typical of the greater Oslo area (for a geological map in relation to sample locations see Reimann et al., 2006 or 2007a,b,c). The most remarkable geological feature is the Oslo Rift. A late Carboniferous succession of shale, sandstone and conglomerate (the Asker
S CIE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 3 ( 2 00 8 ) 1 9 1–1 97
Group) marks the onset of deposition related to the Oslo Rift (Dons and Larsen, 1978). The sedimentary sequence was deposited on a levelled surface and is overlain by basalt followed by a thick sequence of latite lava flows (rhomb porphyry) with thin beds of interlayered sand. Further evolution of this Permian rift is characterised by a number of volcanic centres with basaltic to subordinate rhyolitic lavas preserved in collapsed calderas rimmed by major ring faults and dykes. A variety of plutonic rocks cut the units described above (Dons and Larsen, 1978). Precambrian gneisses occur at the southern- and northernmost ends of the transect. During the last ice age the whole study area was icecovered. The thickness of the ice reached its maximum about 18.000–20.000 years ago, and moraine material was deposited in the study area. Due to the melting of the ice the land rose, and the previous coastline in the area can now be found at 150–200 m above sea level (a.s.l.) (Lundmark, 1986). Below this level, large deposits of marine clays can be found. Today most of these areas are used for agriculture and were excluded for sampling. The geological map of Oslo and surrounding areas (Lutro and Nordgulen, 2004) provides an excellent and more detailed geological overview of the study area.
2.
Methods
2.1.
Sampling
Sample sites were selected in accordance with the methods used in the European moss monitoring project (e.g., Rühling, 1994; Zechmeister, 1997). A minimum distance of 300 m to major roads and larger settlements is required, a minimum distance of 100 m to minor roads and houses and a minimum distance of 5 m to forest roads. With two exceptions (site 134, which is only 25 m from a major road (E6) and site 132, which is about 200 m from a major road (E4)) all sample sites, even those in the city of Oslo, fulfil these requirements. Otherwise, at the selected sample density, it was possible to collect the samples at largely “untouched” natural sites in the surrounding forests, which reach into the city, and in large forested parklands. To avoid possible seasonal influences (e.g., Zechmeister et al., 2003) all samples were collected in as short a time span as possible (14 days) in the autumn of 2005. All sites were located in forest ecosystems with a required minimum number of typical (for the area) plant species (presence of Norwegian spruce (P. abies (L.) Karst.), birch (Betula pendula Roth and B. pubescens Ehrh.) and European mountain ash (Sorbus aucuparia L.) in the tree layer; bracken or other species of fern (Pteridium aquilinum (L.) Kuhn), cowberry (Vaccinium vitis-idaea L.), blueberry (bilberry) (Vaccinium myrtillus L.) and terrestrial moss (Hylocomium splendens (Hedw.) Schimp.) in the ground vegetation layer. Even in the city of Oslo it was possible to find sufficiently “natural” forested areas to enable directly comparable sites to be sampled. To collect wood, large twigs (age 10–20 years) were cut from the trees and the bark was completely removed from the twigs using a carbon steel knife. One litre of 2–5 cm long wood cuttings was collected into clear contamination-free polyethylene (PE)-zip-lock plastic bags. Due to the already quite cool fall temperatures there was no problem with condensed water
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collecting in the bags. In the field the PE bags were marked with the site number and a three-digit code for the material. UTM coordinates of all sample sites were recorded using a GARMIN GPS system. Each sample site was recorded in a series of at least three field photos. Every third day the wood samples were posted to the laboratory for immediate drying. After arrival in the laboratory of the Geological Survey of Norway (NGU) all samples were dried to constant weight at T b 40 °C.
2.2.
Analyses
All dried samples were shipped to ACME laboratories in Canada by courier. Here the samples were pulverised to pass a 100-mesh screen in a mild steel mill. The method has been thoroughly checked to avoid any contamination to the plant samples by the mill material. A 0.5 g aliquot of sample material was first leached with concentrated nitric acid for 1 h and then digested in a hot water bath for an additional hour. After cooling, a modified aqua regia solution of equal parts of concentrated ACS grade HCl and HNO3 and de-mineralised H2O was added to each sample (6 mL/g) to leach in a hot (95 °C) water bath for 2 h. After cooling, the solution was made up to a final volume with 5% HCl and then filtered. The sample weight to solution volume ratio is 1 g per 20 mL. The solutions were analysed using a Perkin Elmer Elan 6000 inductively coupled plasma mass spectrometer (ICP-MS) for 39 elements (Ag, Al, As, Au, B, Ba, Bi, Ca, Cd, Co, Cr, Cu, Fe, Ga, Hg, K, La, Mg, Mn, Mo, Na, Ni, P, Pb, Pd, Pt, S, Sb, Sc, Se, Sr, Te, Th, Ti, Tl, U, V, W, and Zn). Loss on ignition (LOI) was determined gravimetrically following the controlled burning of the vegetation samples at 475 °C. For the purpose of this paper the analytical results of “wet ashing” as described above were recalculated to concentrations in the wood ashes rather than using the “dry weight in wood” results. For Pb the authors tested both, wet and dry ashing procedures for 5 samples from the Oslo transect, and the differences between the analytical results were statistically insignificant (Reimann et al., in press).
2.3.
Quality control
The samples were analysed in one large batch of 400 plant samples (terrestrial moss, birch leaves, wood and bark, spruce needles, wood and twigs, European mountain ash leaves, fern leaves). For quality-control purposes the National Institute of Standard and Technology (NIST) standard reference material 1575a Pine Needles and a laboratory internal reference material, Standard V13 Mountain Hemlock Needles, were included evenly spread over the whole batch 8 and 13 times respectively. Detailed results of quality control are reported in Reimann et al. (2006). Average precision for the elements reported here is better than 5%. For 13 elements of the analytical program more than 50% of the values were below the lower limit of detection: Al (100 mg/kg), Bi (0.02 mg/kg), Ga (0.1 mg/kg), Pd (0.5 μg/kg), Pt (0.2 μg/kg), Sc (0.1 mg/kg), Se (0.1 mg/kg). Te (0.02 mg/kg), Th (0.01 mg/kg), Tl (0.02 mg/kg), U (0.01 mg/kg), V (2 mg/kg) and W (0.1 mg/kg). Thus only the concentrations for 26 elements (Ag, As, Au, B, Ba, Ca, Cd, Co, Cr, Cu, Fe, Hg, K, La, Mg, Mn, Mo, Na, Ni, P, Pb, S, Sb, Sr, Ti, and Zn) are presented here.
194 2.4.
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Data analysis
3.
Due to the nature of the data (40 samples, 26 parameters), compositional data with closure effects, and expected strong spatial dependencies formal statistics and statistical tests were avoided. Rather methods of exploratory data analysis (EDA, Tukey, 1977) and robust measures of central tendency and spread are used.
Results and discussion
The analytical results for birch and spruce wood ashes are presented in Table 1. Fig. 1 shows cumulative probability plots of 8 selected elements in birch and spruce wood ashes, clearly demonstrating the species-dependent uptake characteristics. Table 2 shows the ratio of the median values of birch to the
Table 1 – Element concentrations in 40 samples (birch: 38 samples) of birch and spruce wood ashes, collected along a 120 km long transect in S-Norway Element Ag mg/kg As mg/kg Au mg/kg B mg/kg Ba mg/kg Ca wt.% Cd mg/kg Co mg/kg Cr mg/kg Cu mg/kg Fe wt.% Hg mg/kg K wt.% La mg/kg Mg wt.% Mn wt.% Mo mg/kg Na wt.% Ni mg/kg P wt.% Pb mg/kg S wt.% Sb mg/kg Sr mg/kg Ti mg/kg Zn mg/kg
Material Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce
DL
Min
5%
25%
Median
Mean
75%
95%
Max
MAD
CVR_%
0.002
1 4 2 2 0.03 0.02 276 268 2060 1450 8.3 24 26 4 3 1 147 183 138 292 0.15 0.09 0.2 0.1 6.4 5.6 1.1 0.9 1.9 2.7 0.56 0.31 1.2 1.0 0.28 0.26 18 14 1.8 0.56 37 8 1.2 1.1 0.9 1.8 488 711 92 91 3910 2630
1.8 5 2.9 2 0.07 0.02 459 278 2280 2810 18 26 30 8 6 1.95 236 199 343 314 0.21 0.1 0.3 0.1 13 6 1.2 1.0 5.5 2.9 0.92 1.5 1.3 1.1 0.46 0.38 24.8 21.9 4.3 0.58 53.7 14.8 1.5 1.3 2.29 1.9 912 1020 230 96.9 5970 3530
2.6 6 6 6 0.11 0.05 606 398 3610 5120 20 28 45 21 20 5 288 246 398 504 0.26 0.13 0.4 0.4 18 9 1.6 1.2 6.6 3.4 1.9 2.3 2.6 1.3 0.52 0.46 31 26 4.9 1 170 37 2.9 2.1 2.6 2.3 1160 1320 268 118 9380 4500
3.3 9 10 12 0.14 0.10 652 484 4560 6620 22 31 57 31 34 8 306 268 473 594 0.31 0.14 0.6 0.6 19 10 2.6 1.3 7.7 4.2 3.7 4 3.3 2.1 0.61 0.53 37 28 6.2 1.5 516 67.5 5.9 2.6 3.1 2.6 1420 1590 306 132 14,600 5060
4 9 15 12 0.23 0.13 689 462 5600 7430 22.3 32 65 37 38 9.93 316 266 465 579 0.39 0.18 0.7 0.6 19.3 10 7.1 1.7 7.7 4.3 3.8 4.4 5.9 2.3 0.61 0.52 49 34 6.4 1.4 965 144 5.7 3.9 2.99 2.51 1490 1710 318 135 14,300 5350
5 13 19 16 0.25 0.14 770 536 6420 9640 24.5 35 76 50 51 12 352 290 515 674 0.32 0.23 0.9 0.8 21 11 4.9 1.4 8.8 5.1 5.4 5.7 6.4 2.6 0.65 0.57 62 40.5 7.5 1.7 834 162 7.5 5.6 3.3 2.8 1670 1940 330 142 16,700 6200
7 15 46 21 0.50 0.19 985 567 11300 13500 28 38 115 73 81 21 450 330 642 782 0.35 0.29 1.3 1.0 24 13 27 2.6 9.9 5.7 7.6 9.1 15 5.5 0.75 0.73 89 60 10 2.2 2320 423 10 8.7 3.5 2.9 2330 3170 494 209 22,600 7590
8 18 52 44 1.5 1.1 1050 588 20700 17800 28.4 43 203 184 98 37 508 342 650 860 0.37 0.3 1.7 1.2 25 13.4 86 15 13 6.8 9.3 12 41 8 1.1 0.82 156 80 10.5 2.4 13700 1290 14 9 7 3 3400 3840 652 268 43,500 7910
1.33 4.3 7.41 7.41 0.09 0.07 110 83.8 1680 3310 3.54 4.02 23.7 14.8 23.7 4.45 51.9 33.4 87.5 123 0.03 0.04 0.45 0.3 2.99 1.75 1.63 0.15 1.55 1.32 2.86 2.69 1.93 1.04 0.08 0.08 16.3 5.19 1.93 0.57 489 63 4.09 1.4 0.3 0.3 430 445 48.9 16.3 5920 1280
40 49 74 62 64 70 17 17 37 50 16 13 42 48 70 56 17 12 18 21 11 27 74 50 16 18 63 11 20 31 77 67 58 50 13 15 44 19 31 38 95 93 69 54 10 11 30 28 16 12 41 25
0.02 0.0002 1 0.1 0.01 0.01 0.01 0.1 0.01 0.001 0.001 0.01 0.01 0.001 0.00001 0.01 0.01 0.1 0.01 0.01 0.01 0.02 0.5 1 0.1
DL: detection limit, Min: minimum, Max: maximum, MAD: median absolute deviation and CVR: robust coefficient of variation.
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195
Fig. 1 – Cumulative probability plots of 8 selected elements in birch and spruce wood ashes, demonstrating the species-dependent uptake characteristics. The maximum admissible concentrations for sewage sludge used as fertiliser in agriculture are shown in some diagrams as a stippled line.
median values for spruce wood ashes, sorted according to relative enrichment (depletion) in birch. Not unexpectedly there are large species-dependent differences in the median concentrations between birch and spruce wood ashes although the samples were collected at the same sites (see Fig. 1). The major nutrients Ca, K, Mg, P, S and Mn all occur in wt.% concentrations in the wood ashes, median concentrations of Ca reach over 30 wt.% in spruce ashes (Table 1). Barium, Fe, Na, Sr and Zn all reach median concentrations well above 1000 mg/kg in the wood ashes (Table 1). Also the “heavy metals” Cd, Cr, Cu and Pb all reach disturbingly high median concentrations in the wood ashes. When comparing birch and spruce, Pb, followed by P, Co and Zn are most enriched in birch wood ashes, while Ag is
relatively enriched in spruce wood ashes (Fig. 1, Table 2). Lead shows the largest variation of all elements in both, birch and spruce wood ashes (Fig. 1). Lanthanum shows the largest difference in variation between the two species, a high variation in birch wood ashes and almost no variation in spruce wood ashes (Table 2). Maximum concentrations reach surprisingly high levels for some of the toxic trace elements. Lead reaches more than 1.3 wt.% in birch wood ashes (median 516 mg/kg) at site 127 and even 1290 mg/kg in spruce wood ashes from the same site. At this site a small Pb-mineralisation was subsequently detected in the local bedrock during follow-up work (Reimann et al., in press). Such small Pbmineralisations are quite typical for the larger Oslo area (e.g., Nilsen and Bjørlykke, 1991). Neither birch nor spruce is an
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Table 2 – Median concentrations in birch and spruce wood ashes (all in mg/kg) and the ratio birch wood ashes/spruce wood ashes
Ag Ba Ca Cu As Sr Mn Hg Cr Na Sb Ni B Au Mo Cd Mg K La Fe S Ti Zn Co P Pb
Birch
Spruce
Ratio
Median
Median
Birch/spruce
3.3 4560 220,000 473 10 1420 36,500 0.6 306 6120 3.1 37 652 0.14 3.3 57 77,600 188,000 2.6 3060 58,800 306 14,600 34 62,000 516
8.7 6620 311,000 594 12 1590 40,500 0.6 268 5260 2.6 28 484 0.1 2.1 31 42,200 101,000 1.3 1400 26,100 132 5060 8 14,500 67.5
0.4 0.7 0.7 0.8 0.8 0.9 0.9 1.0 1.1 1.2 1.2 1.3 1.3 1.5 1.6 1.8 1.8 1.9 2.0 2.2 2.3 2.3 2.9 4.3 4.3 7.6
The table is sorted according to an increasing median ratio.
accumulator of Pb; spruce clearly avoids uptake of Pb (Reimann et al., 2007a). Larch and pine are two widespread, documented accumulators of Pb, and Pb-concentrations 300 times higher than in other species collected at the same localities are reported in literature (e.g., Kovalevsky, 1987). The maximum concentrations of Cd are also amazingly high in both, birch and spruce wood ashes: 203 mg/kg and 184 mg/kg respectively, without any apparent source. Zinc reaches a maximum concentration of 4.3 wt.% in birch wood ashes and 7910 mg/kg in spruce wood ashes. A general statement that wood ashes have a “relatively low metal content” (e.g., Omil et al., 2007) is thus not justified. For agricultural soils and for sewage sludge that is used as fertiliser in agriculture clearly defined maximum allowable concentrations have been defined for a number of metals (Cd, Cu, Ni, Pb, Zn, Hg and Cr) (Council Directive 86/278/EEC of 12 June 1986). No such levels have been defined for forest soils or for materials (e.g., wood ashes) that are spread in the forest. It is, however, generally forbidden to spread sewage sludge in forests. The concentrations of many elements reported in wood ashes here even exceed the upper limits as defined for sewage sludge quality class III in the Norwegian regulations (FOR 2003-07-04 nr. 951) for use in agriculture. For Cd only 1 sample of the wood ashes is below the upper limit of 5 mg/kg. For Pb about 70% of the birch wood ashes and still 25% of the spruce wood ashes exceed the maximum allowable concentration of 200 mg/kg. For Zn all samples for both species exceed the maximum allowable concentration of 1500 mg/kg.
Some of the wood ashes fall into the “toxic waste” class in recently defined concentration classes for soils in Norwegian kindergardens, playgrounds and school yards in industrial areas (Pb and Zn — ca. 5% of all wood ash samples exceed the levels defined for class 5 (Alexander, 2007)). It appears thus necessary to consider whether these ashes should be treated as dangerous waste rather than as fertilisers (see also discussion in Ribbing, 2007). The picture becomes even more threatening when considering how ashes spread in a forest ecosystem will disturb the natural biogeochemical cycles (see, for example, Reimann et al., 2007b). The ashes have alkaline properties, one reason why they are considered positive for remediating the acidic soil environment in forests (e.g., Jala and Goyal, 2006; Omil et al., 2007). Increasing the pH in a forest ecosystem will boost the decay kinetics of organic material, this can lead to a sudden release of metals. When adding ashes with high metal concentrations many of the metals will strongly bind to the organic matrix. Over time this will lead to a disproportionate enrichment of heavy metals in forest soils and a decrease of binding sites for nutrients. While the nutrients and all other elements, including the “heavy metals” are usually released very slowly (over tens of years) during the natural decay of organic material, they are all at once available when spreading ashes in the forest. As a result there will be a sudden excess of many elements. Local conditions and the different binding strengths of the elements to the soil organic matrix will determine the fate of this excess. None of the possible long-term effects is beneficial. It is of course justified to argue that any forest fire will have the same effect, however, forest fires are a relatively rare event compared to adding thousands of tons of wood ashes as fertilisers to forests. The forests protect groundwater resources. They are only able to do this, when such interferences are kept to a minimum. Usually the organic material in a forest deteriorates quite slowly — over tens of years. Differences in natural conditions (e.g., climate, drainage conditions) can lead to very different metal concentrations in the organic layer of forest soils without any human interference (Reimann et al., 2007b), even if the metal concentrations are more or less the same in the underlying minerogenic soils. It is a major anthropogenic interference with the natural equilibrium of decay and release of organic matter when burning the wood and spreading the resulting ashes in the forest. As a minimum such wood ashes clearly need an analytical certificate. Well considered maximum allowable concentrations for many elements need to be established before such ashes can be spread in forests or any other near-natural ecosystem. The import of metals into forest ecosystems via such “fertilisation” will in any case be much higher than any anthropogenic impact of heavy metals from the atmosphere via long range transport.
4.
Conclusions
Wood ashes are rich in nutrients but can also contain stunningly high natural concentrations of potentially toxic metals. The elements reaching peak concentrations will vary from plant species to plant species. Additionally, wood ashes are alkaline. When spreading wood ashes in a forest, even in the forest from which the wood originated, a one-time addition of these metals
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in combination with changing the pH of the soil is a strong interference with the natural biogeochemical cycles. The use of fly ash as “fertiliser” in forestry clearly needs legislation. It is the metal concentrations and their availability that need to be considered, they are by no means harmless due to their “natural” origin. “Green energy” may not always be as green as it looks at first glance.
REFERENCES Alexander J. Anbefalte kvalitetskriterier for jord i barnehager, lekeplasser og skoler på industristeder basert på helsevurderingar (Guidelines for soils in day-care centers, playgrounds and school yards in industrial areas based on health risk assessment). Nasjonalt folkehelseinstitutt (Norwegian Institute of Public Health),; 2007. September, (in Norwegian). Brooks RR. Geobotany and biogeochemistry in mineral exploration. New York: Harper and Row; 1972. 290 pp. Brooks RR, Dunn CE, Hall GEM, editors. Biological systems in mineral exploration and processing. New York: Ellis Horwood; 1995. 538 pp. Broyer TC, Johnson CN, Paull RE. Some aspects of lead in plant nutrition. Plant Soil 1972;36:301. Council Directive 86/278/EEC of 12 June 1986 on the protection of the environment, and in particular of the soil, when sewage sludge is used in agriculture. Dons JA, Larsen BT, editors. The Oslo Paleorift, a review and guide to excursions, vol. 337. Nor geol unders Bull; 1978. p. 1–199. Dunn CE. Biogeochemistry in mineral exploration. Amsterdam: Elsevier; 2007. 480 pp. Dunn CE, Adcock SW, Spirito WA. Reconnaissance Biogeochemical Survey Southeastern Cape Breton Island Nova Scotia. Part 1 — black spruce bark. Geol Surv Can Open File 1992;2558 39 pp. FOR 2003-07-04 nr. 951; Forskrift om gjødselsvarer mv. av organisk opphav. http://www.lovdata.no/cgi-wift/ldles?doc=/ sf/sf/sf-20030704-0951.html, 2003. Hillring B, Vinterbäck J. Wood pellets in the Swedish residential market. For Prod J 1998;48:67–72. Hytönen J. Effects of wood, peat and coal ash fertilization on Scots pine foliar nutrient concentrations and growth on afforested former agricultural peat soils. Silva Fenn 2003;37:219–34. Jala S, Goyal D. Fly ash as a soil ameliorant for improving crop production — a review. Biosource Technol 2006;97:1136–47. Kovalevsky A.L. Biogeochemical exploration for mineral deposits. English Translation of 1974 Russian text, Nedra Publishers, Moscow, Russia. Amerind Publishing Co. PVT. Ltd., New Delhi, India. 1979. Kovalevsky AL. Biogeochemical exploration for mineral deposits. 2nd edition. Utrecht, The Netherlands: VNU Science Press; 1987. 220 pp. Kuokkanen T, Pöykiö R, Nurmesniemi H, Rämö J. Sequential leaching of heavy metals and sulfur in bottom ash and fly ash from the combustion of wood and peat at a municipal district heating plant. Chem Speciat Bioavailab 2006;18:131–42. Lundmark JE. Skogmarkens ekologi. Ståndortsanpasset skogbruk. Del 1 — Grunder. Jönköping: Skogsstyrelsen; 1986. 158 pp. Lutro O, Nordgulen Ø. Oslofeltet, berggrunnskart M 1: 250.000. Trondheim: Norges geologiske undersøkelse; 2004. Mertens J, Vervaeke P, Meers E, Tack FMG. Seasonal changes of metals in willow (Salix sp.) stands for phytoremediation on dredged sediments. Environ Sci Technol 2006;40(6):1962–8. Moen A. Nasjonalatlas for Norge: Vegetasjon. Hønefoss: Statens Kartverk; 1998. Moilanen M, Silfverberg K, Hokkanen TJ. Effects of wood-ash on the tree growth, vegetation and substrate quality of a drained mire: a case study. For Ecol Manag 2002;171:321–38.
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Moilanen M, Silfverberg K, Hvkkd H, Issakainen J. Wood ash as a fertilizer on drained mires — growth and foliar nutrients of Scots pine. Can J For Res 2005;35:2734–42. Nilsen O, Bjørlykke A. Silver mineralization in the ancient Akersberg mine, Oslo. Nor Geol Tidskr 1991;71:121–8. Obernberger I, Biedermann F, Widmann W, Riedl R. Concentrations of inorganic elements in biomass fuels and recovery in the different ash fractions. Biomass and Bioenergy 1997;12:211–24. Omil B, Pineiro V, Merino A. Trace elements in soils and plants in temperate forest plantations subjected to single and multiple application of mixed wood ash. Sci Total Environ 2007;381:157–68. Reimann C, Kashulina G, Caritat P de, Niskavaara H. Multi-element, multi-medium regional geochemistry in the European arctic: element concentration, variation and correlation. Appl Geochem 2001a;16:759–80. Reimann C, Koller F, Frengstad B, Kashulina G, Niskavaara H, Englmaier P. Comparison of the element composition in several plant species and their substrate from a 1,500,000 km2-area in Northern Europe. Sci Total Environ 2001b;278:87–112. Reimann C, Arnoldussen A, Boyd R, Finne TE, Nordgulen Ø, Volden T, Englmaier P. The influence of a city on element contents of a terrestrial moss (Hylocomium splendens). Sci Total Environ 2006;369:419–32. Reimann C, Arnoldussen A, Boyd R, Finne TE, Koller F, Nordgulen Ø, Englmaier P. Element contents in leaves of four plant species (birch, mountain ash, fern and spruce) along anthropogenic and geogenic concentration gradients. Sci Total Environ 2007a;377:416–33. Reimann C, Arnoldussen A, Englmaier P, Filzmoser P, Finne TE, Garrett RG, et al. Element concentrations and variations along a 120 km long transect in south Norway — anthropogenic vs. geogenic vs. biogenic element sources and cycles. Appl Geochem 2007b;22:851–71. Reimann C, Arnoldussen A, Finne TE, Koller F, Nordgulen Ø, Englmaier P. Element contents in birch leaves, bark and wood under different anthropogenic and geogenic conditions. Appl Geochem 2007c;22:1549–66. Reimann C., Flem B., Arnoldussen A., Englmaier P., Finne T.E., Koller F., et al., Nordgulen Ø. The biosphere: a homogeniser of Pb-isotope signals. Appl Geochem in press. Ribbing C. Environmentally friendly use of non-coal ashes in Sweden. Waste Management 2007;27:1428–35. Rühling Å, editor. Atmospheric heavy metal deposition in Europe— estimations based on moss analysis, vol. 9. Nord; 1994. 52 pp. Saarsalmi A, Derome J, Levula T. Effect of wood ash fertilization on stand growth, soil, water and needle chemistry, and berry yields of lingonberry (Vaccinium vitis-idaea L,) in a Scots pine stand in Finland. In: Mandre M, editor. Utilisation of industrial wastes in forestry, vol. 42. Forestry studies; 2005. p. 13–33. Tukey JW. Exploratory data analysis. Reading: Addison-Wesley; 1977. Vervaeke P, Tack FMG, Navez F, Martin J, Verloo MG, Lust N. Fate of heavy metals during fixed bed downcraft gasification of willow wood harvested from contaminated sites. Biomass and Bioenergy 2006;30:58–65. Zechmeister H. Schwermetalldeposition in Österreich- erfasst durch biomonitoring mit Moosen (Aufsammlung 1995). Umweltbundesamt Wien, vol. 94. . Monographien Band; 1997. (in German with English summary). Zechmeister HG, Hohenwallner D, Riss A, Hanus-Illnar A. Variations in heavy metal concentrations in the moss species Abietinella abietina (Hedw.) Fleisch. according to sampling time, within site variability and increase in biomass. Sci Total Environ 2003;301:55–65.
a v a i l a b l e a t w w w. s c i e n c e d i r e c t . c o m
w w w. e l s e v i e r. c o m / l o c a t e / s c i t o t e n v
Element levels in birch and spruce wood ashes — green energy? Clemens Reimanna,⁎, Rolf Tore Ottesena , Malin Anderssona , Arnold Arnoldussenb , Friedrich Kollerc , Peter Englmaierd a
Geological Survey of Norway (NGU), N-7491 Trondheim, Norway Norwegian Forest and Landscape Institute, Raveien 9, N-1431 Ås, Norway c Centre for Earth Sciences, Dept. of Lithospheric Research, University of Vienna, Althanstr. 14, A-1090 Vienna, Austria d Department of Freshwater Ecology, Faculty of Life Science, University of Vienna, Althanstr. 14, A-1090 Vienna, Austria b
AR TIC LE I N FO
ABS TR ACT
Article history:
Production of wood ash has increased strongly in the last ten years due to the increasing
Received 30 October 2007
popularity of renewable and CO2-neutral heat and energy production via wood burning.
Received in revised form
Wood ashes are rich in many essential plant nutrients. In addition they are alkaline. The
8 January 2008
idea of using the waste ash as fertiliser in forests is appealing. However, wood is also known
Accepted 8 January 2008
for its ability to strongly enrich certain heavy metals from the underlying soils, e.g. Cd,
Available online 11 February 2008
without any anthropogenic input. Concentrations of 26 chemical elements (Ag, As, Au, B, Ba, Ca, Cd, Co, Cr, Cu, Fe, Hg, K, La, Mg, Mn, Mo, Na, Ni, P, Pb, S, Sb, Sr, Ti, and Zn) in 40 samples
Keywords:
each of birch and spruce wood ashes collected along a 120 km long transect in southern
Wood ashes
Norway are reported. The observed maximum concentrations are 1.3 wt.% Pb, 4.4 wt.% Zn
Birch
and 203 mg/kg Cd in birch wood ashes. Wood ashes can thus contain very high heavy metal
Spruce
concentrations. Spreading wood ashes in a forest is a major anthropogenic interference with
Heavy metals
the natural biogeochemical cycles. As with the use of sewage sludge in agriculture the use of
Fertiliser
wood ashes in forests clearly needs regulation.
Action levels
1.
Introduction
Wood burning is an easily available domestic source of heat and energy. Because it is carbon dioxide neutral and renewable it belongs to the group of energy sources for which the term “green energy”: (see, e.g.,: http:// wikipedia.org/wiki/ Green_energy) has been coined. Due to political decisions to reduce greenhouse gas emissions, energy and heat production systems are at present undergoing a rapid transformation in several western European countries. Thermal and electrical power plants based on biofuel (e.g., wood or wood pellets) are becoming more and more popular. For example, in Sweden the production of wood pellets increased in just 4 years from 10,000 t in 1992 to 500,000 t in 1996 (Hillring and Vinterbäck, 1998) and has reached 1,000,000 t in 2004 (http://www.
© 2008 Elsevier B.V. All rights reserved.
pelletcentre.info/CMS/site.asp?p=2316). In Germany pellet production saw a more than ten-fold increase from less than 100,000 t in 2003 to a planned production of more than 1,200,000 t in 2007 (http://www.pelletcentre.info/CMS/site. asp?p=5418). Wood ashes are a waste product of wood burning. Production of wood ashes is already high in many forested areas and it will increase further in the near future due to the massive promotion of biomass fuels. Wood ashes are rich in many essential plant nutrients, in addition they are alkaline and can thus be utilised as a soil fertiliser and liming material. They have been used to improve the nutritional status of boreal forests (e.g., Moilanen et. al., 2002; Hytönen, 2003; Moilanen et al., 2005, Saarsalmi et al., 2005). Large amounts of nutrients are exported from the forests during logging activities. The
⁎ Corresponding author. E-mail address: [email protected] (C. Reimann). 0048-9697/$ – see front matter © 2008 Elsevier B.V. All rights reserved. doi:10.1016/j.scitotenv.2008.01.015
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idea to return the nutrients to the forests and to decrease the acidic condition of the forest soils at the same time using wood ashes as a fertiliser is thus appealing at first glance. However, not all elements that occur in wood ashes are nutrients. In environmental sciences it is often assumed that heavy metals are not taken up in plants and that their levels are thus very low in wood ashes (e.g., Omil et al., 2007). In contrast, in exploration geochemistry it has long been documented (e.g., Brooks, 1972; Kovalevsky, 1979, 1987; Brooks et al., 1995; Dunn, 2007) that plants take up all metals from the soils and can enrich some of these elements, especially the so called “heavy metals” to surprisingly high levels, without any need for anthropogenic contamination (Reimann et al., 2001b, 2007a,b,c). An example of an element that is strongly enriched in organic materials is lead (Pb) (Reimann et al., 2001a, 2007b). Despite the toxic effects of Pb, it occurs naturally in all plants, and in small traces Pb may even be an essential element (Broyer et al., 1972). It is taken up mainly by root hairs and stored as a pyrophosphate in cell walls (Dunn, 2007). Dunn et al. (1992) report a median value of 311 mg/kg Pb in spruce bark ashes (with local concentrations reaching almost 0.7 wt.%) from samples collected on Cape Breton Island, Nova Scotia, Canada, far from any likely anthropogenic contamination source. The highest Pb concentrations occurred near the Java lead deposit (Dunn et al., 1992). High values of environmentally relevant heavy metals in the ashes of biomass fuels have been reported previously (e.g., Obernberger et al., 1997). Other authors (e.g., Ribbing, 2007) discuss some metal values in wood ashes in light of the new EU waste regulations and come to the “astounding result that even pure-wood ashes can be regarded as hazardous waste” when taking these regulations seriously. The latter comment (“astounding”) highlights a widespread misconception in environmental sciences, namely that Mother Nature is always benign. Toxic metal concentrations are toxic metal concentrations, independent of source. Many plants are since long known as accumulators of certain chemical elements (Brooks, 1972; Brooks et al., 1995). Even for potentially toxic metals plant species can be found that accumulate or even hyperaccumulate them (Brooks et al., 1995). The degree of uptake is species dependent, some plants are well protected against certain metals, others are not. For example the very common willow is known to be a hyperaccumulator of Cd and is often considered for phytoremediation purposes (e.g., Mertens et al., 2006). Pine and larch are two wide-spread known accumulators of Pb (Kovalevsky, 1987; Dunn, 2007). There is thus every reason to be concerned about metal concentrations in wood ashes when the ash is returned to the forest. Some authors discuss the combustion process and suggest that it may be possible to fractionate the heavy metals into certain ash fractions (Obernberger et al., 1997; Vervaeke et al., 2006). Others discuss the importance of binding of metals in the ash when utilising ash as fertiliser in agriculture or forestry (Kuokkanen et al., 2006). When using wood ashes as a fertiliser there exists in any case a substantial risk of introducing high levels of toxic trace elements to the forest surface soils at the time of fertilising. The long-term fate of these elements will largely depend on binding, retention and uptake of these elements into the plants. Adding an alkaline fertiliser to (or changing the pH of)
forest soils will severely disturb the kinetics of the natural biogeochemical cycles of many elements (Reimann et al., 2007b). The development of acidic soils in northern forests is a natural process, due to the slow decay of organic material in a wet and cold climate. Spruce needles especially decay very slowly, an important reason why strongly acidic soils can develop in spruce monocultures. When liming these soils, the more alkaline conditions will result in a change in the composition of the organic fraction. The humic substances will react less acidically and their sorption capacity will decline. In contrast to common belief liming will here not result in a better binding but rather in a more rapid release of many toxic trace elements that are bound in the soil organic compartment to the receiving waters (Reimann et al., 2007b). Birch (Betula pubescens Ehrh.) and spruce (Picea abies (L.) Karsten) wood samples were collected at 3 km intervals along a 120 km long transect, cutting the city of Oslo, Norway. To obtain an overview of the potential environmental hazard when using the ashes of the most usual firewood in S. Norway as a fertiliser, element concentrations and variation for 26 chemical elements in the wood ashes are reported here.
1.1.
Location and land use
The study area is located in southern Norway, in the surroundings of, and including, the city of Oslo (for a map of the sample locations see Reimann et al., 2006 or 2007a,b,c). Oslo has about 600,000 inhabitants and is Norway's largest city. Another 1.4 million people, of Norway's total population of 4.5 million, live close to the city. In general the region to the south of Oslo is one of the most intensely utilised and richest agricultural areas of Norway with many interspersed suburban developments. Directly north of Oslo the terrain rises abruptly to an elevation of 400–700 m above sea level (a.s.l.) and land use changes completely. This area, known as “Nordmarka”, is dominated by forests and many lakes and is extensively utilised for nonmotorised recreation purposes by the Oslo population. Forestry is another important activity in the area. Further north, in the Randsfjord area, agriculture is again the dominant land use. In contrast to the Ås area, animal husbandry is widespread and the area is too far inland from Oslo to feel the pressure of spreading urbanisation. The wood samples are thus collected from an area that can be considered as typical for the firewood-production that is burned in the greater Oslo area. The annual precipitation in the investigated area is in the range of 700–1000 mm. The yearly average temperature is 5.7 °C in Oslo. Due to the higher elevation and higher precipitation, the climate in Nordmarka is markedly colder and wetter than along the rest of the transect. For detailed maps on precipitation, temperature and vegetation see Moen (1998). The dominant wind direction is towards the north, i.e. from Oslo towards Nordmarka and Randsfjord.
1.2.
Geology
The transect was selected so that it crossed different lithologies, typical of the greater Oslo area (for a geological map in relation to sample locations see Reimann et al., 2006 or 2007a,b,c). The most remarkable geological feature is the Oslo Rift. A late Carboniferous succession of shale, sandstone and conglomerate (the Asker
S CIE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 3 ( 2 00 8 ) 1 9 1–1 97
Group) marks the onset of deposition related to the Oslo Rift (Dons and Larsen, 1978). The sedimentary sequence was deposited on a levelled surface and is overlain by basalt followed by a thick sequence of latite lava flows (rhomb porphyry) with thin beds of interlayered sand. Further evolution of this Permian rift is characterised by a number of volcanic centres with basaltic to subordinate rhyolitic lavas preserved in collapsed calderas rimmed by major ring faults and dykes. A variety of plutonic rocks cut the units described above (Dons and Larsen, 1978). Precambrian gneisses occur at the southern- and northernmost ends of the transect. During the last ice age the whole study area was icecovered. The thickness of the ice reached its maximum about 18.000–20.000 years ago, and moraine material was deposited in the study area. Due to the melting of the ice the land rose, and the previous coastline in the area can now be found at 150–200 m above sea level (a.s.l.) (Lundmark, 1986). Below this level, large deposits of marine clays can be found. Today most of these areas are used for agriculture and were excluded for sampling. The geological map of Oslo and surrounding areas (Lutro and Nordgulen, 2004) provides an excellent and more detailed geological overview of the study area.
2.
Methods
2.1.
Sampling
Sample sites were selected in accordance with the methods used in the European moss monitoring project (e.g., Rühling, 1994; Zechmeister, 1997). A minimum distance of 300 m to major roads and larger settlements is required, a minimum distance of 100 m to minor roads and houses and a minimum distance of 5 m to forest roads. With two exceptions (site 134, which is only 25 m from a major road (E6) and site 132, which is about 200 m from a major road (E4)) all sample sites, even those in the city of Oslo, fulfil these requirements. Otherwise, at the selected sample density, it was possible to collect the samples at largely “untouched” natural sites in the surrounding forests, which reach into the city, and in large forested parklands. To avoid possible seasonal influences (e.g., Zechmeister et al., 2003) all samples were collected in as short a time span as possible (14 days) in the autumn of 2005. All sites were located in forest ecosystems with a required minimum number of typical (for the area) plant species (presence of Norwegian spruce (P. abies (L.) Karst.), birch (Betula pendula Roth and B. pubescens Ehrh.) and European mountain ash (Sorbus aucuparia L.) in the tree layer; bracken or other species of fern (Pteridium aquilinum (L.) Kuhn), cowberry (Vaccinium vitis-idaea L.), blueberry (bilberry) (Vaccinium myrtillus L.) and terrestrial moss (Hylocomium splendens (Hedw.) Schimp.) in the ground vegetation layer. Even in the city of Oslo it was possible to find sufficiently “natural” forested areas to enable directly comparable sites to be sampled. To collect wood, large twigs (age 10–20 years) were cut from the trees and the bark was completely removed from the twigs using a carbon steel knife. One litre of 2–5 cm long wood cuttings was collected into clear contamination-free polyethylene (PE)-zip-lock plastic bags. Due to the already quite cool fall temperatures there was no problem with condensed water
193
collecting in the bags. In the field the PE bags were marked with the site number and a three-digit code for the material. UTM coordinates of all sample sites were recorded using a GARMIN GPS system. Each sample site was recorded in a series of at least three field photos. Every third day the wood samples were posted to the laboratory for immediate drying. After arrival in the laboratory of the Geological Survey of Norway (NGU) all samples were dried to constant weight at T b 40 °C.
2.2.
Analyses
All dried samples were shipped to ACME laboratories in Canada by courier. Here the samples were pulverised to pass a 100-mesh screen in a mild steel mill. The method has been thoroughly checked to avoid any contamination to the plant samples by the mill material. A 0.5 g aliquot of sample material was first leached with concentrated nitric acid for 1 h and then digested in a hot water bath for an additional hour. After cooling, a modified aqua regia solution of equal parts of concentrated ACS grade HCl and HNO3 and de-mineralised H2O was added to each sample (6 mL/g) to leach in a hot (95 °C) water bath for 2 h. After cooling, the solution was made up to a final volume with 5% HCl and then filtered. The sample weight to solution volume ratio is 1 g per 20 mL. The solutions were analysed using a Perkin Elmer Elan 6000 inductively coupled plasma mass spectrometer (ICP-MS) for 39 elements (Ag, Al, As, Au, B, Ba, Bi, Ca, Cd, Co, Cr, Cu, Fe, Ga, Hg, K, La, Mg, Mn, Mo, Na, Ni, P, Pb, Pd, Pt, S, Sb, Sc, Se, Sr, Te, Th, Ti, Tl, U, V, W, and Zn). Loss on ignition (LOI) was determined gravimetrically following the controlled burning of the vegetation samples at 475 °C. For the purpose of this paper the analytical results of “wet ashing” as described above were recalculated to concentrations in the wood ashes rather than using the “dry weight in wood” results. For Pb the authors tested both, wet and dry ashing procedures for 5 samples from the Oslo transect, and the differences between the analytical results were statistically insignificant (Reimann et al., in press).
2.3.
Quality control
The samples were analysed in one large batch of 400 plant samples (terrestrial moss, birch leaves, wood and bark, spruce needles, wood and twigs, European mountain ash leaves, fern leaves). For quality-control purposes the National Institute of Standard and Technology (NIST) standard reference material 1575a Pine Needles and a laboratory internal reference material, Standard V13 Mountain Hemlock Needles, were included evenly spread over the whole batch 8 and 13 times respectively. Detailed results of quality control are reported in Reimann et al. (2006). Average precision for the elements reported here is better than 5%. For 13 elements of the analytical program more than 50% of the values were below the lower limit of detection: Al (100 mg/kg), Bi (0.02 mg/kg), Ga (0.1 mg/kg), Pd (0.5 μg/kg), Pt (0.2 μg/kg), Sc (0.1 mg/kg), Se (0.1 mg/kg). Te (0.02 mg/kg), Th (0.01 mg/kg), Tl (0.02 mg/kg), U (0.01 mg/kg), V (2 mg/kg) and W (0.1 mg/kg). Thus only the concentrations for 26 elements (Ag, As, Au, B, Ba, Ca, Cd, Co, Cr, Cu, Fe, Hg, K, La, Mg, Mn, Mo, Na, Ni, P, Pb, S, Sb, Sr, Ti, and Zn) are presented here.
194 2.4.
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Data analysis
3.
Due to the nature of the data (40 samples, 26 parameters), compositional data with closure effects, and expected strong spatial dependencies formal statistics and statistical tests were avoided. Rather methods of exploratory data analysis (EDA, Tukey, 1977) and robust measures of central tendency and spread are used.
Results and discussion
The analytical results for birch and spruce wood ashes are presented in Table 1. Fig. 1 shows cumulative probability plots of 8 selected elements in birch and spruce wood ashes, clearly demonstrating the species-dependent uptake characteristics. Table 2 shows the ratio of the median values of birch to the
Table 1 – Element concentrations in 40 samples (birch: 38 samples) of birch and spruce wood ashes, collected along a 120 km long transect in S-Norway Element Ag mg/kg As mg/kg Au mg/kg B mg/kg Ba mg/kg Ca wt.% Cd mg/kg Co mg/kg Cr mg/kg Cu mg/kg Fe wt.% Hg mg/kg K wt.% La mg/kg Mg wt.% Mn wt.% Mo mg/kg Na wt.% Ni mg/kg P wt.% Pb mg/kg S wt.% Sb mg/kg Sr mg/kg Ti mg/kg Zn mg/kg
Material Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce Birch Spruce
DL
Min
5%
25%
Median
Mean
75%
95%
Max
MAD
CVR_%
0.002
1 4 2 2 0.03 0.02 276 268 2060 1450 8.3 24 26 4 3 1 147 183 138 292 0.15 0.09 0.2 0.1 6.4 5.6 1.1 0.9 1.9 2.7 0.56 0.31 1.2 1.0 0.28 0.26 18 14 1.8 0.56 37 8 1.2 1.1 0.9 1.8 488 711 92 91 3910 2630
1.8 5 2.9 2 0.07 0.02 459 278 2280 2810 18 26 30 8 6 1.95 236 199 343 314 0.21 0.1 0.3 0.1 13 6 1.2 1.0 5.5 2.9 0.92 1.5 1.3 1.1 0.46 0.38 24.8 21.9 4.3 0.58 53.7 14.8 1.5 1.3 2.29 1.9 912 1020 230 96.9 5970 3530
2.6 6 6 6 0.11 0.05 606 398 3610 5120 20 28 45 21 20 5 288 246 398 504 0.26 0.13 0.4 0.4 18 9 1.6 1.2 6.6 3.4 1.9 2.3 2.6 1.3 0.52 0.46 31 26 4.9 1 170 37 2.9 2.1 2.6 2.3 1160 1320 268 118 9380 4500
3.3 9 10 12 0.14 0.10 652 484 4560 6620 22 31 57 31 34 8 306 268 473 594 0.31 0.14 0.6 0.6 19 10 2.6 1.3 7.7 4.2 3.7 4 3.3 2.1 0.61 0.53 37 28 6.2 1.5 516 67.5 5.9 2.6 3.1 2.6 1420 1590 306 132 14,600 5060
4 9 15 12 0.23 0.13 689 462 5600 7430 22.3 32 65 37 38 9.93 316 266 465 579 0.39 0.18 0.7 0.6 19.3 10 7.1 1.7 7.7 4.3 3.8 4.4 5.9 2.3 0.61 0.52 49 34 6.4 1.4 965 144 5.7 3.9 2.99 2.51 1490 1710 318 135 14,300 5350
5 13 19 16 0.25 0.14 770 536 6420 9640 24.5 35 76 50 51 12 352 290 515 674 0.32 0.23 0.9 0.8 21 11 4.9 1.4 8.8 5.1 5.4 5.7 6.4 2.6 0.65 0.57 62 40.5 7.5 1.7 834 162 7.5 5.6 3.3 2.8 1670 1940 330 142 16,700 6200
7 15 46 21 0.50 0.19 985 567 11300 13500 28 38 115 73 81 21 450 330 642 782 0.35 0.29 1.3 1.0 24 13 27 2.6 9.9 5.7 7.6 9.1 15 5.5 0.75 0.73 89 60 10 2.2 2320 423 10 8.7 3.5 2.9 2330 3170 494 209 22,600 7590
8 18 52 44 1.5 1.1 1050 588 20700 17800 28.4 43 203 184 98 37 508 342 650 860 0.37 0.3 1.7 1.2 25 13.4 86 15 13 6.8 9.3 12 41 8 1.1 0.82 156 80 10.5 2.4 13700 1290 14 9 7 3 3400 3840 652 268 43,500 7910
1.33 4.3 7.41 7.41 0.09 0.07 110 83.8 1680 3310 3.54 4.02 23.7 14.8 23.7 4.45 51.9 33.4 87.5 123 0.03 0.04 0.45 0.3 2.99 1.75 1.63 0.15 1.55 1.32 2.86 2.69 1.93 1.04 0.08 0.08 16.3 5.19 1.93 0.57 489 63 4.09 1.4 0.3 0.3 430 445 48.9 16.3 5920 1280
40 49 74 62 64 70 17 17 37 50 16 13 42 48 70 56 17 12 18 21 11 27 74 50 16 18 63 11 20 31 77 67 58 50 13 15 44 19 31 38 95 93 69 54 10 11 30 28 16 12 41 25
0.02 0.0002 1 0.1 0.01 0.01 0.01 0.1 0.01 0.001 0.001 0.01 0.01 0.001 0.00001 0.01 0.01 0.1 0.01 0.01 0.01 0.02 0.5 1 0.1
DL: detection limit, Min: minimum, Max: maximum, MAD: median absolute deviation and CVR: robust coefficient of variation.
S CIE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 3 ( 2 00 8 ) 1 9 1–1 97
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Fig. 1 – Cumulative probability plots of 8 selected elements in birch and spruce wood ashes, demonstrating the species-dependent uptake characteristics. The maximum admissible concentrations for sewage sludge used as fertiliser in agriculture are shown in some diagrams as a stippled line.
median values for spruce wood ashes, sorted according to relative enrichment (depletion) in birch. Not unexpectedly there are large species-dependent differences in the median concentrations between birch and spruce wood ashes although the samples were collected at the same sites (see Fig. 1). The major nutrients Ca, K, Mg, P, S and Mn all occur in wt.% concentrations in the wood ashes, median concentrations of Ca reach over 30 wt.% in spruce ashes (Table 1). Barium, Fe, Na, Sr and Zn all reach median concentrations well above 1000 mg/kg in the wood ashes (Table 1). Also the “heavy metals” Cd, Cr, Cu and Pb all reach disturbingly high median concentrations in the wood ashes. When comparing birch and spruce, Pb, followed by P, Co and Zn are most enriched in birch wood ashes, while Ag is
relatively enriched in spruce wood ashes (Fig. 1, Table 2). Lead shows the largest variation of all elements in both, birch and spruce wood ashes (Fig. 1). Lanthanum shows the largest difference in variation between the two species, a high variation in birch wood ashes and almost no variation in spruce wood ashes (Table 2). Maximum concentrations reach surprisingly high levels for some of the toxic trace elements. Lead reaches more than 1.3 wt.% in birch wood ashes (median 516 mg/kg) at site 127 and even 1290 mg/kg in spruce wood ashes from the same site. At this site a small Pb-mineralisation was subsequently detected in the local bedrock during follow-up work (Reimann et al., in press). Such small Pbmineralisations are quite typical for the larger Oslo area (e.g., Nilsen and Bjørlykke, 1991). Neither birch nor spruce is an
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Table 2 – Median concentrations in birch and spruce wood ashes (all in mg/kg) and the ratio birch wood ashes/spruce wood ashes
Ag Ba Ca Cu As Sr Mn Hg Cr Na Sb Ni B Au Mo Cd Mg K La Fe S Ti Zn Co P Pb
Birch
Spruce
Ratio
Median
Median
Birch/spruce
3.3 4560 220,000 473 10 1420 36,500 0.6 306 6120 3.1 37 652 0.14 3.3 57 77,600 188,000 2.6 3060 58,800 306 14,600 34 62,000 516
8.7 6620 311,000 594 12 1590 40,500 0.6 268 5260 2.6 28 484 0.1 2.1 31 42,200 101,000 1.3 1400 26,100 132 5060 8 14,500 67.5
0.4 0.7 0.7 0.8 0.8 0.9 0.9 1.0 1.1 1.2 1.2 1.3 1.3 1.5 1.6 1.8 1.8 1.9 2.0 2.2 2.3 2.3 2.9 4.3 4.3 7.6
The table is sorted according to an increasing median ratio.
accumulator of Pb; spruce clearly avoids uptake of Pb (Reimann et al., 2007a). Larch and pine are two widespread, documented accumulators of Pb, and Pb-concentrations 300 times higher than in other species collected at the same localities are reported in literature (e.g., Kovalevsky, 1987). The maximum concentrations of Cd are also amazingly high in both, birch and spruce wood ashes: 203 mg/kg and 184 mg/kg respectively, without any apparent source. Zinc reaches a maximum concentration of 4.3 wt.% in birch wood ashes and 7910 mg/kg in spruce wood ashes. A general statement that wood ashes have a “relatively low metal content” (e.g., Omil et al., 2007) is thus not justified. For agricultural soils and for sewage sludge that is used as fertiliser in agriculture clearly defined maximum allowable concentrations have been defined for a number of metals (Cd, Cu, Ni, Pb, Zn, Hg and Cr) (Council Directive 86/278/EEC of 12 June 1986). No such levels have been defined for forest soils or for materials (e.g., wood ashes) that are spread in the forest. It is, however, generally forbidden to spread sewage sludge in forests. The concentrations of many elements reported in wood ashes here even exceed the upper limits as defined for sewage sludge quality class III in the Norwegian regulations (FOR 2003-07-04 nr. 951) for use in agriculture. For Cd only 1 sample of the wood ashes is below the upper limit of 5 mg/kg. For Pb about 70% of the birch wood ashes and still 25% of the spruce wood ashes exceed the maximum allowable concentration of 200 mg/kg. For Zn all samples for both species exceed the maximum allowable concentration of 1500 mg/kg.
Some of the wood ashes fall into the “toxic waste” class in recently defined concentration classes for soils in Norwegian kindergardens, playgrounds and school yards in industrial areas (Pb and Zn — ca. 5% of all wood ash samples exceed the levels defined for class 5 (Alexander, 2007)). It appears thus necessary to consider whether these ashes should be treated as dangerous waste rather than as fertilisers (see also discussion in Ribbing, 2007). The picture becomes even more threatening when considering how ashes spread in a forest ecosystem will disturb the natural biogeochemical cycles (see, for example, Reimann et al., 2007b). The ashes have alkaline properties, one reason why they are considered positive for remediating the acidic soil environment in forests (e.g., Jala and Goyal, 2006; Omil et al., 2007). Increasing the pH in a forest ecosystem will boost the decay kinetics of organic material, this can lead to a sudden release of metals. When adding ashes with high metal concentrations many of the metals will strongly bind to the organic matrix. Over time this will lead to a disproportionate enrichment of heavy metals in forest soils and a decrease of binding sites for nutrients. While the nutrients and all other elements, including the “heavy metals” are usually released very slowly (over tens of years) during the natural decay of organic material, they are all at once available when spreading ashes in the forest. As a result there will be a sudden excess of many elements. Local conditions and the different binding strengths of the elements to the soil organic matrix will determine the fate of this excess. None of the possible long-term effects is beneficial. It is of course justified to argue that any forest fire will have the same effect, however, forest fires are a relatively rare event compared to adding thousands of tons of wood ashes as fertilisers to forests. The forests protect groundwater resources. They are only able to do this, when such interferences are kept to a minimum. Usually the organic material in a forest deteriorates quite slowly — over tens of years. Differences in natural conditions (e.g., climate, drainage conditions) can lead to very different metal concentrations in the organic layer of forest soils without any human interference (Reimann et al., 2007b), even if the metal concentrations are more or less the same in the underlying minerogenic soils. It is a major anthropogenic interference with the natural equilibrium of decay and release of organic matter when burning the wood and spreading the resulting ashes in the forest. As a minimum such wood ashes clearly need an analytical certificate. Well considered maximum allowable concentrations for many elements need to be established before such ashes can be spread in forests or any other near-natural ecosystem. The import of metals into forest ecosystems via such “fertilisation” will in any case be much higher than any anthropogenic impact of heavy metals from the atmosphere via long range transport.
4.
Conclusions
Wood ashes are rich in nutrients but can also contain stunningly high natural concentrations of potentially toxic metals. The elements reaching peak concentrations will vary from plant species to plant species. Additionally, wood ashes are alkaline. When spreading wood ashes in a forest, even in the forest from which the wood originated, a one-time addition of these metals
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in combination with changing the pH of the soil is a strong interference with the natural biogeochemical cycles. The use of fly ash as “fertiliser” in forestry clearly needs legislation. It is the metal concentrations and their availability that need to be considered, they are by no means harmless due to their “natural” origin. “Green energy” may not always be as green as it looks at first glance.
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