The heavy metals that may accumulate in wood ash are of special concern when it is used for fertilisation purposes. Compared with coal ashes, ashes derived from wood are lower in heavy metals, but are more alkaline (Campbell 1990). High concentra­tions of As, Cd, Cr, Pb, Zn and Cu may, however, occur owing to the incineration of surface-treated waste wood and wood treated with industrial preservatives (Krook et al. 2006). Cu concentrations in biomass ashes were frequently shown to exceed critical values according to national regulations in Austria (Neurauter et al. 2004) and Germany (Ministerium fur Umwelt und Verkehr Baden-Wurttemberg 2003). Average Cu contents found in three studies dealing with ash composition are illustrated in Fig. 1.1 (Neurauter et al. 2004; Niederberger 2002; Toth(Sva 2005). Whereas the incineration of pure wood led to moderate Cu concentrations in the resulting ash, high Cu contents were found when other biomass, especially roadside greenery and material derived from wood processing, was combusted together with natural wood (samples 9, 10).

Wood ash is better applicable for fertilisation purposes if it is separated into fly and bottom ash during combustion, as heavy metals — except for Zn — accumulate in the fly ash (Pitman 2006; Stockinger et al. 2006). Fly ash is the lightest fraction formed during combustion, being deposited within the boiler or in the filters (Pitman 2006). Ashes may also include organic pollutants such as polychlorinated dibenzodioxin, biphenyls, dibenzofuran and polycyclic aromatic hydrocarbons (PAHs), which are of interest because of their toxic, mutagenic and carcinogenic effects (Lavric et al. 1994; Enell et al. 2008). High amounts of PAHs are ascribed to a poor combustion performance (Sarenbo 2009). Wood ashes may pose a risk not only because of the direct input of organic pollutants, but also because a rise in soil pH following wood ash amendments enhances remobilisation of PAHs and polychlorinated biphenyls (Bundt et al. 2001). An elevated pH also affects metal solubility in soil; however, changes in solubility do not necessarily correlate with incorporation of heavy metals in plants grown on the respective soils (Dimitriou et al. 2006). Another essential issue in regard of ash amendments to soils is leaching of toxic substances to the ground­water (Williams 1997), especially in combination with an elevated pH and high Na content (Morris et al. 2000). Leaching is frequently evaluated in laboratory tests, but

sample

these tend to overestimate or underestimate on-site leaching processes and thus it is difficult to assess the real situation in the field (Reijnders 2005).

Analysis of variance was used to test for effects of multiple wood ash application on soil chemical properties and foliar concentration of macroelements and microele­ments, with the PROC GLM procedure of SAS (SAS Institute 2004).

Normal diameter and total height were measured annually in all trees within the plots at the end of the growing season for a period of 3 years. The breast diameter was measured in two directions with a caliper, to an accuracy of less than 1 mm. Tree height was measurement with a Vertex III hypsometer. A volume equation based on the allometric model of Schumacher and Hall was used to calculate the tree volumes. The parameters of this model were estimated by Castedo (2004) by use of the following expression:

v = 0.000048 x d20062 x ht0’86691,

where v is the bark volume (m3) of individual trees; d is the breast diameter in centimeters, and ht is the total height in meters. The total height and breast diameters were compared by repeated measurement analysis with the PROC GLM procedure of SAS (SAS Institute 2004) after prior confirmation of the assumptions of equal (Levene test), normal (Kolmogorov-Smimov test), and inde­pendent variance. The mathematical model used was as follows:

yij = m + pxit + Ti + Dj + Ti x Dj + eij,

where yij is the random variable representing the value in the jth observation of the ith treatment, m is a constant representing the mean response of the variable; pxit models the linear relationship between the response and the covariate (initial height or diameter measured before the application of ashes), Ti and Dj are the effects of treatment i (control, WA, and WAP) and the time j (0,1,2,.. .4 measurements), respectively, Ti x Dj is the interaction effect of treatment i by time j, and eij is the experimental error.

As already stated, the functional unit in this study is the fertilization of 1 ha of land in Cote d’Ivoire, on which cacao trees are grown together with shadow trees. In Table 8.10, an overview is given of the process steps for the two scenarios. As can be seen from Table 8.10, several process steps are identical for both scenarios. This means that these process steps are not taken into account.

8.4.2 Scenario 1: Recycling of the Ashes as a Fertilizer

The filter ash is transported by means of a heavy lorry trailer from the bioenergy plant to the port. The distance between these is assumed to be 100 km. Table 8.3 gives the emissions per ton kilometer. Per hectare cacao plantation in Cote d’Ivoire, 74 kg cacao shells are produced, resulting in 6.7 kg filter ash, corresponding to 0.67 t km. The filter ash is transported by means of a bulk carrier from the Dutch port to Cote d’Ivoire. The distance is assumed to be 6,000 km (SenterNovem 2007), which results in 40.21 km. The ash is transported by means of a medium-sized lorry from the port in Cote d’Ivoire to the cacao plantation. The distance is assumed to be 800 km, which results in 5.4 t km.

The major part of the PK nutrient demand has to be provided by synthetic fertilizer. Table 8.7 shows the emissions produced by 1 kg of these fertilizers. These emissions are valid for production of the fertilizer in Europe. We assume the production of the fertilizers takes place in Cote d’Ivoire. However, because of lack of better data, the production data for Europe are used. The fertilizer (238 kg) is transported within Cote d’Ivoire by means of a medium-sized lorry. The distance is assumed to be 800 km, corresponding to 190 t km. In Table 8.11, an overview is given of the total emissions in scenario 1.

2.2.1 Treatments and Experimental Design

Two field experiments were conducted on different soil types at the agricultural experimental stations of the University of Rostock (loamy sand) and at the Institute of Organic Farming in Trenthorst (sandy loam). The same soil types were also used for the pot experiments in Rostock; in 2007 the experiments were established on loamy sand and in 2008 they were established on sandy loam (Table 2.2).

The rape meal ash (RMA) was produced at the University of Rostock in a fluidized bed combustion at 860°C. The rye straw ash (SA) was produced via grate firing at 750°C and was supplied by the Leibniz Institute for Agricultural Engineering in Potsdam-Bornim (Germany). The rye cereal ash (CA) was manu­factured at the Agricultural Technical School of Tulln (Austria) also via grate firing at 650-850°C. The fertilization treatments in the field and pot experiments were established in respect of the nutrient contents of the ashes (Table 2.3). Heavy metal contents are given in Table 2.4.

In the 2-year field experiments with different crop plants (see Table 2.5) the three ashes were applied once at the beginning of the experiments and incorporated into the top soil. Nitrogen was given in all treatments according to good fertilization practice.

In the pot experiments six different fertilization treatments were established. Besides the ash treatments, other treatments included triple superphosphate (TSP) as a highly soluble P source, potassium chloride (KCl) as a highly soluble potas­sium source, and a control (CON) without P and potassium. The ashes/fertilizers were applied on the soil surface and mixed into the upper 5 cm of soil. For nitrogen supply, 1.4 g NH4NO3 per pot was given. Mitscherlich pots with 6 kg soil each were used for crop cultivation.

Eight different crops were investigated in the pot experiments. Depending on the favourable growing time of cultivated main crops and catch crops, two experiments per year were established (Table 2.5). The main crops were seeded in April and the catch crops were seeded in August. The crop growing period in the pot experiments was about 7-10 weeks until the time of maximum biomass. In field and pot experiments all treatments were replicated four times.

Acidification involves four main processes, which are (1) leaching of basic cations (Ca2+, Mg2+, K+, and Na+) from the exchangeable complex of the soil (clay, minerals, humus) and their substitution by protons (H+) and cation acids (SU+, Fe3+, Al3+), (2) accumulation of potentially toxic heavy metals (Co, Cd, Zn) in deeper soil layers, (3) accumulation of aluminum sulfates after saturation of exchange sites with Al ions that allows transportation of acidity to deeper soil layers or groundwaters, and (4) transfer of potentially toxic cation acids (Al, heavy metals) toward groundwater and surface water owing to the increased solubilization of compounds formed by these cation acids under very acidic conditions (Mayer 1998). Acid soil infertility is a major limi­tation to crop production on highly weathered and leached soils throughout the world (Von Uexkiill and Mutert 1995). It is a complex interaction of growth-limiting factors including toxic levels of aluminum, manganese, and iron, as well as deficiencies of some essential elements, such as phosphorus, nitrogen, potassium, calcium, and magnesium, and some micronutrients (Kochian et al. 2004). Among these constra­ints, aluminum toxicity and phosphorous deficiency are the most important owing to their ubiquitous existence and overwhelming impact on plant growth (Kochian et al. 2004).

Table 10.4 gives the results of the monolith leaching test on cubic specimens of cement pastes (water-to-binder weight ratio 0.50) made with blended cement [70% (w/w) Portland cement and 30% (w/w) WBFA]. In this table, the concentrations of selected heavy metals (Cd, Cr, Cu, Ni, Pb, and Zn) in each of eight leachates are reported as the average values of three replicate leaching tests.

With respect to the other heavy metals investigated, the higher concentrations of copper, lead, and zinc measured for most leachates were directly related to their higher contents in the original fly ash (Table 10.1).

Using the data in Table 10.4, we calculated the leaching rate of each of the selected heavy metals as an average within each leaching period, and these rates are reported in Fig. 10.3 for each of the eight leachant renewals.

For copper, lead, and zinc, the leaching rates dramatically reduced after the first leachant renewal (the first two renewals for Zn), thus revealing the existence of two different mechanisms governing the leaching process of such heavy metals. At early leaching times (first two renewals), the controlling mechanism appeared to be the release of heavy metal from the outer surface of the monolith specimen by dissolution into the leaching solution or by wash-off, or both. At longer leaching times, the release was probably controlled by diffusion, and the heavy metal ions had to migrate within the pore liquid of the cementitious matrix of the test specimen prior to reaching the liquid bulk. As a result, this leaching phase was characterized by a much lower rate as compared with the initial leaching phase. In the case of
cadmium, chromium, and nickel leaching, no dissolution/wash-off phenomenon was detected during the early release phase.

As shown in Fig. 10.3, after the first or the second leachant renewal, the release rate of each heavy metal did not significantly vary with increasing leaching time. Thus, the high concentrations of heavy metals measured for the seventh and eighth leachates (Table 10.4) were attributable to the much higher contact times between the specimen and the leachant (20 and 28 days for the seventh and eighth renewals, respectively).

With use of the results in Table 10.4, the cumulative mass of each heavy metal released per unit exposed surface area of specimen, Mt (mg/m2), was also calculated and is plotted in Fig. 10.4 as a function of the square root of the cumulative leach time, t (h1/2).

For the leaching of cadmium, chromium, and nickel, there existed straight line relationships between Mt and t1/2, with no intercept on the coordinate axis. This is typical of leaching processes controlled by the diffusion mechanism. Conversely, for copper, lead, and zinc leaching, linear relationships with positive intercepts on the ordinate axis were obtained. This is typical of leaching processes controlled initially by dissolution or wash-off phenomena, or both.

To predict the long-term release of copper, lead, and zinc from monolithic speci­mens, the leaching data in Table 10.4, relative to these metals, were considered over the leach time interval for which diffusion was the release-controlling mechanism. In other words, the first two leachant renewals were considered as preconditioning steps for the subsequent leaching test. In this way, straight line relationships between Mt and t1/2 were obtained for the release of copper, lead, and zinc, as shown in Fig. 10.5.

With use of the linear regression equations resulting from the data in Figs. 10.4 and 10.5, the releases of heavy metals after 100 years of leaching were estimated

and compared with the standard limits (Category I applications) as specified in the Dutch Building Materials Decree (1995). These specifications are commonly taken as a reference for evaluating the environmental quality of cement-based materials incorporating hazardous wastes. Figure 10.6 compares the estimated releases of the selected heavy metals with the Dutch standard limits.

As can be noted, all the releases were well below the corresponding regulatory limits and this proved the good immobilization capacity of heavy metals by the cementitious matrix investigated and, consequently, the good environmental qual­ity of the blended cement formulated with 30% (w/w) WBFA.

10.2 Conclusions

The WBFA is characterized by a significant content of heavy metals of particular environmental concern, such as cadmium, chromium, copper, nickel, lead, and zinc, and by a remarkable amount of water-soluble compounds, such as alkalies, chlorides, and sulfates.

According to the European chemical requirements established for reuse of coal fly ash as a mineral admixture in cement-based materials, the biomass fly ash studied appears to be suitable for the formulation of blended cements, provided that its chloride content be preliminarily reduced.

As indicated by the results of the water elution test on WBFA, a single-stage washing treatment of this ash with deionized water might be sufficient to reduce the chloride content to acceptable levels.

As evidenced by the results of the monolith leaching test on hardened pastes of blended cement [70% (w/w) Portland cement-30% (w/w) WBFA], in spite of the high content of water-soluble compounds of WBFA and the acid pH conditions of the leachant throughout the test (pH 6.0), very low releases of heavy metals were always obtained, thus revealing a high metal immobilization capacity by the cementitious matrix and, consequently, a good environmental quality of the blended cement investigated.

For some heavy metals such as copper, lead, and zinc, the release from a monolithic specimen appears to be governed by two different leaching mechan­isms: dissolution/wash-off at earlier leach times and diffusion at longer leaching times. Conversely, in the case of cadmium, chromium, and nickel leaching, no dissolution/wash-off phenomenon was detected.

The functioning of ecological systems can be studied at small spatiotemporal scales in controlled laboratory microcosms. In particular, the microcosm method has been used as a tool to understand the functioning of decomposer food webs (Scheu 2002; Huhta 2007). Although the experimental approach has limitations such as limited community composition and restricted mobility of animals (Kampichler et al. 2001), idealized model systems help to reduce the natural variability and exclude variables which are considered beyond the scope of the research. The dynamics of decomposer food webs and nutrients in such systems are fairly well known. Decomposer population dynamics in laboratory microcosms containing no auto­trophs, and hence no carbon input, consists of one growth phase and a subsequent decline phase (Nieminen 2008b). The length of the population growth phase depends on the initial population size (Nieminen 2002). For example, the biomass of fungal-feeding nematodes increased exponentially for 3.5 weeks in organic forest soil, then crashed and remained low for 16 weeks (Nieminen 2008b). In a laboratory experiment using larger pieces of forest floor, nematode populations crashed after 3 weeks, then collembolan and enchytraeid populations crashed (Huhta et al. 1983). In such a system lacking carbon input and nutrient uptake by plants, inorganic N accumulates in soil (Nieminen 2008b).

In such severely carbon limited systems, addition of organic carbon available to microbes increases microbial biomass, and reduces the N concentration in soil (Sparling and Williams 1986; Schmidt et al. 1997; Dunn et al. 2006). In an early experiment, glucose increased both the biomass of fluorescein diacetate active mycelium and yeasts as well as fungal-feeding nematodes in pine microcosms (Baath et al. 1978). However, the soil properties were quite unnatural, for example the pH was high (above 6.8), and algae thrived in microcosms but enchytraeids were lacking (Baath et al. 1978). In coniferous forest soil, the pH and moisture usually limit yeast and algal growth. Extra carbon input is also reflected in the biomasses of higher trophic levels of the soil decomposer food web. Addition of carboxymethyl cellulose to microcosms containing mineral and organic soil, needle litter and a Scots pine seedling increased the biomass of both saprotrophic fungi and hyphal-feeding nematodes (Nieminen and Setala 2001). Although the biomass of filamentous fungi increases after cellulose addition, addition of labile carbon such as sucrose can enhance the growth of early successional (r-strategist) microbes such as bacteria (Moore-Kucera and Dick 2008; Nottingham et al. 2009) and Zygomy — cota fungi (Hanson et al. 2008). Nieminen (2010) found that a sucrose addition equalling 100 kg C ha-1 maintained a stable nematode population for one growing season, and an increasing enchytraeid population. Sucrose addition did not alter net N mineralization rates, indicating that N mineralization by increased animal popula­tions exactly balanced the N immobilization in microbial biomass. When the carbon addition rate is increased above 100 kg C ha-1 in Norway spruce forest soil, soil animals, which have orders of magnitude lower growth rates than microbes, cannot consume all the extra microbial biomass in one growing season, and as a conse­quence, N is immobilized in microbial biomass.

Wood ash is a by-product of the wood industry resulting from burning of wood residues for energy production (Nkana et al. 2002). Most of the inorganic nutrients and trace elements in wood are retained in the ash during combustion; the quality of the end product depends on the quality of the wood, the tree species, and the burning process (Perkiomaki et al. 2004). The ash and the metal contents are generally higher in bark than in stemwood (Hakkila 1986; Werkelin et al. 2005). Fly ash contains higher levels of dioxins and heavy metals than the bottom ash (Pitman 2006). For agricultural and horticultural purposes, only bottom ash should be used according to Stockinger et al. (2006). Wood ash is a significant source of the nutrients phosphorous, potassium, magnesium, and calcium, and its properties resemble those of lime (Naylor and Schmidt 1986; Ohno and Erich 1990). Ash fertilization can compensate for the nutrient losses caused by harvesting operations, nutrient leaching, and soil acidification (Saarsalmi et al. 2006).

Although roads are not a particularly suitable use for residues for logistical reasons, they are the type of civil works with the best developed set of requirements, against which constructions with combustion residues may be assessed. Fly ash, preferably from solid biofuels or sludges from the pulp and paper industry, has been used in non-surfaced gravel roads and bottom ash is now also beginning to be used.

Non-surfaced roads have been built using fly ash in Sweden for some time. However, through the introduction of the Finnish experience with this technology (Lahtinen 2001) as the starting point within the Ash Programme, a convenient impulse was given to renewed development. In successive projects, fly ash from biomass has been characterised, recipes have been developed and a few test roads have been built (Macsik and Svedberg 2006; Macsik et al. 2009). A short summary of the results is as follows: bearing capacity and freeze-thaw resistance have increased, fly ash replaces natural materials of about twice its volume, which leads to a significant reduction in weight and height. For the good properties of biomass fly ash to be optimally exploited, and for conservation of fly ash resources, ash should be used to stabilise bad or worn-out road materials. Adverse impacts on the environment could not be observed during the monitoring of construction and use of the roads.

Bottom ash from solid biofuels is also used for roadwork. An example is a private road north of Norrtalje, where approximately 5,000 t of mixed bottom and fly ash from a grate furnace have been used since 2006; see Fig. 11.1. It was possible to run 40-t lorries on the road even when it was flooded by melting snow, which is remarkably good. Infiltration in the body of the work is very slow, which indicates that the environmental impact should be minor. The impact of an ash pile stored for 7 years on the place has been investigated: the uptake of heavy metals in ash by plants and berries did not lead to any increased levels in the plants.

The ash of solid biofuels also has binding properties and it has been utilised in concrete applications. One project involved replacing Portland cement in panel stope mine filling, where large blocks or “stopes” of ore are removed, creating a large cavity. These stopes are backfilled using concrete (cement and mine tail­ings) to stabilise the mine. In full-scale trials, biomass fly ash from grate furnaces could replace 50% of the Portland cement. The other use of solid biofuel ashes demonstrated in the Ash Programme is as filler in low-quality concrete. However, the chloride content of the ash may pose corrosion problems for steel reinforce­ment bars.

Fig. 11.1

As versatile as wood ash is, its potential areas of application are:

• Ash application in forest ecosystems

Wood ash is commonly applied to forest ecosystems to return nutrients extracted through whole-tree harvesting and to counteract soil acidification (Sect. 1.3.1).

• Wood ash as fertiliser or fertiliser supplement in agroecosystems

Wood ash rich in nutrients but displaying a low concentration of heavy metals or organic pollutants is also suitable as fertiliser or fertiliser supplement for agricul­tural and horticultural purposes (Sect. 1.3.2).

• Wood ash for geotechnical constructions and industrial processes

Typical applications in this field are the construction of roads and parking areas, the use of ash as a surface layer in landfills and admixture of ash for concrete, brick or cement production (Sect. 1.3.3).

1.3.1 Ash Application in Forest Ecosystems

The effect of wood ash application on forest ecosystems has been intensively studied in northern European countries where ash is used as fertiliser in boreal forests (Aronsson and Ekelund 2004). Owing to extensive forest harvesting (espe­cially whole-tree harvesting), reuse of ashes was established to avoid base element depletion of forest soils, leading to increasing acidity as well as decreasing amounts of nutrients and organic matter in the soil, thus threatening forest productivity (Stupak et al. 2008).