Soil pH and Ca, Mg, K, P, and S levels increased slightly after ash application. No leaching loss was observed after several applications of ash. Furthermore, microbial activity increased and ammonium and nitrate concentrations decreased as a conse­quence of N immobilization. No effects due to heavy metals were observed in the soil solution, and the availability of Mn and Zn increased slightly after the third application of ash.

3.3.1 Nutrition

The Pinus radiata plantations were initially deficient in P, K, and Mg (Table 6.4), as often found in this type of acid soil plantation (Mesanza et al. 1993; Romanya and

Vallejo 1996). In contrast, foliar N concentrations were sufficient (15 mg g-1). Repeated application of wood ash to the soil did not have any effect on foliar concentration 3 years after the initial treatment. Although the concentrations of basic cations such as Ca2+ and Mg2+ increased significantly in the soil, foliar analysis did not show any significant changes in the needles after the treatment. This may be the result of tree growth and the consequent dilution effect.

Supplemental fertilization with slow release of phosphorus (WAP treatment) increased the foliar concentration of the element throughout the study (Fig. 6.3). The differences between treatments are more significant for the soils over migma — tites than for those over lutites. However, repeated application of wood ash did not increase the levels of P in needles.

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.

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).

10.3.1 Chemical Characterization ofWBFA

As shown in Table 10.1, the WBFA was characterized by a significant presence of heavy metals of particular environmental concern, such as cadmium, chromium, copper, lead, and zinc. Zinc was the predominant heavy metal (2,274 mg/kg), whereas cadmium was the heavy metal with the lowest concentration (9 mg/kg).

With regard to the reuse of WBFA in cement-based materials, it must also be considered that several other chemical species, such as chlorides, sulfates, alkalies, and magnesium oxide contained in this waste (Table 10.1), may exert adverse effects on cement hydration and concrete durability.

It is undoubted that the presence of a significant amount of water-soluble compounds, such as chlorides and alkalies, could promote the formation of a high porosity within the hardened cementitious mixes, thus penalizing mechanical strength development and durability. High contents of water-soluble chlorides can also be deleterious for steel-reinforced concrete, since they will promote the corrosion of iron reinforcing bars. High contents of available alkalies (i. e., the amount of alkalies released into the pore liquid of cementitious matrices) are responsible for the development of deleterious expansion associated with alkali-silica reaction in concretes made with aggregates containing some alkali-reactive forms of silica or silicate, or both. Deleterious expansive phenomena in concrete can also arise from very slow dissolution of significant amounts of sulfates or magnesium oxide, with subsequent precipitation of very expansive phases, such as ettringite (CaSO4-32H2O) and brucite [Mg(OH)2].

In the USA, the current restrictions concern the use of fly ashes originating from combustion of fossil fuels (bituminous and subbituminous coal, peat, and lignite), and these restrictions are specified in the ASTM C 618 method. In Europe, the current restrictions concern both the fly ashes derived from fossil fuels and those resulting from the combustion of biomass and fossil fuel blends (cofiring), with a biomass content not higher than 20 wt%. These restrictions are specified in the EN 450-1 method.

For the above reasons there exists no standard method dealing with the chemical requirements for reuse of fly ash from pure biomass combustion in cementitious mixes. However, it is reasonable that the quality of fly ash from pure biomass burning should follow the same regulation as fly ash from fossil fuel combustion.

Table 10.3 compares the specific chemical characteristics of the WBFA (also reported in Table 10.1 in terms of elemental composition) with the corresponding

limits (chemical requirements) specified by ASTM C 618 and EN 450-1. In this table, the chemical characteristics of the solid residue resulting from the elution of WBFA with deionized water (elution test) are also reported.

The chemical characteristics of this solid residue, also referred to as washed WBFA, were evaluated on the basis of the results of the elution test reported in Fig. 10.2, in terms of percentage removals of alkalies, chlorides and sulfates, and percentage weight loss of WBFA.

As shown in Table 10.3, with exception made for the loss on ignition, the American specifications appear to be less severe than the European specifications. In particular, ASTM C 618 does not restrict chloride, free calcium oxide, magne­sium oxide, and soluble phosphate contents. Moreover, the sulfate limit (5%) is more than the maximum allowable (3%) by European regulation. As far as the alkali content is concerned, the American limit [1.5% (w/w) as available alkalies] appears to be comparable to the European limit [5.0% (w/w) as acid-soluble alkalies] if it is considered that, depending on the type of coal fly ash, the content of available alkalies, as determined by the ASTM C311 test method, may vary from 20 to 50% of total alkalies (acid-soluble alkalies), as determined by the EN 196-2 test method (Berra et al. 1992).

According to the chemical requirements prescribed by EN 450-1, which are mostly more severe than those required by ASTM C 618, and remembering that, to date, these requirments do not apply to pure biomass fly ashes, the WBFA was found to meet all chemical requirements, except for the chloride content (1.07% against the limit of 0.10%) (Table 10.3).

In spite of the high release of water-soluble chlorides (82.2% removal) accom­panied by a relatively low percentage weight loss of fly ash (12%) (Fig. 10.2), even the washed WBFA failed to meet the chemical requirement for chlorides, but the chloride content (0.20%) was only slightly higher than the limit of 0.10% (w/w) (Table 10.3).

As compared with WBFA, washed WBFA was also characterized by much lower contents of alkalies [1.43% (w/w) Na2Oeq against 3.7%] and sulfates [0.75% (w/w) SO3 against 2.9%]. Conversely, washed WBFA was more rich in MgO [3.9% (w/w) against 3.5% for WBFA], but the MgO content remained below the limit of 4.0% (Table 10.3). The washed WBFA could also be richer in heavy metals than WBFA, in consideration of the expected low release of these substances into aqueous solutions.

As far as the use of washed WBFA in cement-based materials is concerned, it is likely that a washing treatment of fly ash with a liquid-to-solid ratio above 10 L/kg could reduce the chloride content below the 0.1% limit (Table 10.3). In that case, the wastewater resulting from the washing of fly ash could contain chloride and sulfate concentrations below the limits established for wastewater disposal. How­ever, this wastewater should be treated for pH correction and, probably, for heavy metal removal. In this regard, the pH of WBFA, defined as the pH of the aqueous suspension of fly ash with an L/S ratio of 10 L/kg, was 12.9. This high pH, which is compatible with the use of WBFA in cement-based materials, was attributable to the release of alkalies and calcium oxide from fly ash.

The electrical conductivity of the eluate (1.53 S/m) evidenced significant release of electrolytic compounds from WBFA.

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).

11.4.1 Landfills and Mining Waste Deposits

Combustion residues are used to provide a barrier against the penetration of water or oxygen, or both, into the body of the landfill or of the heap or impoundment of mining waste. The properties that make fly ash from solid biofuels suitable are their high pH and their self-binding properties. The potential for use in several contexts is quite high and the demand could easily exceed the availability of suitable residues.

Fly ash mixed into sewage sludge in equal dry substance proportions raises the pH of the sludge, thus preventing its biological degradation. The percolation rates achieved in field experiments with fly ash mixed into sewage sludge are on the order of 12 l/m2/year, both initially and after a few years, which is sufficiently low for sealing layers on landfills for non-hazardous waste (Carling et al. 2006; Macsik et al. 2005). Its shear strength is acceptable and it withstands settling in the body of the landfill.

On the Tveta landfill, the functional requirements on a sealing layer are fulfilled using monolithic layers of ash through diffusional processes, i. e. particle size distribution, moisture content and reactivity of the combustion residues, all con­tributing to the minimisation of the pore volume (Tham and Andreas 2008).

A sealing layer of fly ash from solid biofuels, covered by a protective layer of vegetated sewage sludge, prevents oxygen from reaching sulphidic tailings from mining. The experimental object is the 80-ha large tailings impoundment of Gillervattnet, where an experimental field covering 3 ha was established in 2003-2005. The hardened layer of ash and the toxicity of the fresh ash prevent roots from breaking through the sealing layer and providing channels for air entry.

However, reed canary grass has some capacity to weaken even hardened fly ash sealing layers with a resistance of approximately 5 MPa. The study suggests that the secretion of saccharides by some plant roots may contribute to this effect (Greger et al. 2006, 2009). It is thus best to avoid these particular plants.

Repeated application of ash did not produce significant changes in foliar concen­trations of any of the elements. However, the application produced an increase in the concentration of Zn, Cu, and Ni in needles in the plots on lutites (Table 6.5). This was not observed in 2006. The foliar concentration of Cd was closely related to soil acidity and therefore the concentration of Cd decreased from the third applica­tion of ash onwards.

The filter ash is transported by means of a heavy lorry trailer from the bioenergy plant to a mine in Germany. The distance is assumed to be 250 km. The amount of ash transported is 6.7 kg, which corresponds to 1.68 t km. The total PK nutrient demand has to be provided by synthetic fertilizer (243 kg) under the same condi­tions as described in scenario 1. The transport distance corresponds to 194 t km. In Table 8.11, an overview is given of the total emissions in scenario 2.

8.4.3 Comparison of the Two Scenarios

In Table 8.11, the emissions caused by the two scenarios are compared. The emissions caused by scenario 1 and by scenario 2 are about the same, which is not surprising because of the low level of replacement of the PK fertilizer that can be obtained. However, the potential CO2 reduction per kilogram of filter ash is significant, namely, about 0.40 kg CO2/kg filter ash.

In this study some assumptions have been made. In the sensitivity analysis, the influence of these assumptions on the emissions is determined.

• All of the phosphorous and potassium present in the ash is available as nutrient. If the availability of both elements drops below 40-45%, then the CO2 and SO2 emissions in scenario 1 will be higher than those in scenario 2.

• The transport distance within Cote d’Ivoire is 800 km. If the transport distance is 500 km, the CO2 emissions in scenarios 1 and 2 decrease to 160.3 and 161.9 kg, respectively. If the transport distance increases to 1,000 km, the CO2 emissions are 179.3 and 180.9 kg, respectively. This means that the absolute value in both scenarios will be influenced to a small extent, but the difference between the values in both scenarios remains the same.

• The transport distance within Cote d’Ivoire is the same for the filter ash and the PK fertilizer. If the transport distance of the ash is about 1,500 km more than that for the PK fertilizer, then the CO2 emissions in scenario 1 will be higher than those in scenario 2.

Aluminum is a light metal that makes up 7% of the earth’s crust and is the third most abundant element, after oxygen and silicon, plant roots are therefore almost always exposed to aluminum in some form (Ma et al. 2001). Aluminum exists in soils in many harmless forms, including hydrous oxides, aluminosilicates, sulfates, and phosphates (Haynes and Mokolobate 2001). Aluminum inhibits plant growth by affecting plant roots and development (Delhaize et al. 1993; Ma et al. 2001), inhibiting both cell divisions in the apical root meristem and cell elongation (Blamey et al. 1983). Roots become stubby and brittle; root tips and lateral roots become thick and occasionally necrotic brown (Mossor-Pietraszewska 2001). These effects restrict the ability of the plant to take up nutrients and water, leading to nutrient and/or water stresses (Rout et al. 2001; Haynes and Mokolobate 2001). Plants in acid soil, owing to aluminum solubility at low pH, exhibit a variety of nutrient-deficiency symptoms, with a conse­quent decrease in yield. Aluminum toxicity is linked with phosphorus, calcium, magnesium, or iron deficiency syndrome (Rout et al. 2001). To overcome the lack of productivity in aluminum toxic soil, the first step is to treat acidity, which will promote better root growth and function and will allow nutrients and water to be taken up more effectively by the plant.

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.