1. Introduction
Europe’s water is becoming a critical resource to be preserved. Economic activities, urbanization, and global warming affect the quality and availability of European freshwaters. For that, the collection and treatment of waste waters is one key element in the development of a sustainable water cycle [
1]. In the European Union (EU) context, industry is considered a key stakeholder in water consumption, from both a quantitative and qualitative point of view. The uptake of water by industry in Europe represents about 54% of the total uptake for human activities, equal to about 96 billion cubic meters per year [
2].
According to the Urban Waste Water Treatment Directive, the wastewater treatment processes in the EU have strongly improved from 1990 to 2014 [
3]. and the number of wastewater treatment plants (WWTPs) has increased. Nowadays, 26523 urban wastewater treatment plants are operating in the EU [
4]. In addition, at least 30,000 large industrial facilities with related wastewater treatment systems should be considered [
5].
The growing number of wastewater treatment plants in the EU brings about a rapid increase in the amount of sludge being produced. In this respect, the annual production of dry sewage sludge in the 32 member countries of the European Environment Agency (EEA) was estimated at around 11.1 million t in 2018 [
6]. Of the total sludge produced, 34% was used in agriculture, 31% was incinerated, 12% was used in compost and other applications, 12% was disposed of in landfills, and 10% was used in different ways [
7].
The costs attributed to sludge management range from 20% to 65% of the overall expenses of a wastewater treatment plant [
8], and they are mainly related to the high disposal cost, which currently ranges between EUR 160 and 310 t
−1 [
9]. In fact, sludge contains, on a dry basis, 50–70% organic matter, 30–50% inorganics, 3.4–4.0% N, 0.5–2.5% P, 1–3% Al, and micronutrients [
10,
11], but also sulfur, heavy metals, and organic compounds that are toxic for the environment, such as polycyclic aromatic hydrocarbons (PAHs), surfactants, pharmaceuticals, and others [
12].
Another relevant cost item for WWTPs is energy consumption, which contributes to WWTP operational costs by about 7–33% [
13]. For this reason, to improve the economics and sustainability of both industrial and municipal WWTPs, sludge management should be based on the effective recovery of nutrients and carbon while minimizing energy consumption. The recovery of phosphorus from sludge is particularly supported by several EU member states. Phosphate rock extraction increased from 3.4 million metric t in 1913 to 245 million metric t in 2014 [
14], with a projection of 425 million metric t by 2050 [
15], so that the EU has identified phosphate rock and phosphorous among the 34 critical raw materials of high importance for the EU economy and of high risk associated with their supply [
16].
However, Regulation (EU) 2019/1009 on fertilizing products (EU FPR) [
17] excludes both municipal and industrial sludge from being used for CE organic fertilizer production. Despite the fact that sludge is still used in agriculture, or at small percentages, for compost production in many EU countries, EU FPR prevents compost, digestate, or biochar from sludge from being certified as CE fertilizing products. On the contrary, inorganic fertilizers and precipitated phosphate salts, extracted from wastewater, are strongly promoted [
14,
15,
16,
18,
19].
Organic carbon is the other key element of sludge. In fact, organic carbon does not only represent the main element contained in sludge, but it is also a high-value product today. In fact, the carbon value is strictly related to that of CO
2. By the European Union European Trading Scheme (EU ETS) regulation, which establishes the market mechanism attributing CO
2 a price [
20], any ton of CO
2 emitted by the steel, energy, or fertilizer sector corresponds to an allowance that the company in that sector must purchase. The stricter measures adopted by the recently updated version of the EU ETS made CO
2 prices reach prices between EUR 80 and 100 t
−1 [
21]. Since the EU ETS will include municipal waste incineration plants as of 2026, any ton of carbon in the sludge, if incinerated, will represent a high cost for the incineration plant (around EUR 240 t
−1, at the current CO
2 price of EUR 80 t
−1). An alternative use, preferably a carbon storage solution, could reduce fossil coal consumption and avoid CO
2 emissions. In the context of reducing sludge volumes and maximizing the recovery of sludge, water removal represents the first challenge. To this end, innovative sludge dewatering and pre-treatment technologies, such as vacuum preloading, freeze-thaw, and Fenton treatment, have been recently investigated, with promising results in terms of water removal and low energy consumption [
22,
23]. The results achieved by advanced dewatering systems must be considered as the starting point for a sustainable valuation of sludge as a source of valuable materials.
Thermochemical treatments of sludge, such as hydrothermal carbonization (HTC), slow pyrolysis, and hydrothermal liquefaction (HTL), have been the subject of several studies as an alternative to its incineration [
24,
25].
Among them, the HTC and slow pyrolysis processes are known as “carbonization processes”, due to their main aim of maximizing carbon recovery as a solid material.
The HTC process works at temperatures in the range of 180–250 °C and pressures in the range of 10–50 bar and operates in a reaction environment characterized by the presence of liquid water [
13]. The two main products of HTC are the solid, named “hydrochar”, which is a carbonaceous, highly volatile solid, namely >50% dry basis (d.b.), and the water phase, rich in organic carbon [
26]. For this reason, HTC is also seen as a pre-treatment preceding the application of other recovery processes, such as anaerobic digestion of the water phase and pyrolysis of the hydrochar [
27,
28].
Slow pyrolysis needs sludge to be previously dewatered by drying or other pre-treatment systems. The process works at higher temperatures, in a range between 400 and 700 °C, with a residence time in the order of hours and heating rates between 1 and 20 °C/min [
29,
30]. The process always generates three products: a solid carbonaceous matrix (char), a mixture of condensable vapors (water and organic compounds), and a mixture of permanent gases (CO, CO
2, CH
4, H
2, etc.) [
31]. After pyrolysis, most of the sludge’s ashes are concentrated in the char, which usually represents more than 40% d.b. of the processed material, rising to over 50% d.b. in the case of ash-rich municipal sludges [
32]. Consequently, an effect of treating sludge by slow pyrolysis is the increase of the ash content in the produced char and often, by consequence, a reduction of the carbon content in comparison with the processed feedstock.
Char is a carbonaceous product, rich in inorganics when produced by sludge, hydrophobic, and with a low volatile content. Phosphorous, as well as calcium, iron, and silicon, are more concentrated in char compared to the processed feedstock. Due to its high ash content and its low heating value (9–14 MJ kg−1 d.b.), char from sludge slow pyrolysis is neither suitable as a solid biofuel nor applicable in the steel industry. However, the high concentration of both inorganic compounds and carbon makes sludge-derived char an interesting raw material that is usable as a source of nutrients and renewable carbon.
The scientific challenge of sludge management consists then in the valuation of resources from this waste, including carbon and inorganic compounds (with phosphorous in the first place). To this end, a chemical char upgrading step to extract phosphorus and other inorganic compounds could represent not only a promising solution to ensure a full valorization of these raw materials but also an opportunity to improve char quality and unlock its potential application in the cement, steel, and other industry sectors.
Chemical leaching is a process that allows for the separation of the soluble components of a solid material by dissolving them into a liquid phase. A common application of chemical leaching is low-grade coal cleaning, aiming to reduce the amount of mineral matter [
15], which comprises ash and sulfur [
16]. The application of the acid leaching process for phosphorus extraction from pyrolyzed sludge has been tested by the authors in previous studies with promising results. Acid leaching performed on sludge-derived char enabled a high extraction rate of P, Ca (>90%), and other metals; at the same time, the process extracted only 50% of Al and was ineffective on SiO
2 extraction [
32]. The products obtained by the acid leaching process applied to sludge-derived char consist of an acid liquid containing the recovered inorganic elements, such as P, Ca, and Fe, and a biocoal (LBC), with a higher C content compared to the raw char and a reduced ash concentration. The inorganic elements retained in the liquid can be recovered by chemical precipitation, producing an inorganic compound with a high N, P, and K content. Biocoal, with its reduced ash content and, thus, increased carbon content, can be considered an alternative as a precursor for the production of adsorbents and biomaterials.
To the best of the authors’ knowledge, many studies are available concerning the extraction of phosphorus from sludge ash [
33,
34,
35] or the production of biochar by slow pyrolysis [
36,
37]. However, limited literature was found addressing the simultaneous valorization of both carbon and phosphorus retained in sewage sludge, mainly involving hydrothermal carbonization as a thermochemical treatment process [
38,
39], characterized by high volatility and low market value. The aim of the proposed work is to test the integrated slow pyrolysis and acid leaching processes on an ash-rich industrial sludge to obtain a low-ash biocoal and extract most of the valuable inorganic elements. This integrated process was previously tested by the authors on an ash-rich municipal sewage sludge [
25], but never before on a sludge of industrial origin. Moreover, this study investigated the extraction of inorganics from the leachate by chemical precipitation and the upgrading of the biocoal by means of physical activation to produce two high-value products: an inorganic N-P-K fertilizer in compliance with the EU FPR regulation and an activated carbon with high micro- and meso-porosity, reusable for wastewater treatment or as an alternative to fossil coal in different sectors.
2. Materials and Methods
2.1. Feedstock: Industrial Sludge
The feedstock consists of an industrial sludge generated by anaerobic digestion, centrifuging, and drying of the suspended solids from the WWTP of an Italian meat producer factory. The WWTP producing the sludge receives the waste streams of the agri-food production, the meat manufacturing process, and the wastewater of the whole industrial complex, including equipment washing, toilets, and other facilities.
2.2. Experimental Activity
2.2.1. General Description of the Experimental Activity
The experimental campaign started with slow pyrolysis testing at laboratory scale, performed at different operating temperatures. Then, the leaching of the produced chars was tested, aiming to assess the performance of the leaching process on char produced at different pyrolysis temperatures. Based on the results of the leaching and pyrolysis tests, the most promising operating conditions for the slow pyrolysis test were selected. In particular, the temperature and the residence time chosen for the pilot-scale pyrolysis tests are those that enabled the highest degree of demineralization in the laboratory. Subsequently, the char from the slow pyrolysis pilot scale test was leached, adopting more severe conditions compared to those tested in the laboratory in order to maximize the ash extraction efficiency. The dissolved inorganic compounds were recovered from the leachate by precipitation. Finally, the activation process for the biocoal was tested. The activities performed during the experimental campaign are described in the paragraphs below.
2.2.2. Laboratory-Scale Slow Pyrolysis Tests
Sludge slow pyrolysis tests in laboratory (PT400–PT650) were performed on a thermogravimetric analyzer (LECO TGA701). Six pyrolysis tests were carried out at constant residence time (1 h) and heating rate (20 °C min−1), increasing the temperature by 50 °C, from 400 °C to 650 °C, among the tests. The following nomenclature is adopted to identify the laboratory-scale slow pyrolysis tests:
PT400: test performed at the pyrolysis temperature of 400 °C;
PT450: test performed at the pyrolysis temperature of 450 °C;
PT500: test performed at the pyrolysis temperature of 500 °C;
PT550: test performed at the pyrolysis temperature of 550 °C;
PT600: test performed at the pyrolysis temperature of 600 °C;
PT650: test performed at the pyrolysis temperature of 650 °C.
The mass yield was determined for each trial, and the char from each test was then analyzed in laboratory. The carbon recovery rate was calculated for each test, after char characterization, as follows:
where
C BC (% d.b.) and
mass BC (g) are the carbon content and the mass of the char (BC) from the test, respectively, and
C sludge (% d.b.) and
mass sludge (g) are the carbon content and the mass of the processed dry industrial sludge, respectively.
2.2.3. Pilot-Scale Slow Pyrolysis Test
A pilot-scale slow pyrolysis unit (
Figure 1) was operated for the test in a real-world environment. The pyrolysis unit, called SPYRO, is an auger-type reactor of approximately 2 m in length and an inner diameter of 0.15 m. The unit can convert up to 3 kg h
−1 of feedstock and can be operated up to 600 °C. The pyrolysis reactor is coupled with a condensation unit for the recovery of volatiles in the form of pyrolysis liquid. Details of the pilot unit are reported elsewhere [
32].
The operational parameters of the pilot-scale slow pyrolysis test (PT
P) are shown in
Table 1. The rotating speed of the reactor screw was set at 0.4 rpm to achieve a solid residence time of almost 1 h.
The char produced by the trial was collected and weighted to calculate char mass yield (
PMY) as follows:
where
BCP mass (g) is the mass of char and
sludge mass (g) is the mass of dry sludge.
Then the char was analyzed. The pyrolysis oil obtained by the pilot plant condensation system was first collected and weighted, and then separated gravimetrically, obtaining three phases: a light oil phase (LOP), an aqueous phase (AP), and a heavy oil phase (HOP). After collection, the LOP and the HOP were analyzed.
2.2.4. Experimental Procedures for Chemical Leaching Tests
The char samples obtained from slow pyrolysis tests on the lab scale (BC) and on the pilot scale (BCP) were processed by chemical leaching. The experimental campaign was conducted at laboratory scale to extract and separate the desired inorganic elements from the char. First, chemical leaching tests were performed to identify the optimal leaching operating conditions. The following operating conditions were varied during the experiments:
Mass ratio between leaching solution and processed char (liquid:char);
Pure reagent concentration (HNO3) in the leaching solution (molarity, mol L−1);
Operating temperature (temperature, °C);
Contact time (contact time, h).
All the leaching tests were performed using the same procedure. Char was initially oven dried at 105 °C to a constant weight. The leaching solution was prepared by mixing, in a beaker, demineralized water and nitric acid (dosed as a ≥65% solution by Honeywell, ACS grade). The char was ground manually at a size below 500 µm in the case of BC processing and below 250 µm in the case of BCP, and then added to the solution. Each test was performed using a heating plate equipped with a thermocouple and a magnetic stirrer. At the end of the contact time, the contents of the beaker were vacuum filtered to separate the liquid from the solid phase. The separated solid material was then washed with demineralized water for 30 min to remove the residual acid reagent. The separation of washing water was again performed through vacuum filtration. The solid material was finally oven-dried at 105 °C to a constant weight.
2.2.5. Performance Indicators of Chemical Leaching Tests
Leaching test performances were evaluated by two indicators: degree of demineralization (DD) and element extraction efficiency (EE).
The first, calculated for all leaching tests, expresses the percentage of extracted ash against the initial content in the processed char:
where
ash BC (% d.b.) is the ash in the processed char, and
ash LBC (% d.b.) is the ash in the biocoal.
The second indicator was calculated for the leaching test performed on BC
P (LT
P) only and expresses the percentage of the extracted element against the initial content of that element in the processed char:
where
element LL (mg) is the element mass in the leachate (LL), and
element BC,i (mg) is the element mass in the processed char. Both were calculated given the mass of the LL and processed char, and the element concentration in the materials was determined analytically.
2.2.6. Chemical Leaching Tests on BC from Laboratory-Scale Slow Pyrolysis Tests
To select the process conditions to be performed in the slow pyrolysis pilot plant, the study included an assessment of the impact of slow pyrolysis temperatures on the efficacy of the leaching process in terms of inorganic compound extraction. Six chemical leaching tests at the same conditions were performed, processing the chars (BC) produced in the laboratory at different temperatures. In each test, 11.2 g of char were processed. The tests were performed using nitric acid (HNO3) as a reagent and under the following operating conditions: liquid:char 10:1, molarity 0.5 mol L−1, temperature 30 °C, and contact time 1 h. The following nomenclature is adopted to identify the chemical leaching tests:
LT400: test performed on the BC from PT400;
LT450: test performed on the BC from PT450;
LT500: test performed on the BC from PT500;
LT550: test performed on the BC from PT550;
LT600: test performed on the BC from PT600;
LT650: test performed on the BC from PT650.
The samples of biocoal (LBCs) were collected, and their ash content was determined.
2.2.7. Chemical Leaching Tests on BC from PTP
Finally, a chemical leaching test was performed on the char produced by the pilot-scale pyrolysis test (BC
P), aiming to extract the inorganics from the char and maximize the ash reduction in the final solid. In this test, 20 g of char in a 0.8 M HNO
3 acid solution were processed with a contact time of 2 h and a temperature set at 70 °C. The solid mass yield was calculated for the leaching test performed on BC
P (LT
P), assuming that no organic matter is lost during the leaching process (neither in liquid nor in gas form). The equation adopted is the following:
where:
LMY—Leaching Mass Yield (% d.b.);
Ash LBC—Ash content of biocoal (% d.b.);
Mass BCP—Mass of processed char (g);
Ash BCP—Ash content of processed char (% d.b.).
The outcome of the laboratory trials was used to define the process conditions for the pilot-scale slow pyrolysis test and for the chemical leaching test on the processed char.
2.2.8. Chemical Precipitation Test
The leachate from the LTP was adopted as the starting material for the precipitation test, aiming at recovering the inorganic compounds dissolved in the liquid in solid form by increasing the liquid pH. To this end, known volumes of a KOH solution were dosed in the glass beaker containing the acid liquid. During the process, the system was stirred, and the pH was monitored by a pH meter (Metrohm 827 pH). The 15% KOH solution adopted for the test was prepared by KOH pellets (≥97% purity, ACS grade by Merck). The dosage was stopped when a pH of around 9 was achieved in the liquid phase. The precipitated solid was then separated from the liquid phase by centrifugation and finally dried in an oven at 105 °C.
2.2.9. Activation Tests
The LBC from the LT
P was then activated. The activation tests (AT) were performed using a batch tubular quartz reactor placed in a ceramic furnace. The furnace is equipped with k-type thermocouples both inside and outside the reactor, controlled by a programmable logic controller (PLC). First, a known quantity of LBC
P was charged in the tubular reactor, which was then placed in the furnace. Subsequently, the desired process temperature and the heating rate were set. As a first step, nitrogen was fluxed to avoid oxygen entering and feedstock oxidation. Afterwards, CO
2 was fluxed at a flow rate of 3.5 l min
−1 for the set residence time. During the first activation test (AT1), 9 g of LBC
P were processed, while 8 g of LBC
P were processed during AT2. The activation test operating conditions are reported in
Table 2.
The activated carbon (AC) mass yield obtained by the activation test (
AMY) was calculated for both trials as follows:
where
AC mass (g) is the mass of activated carbon and
LBCP mass (g) is the mass of biocoal.
The total AC mass yield (
ACYtot) obtained by slow pyrolysis (PT
P), acid leaching (LTP), and activation (AT1, AT2) is calculated by the following formula:
2.3. Feedstock and Process Product Characterization
The characterization of the sludge used as pyrolysis feedstock and of the products obtained from the tested processes consisted of different physical-chemical analyses.
The industrial sludge was first air dried to determine its moisture content prior to its characterization. The first analysis consisted of the determination of the content of residual moisture, ash, determined at 550 °C (ash 550) and 710 °C (ash 710), volatiles, and fixed carbon (fixed C, calculated as the difference between 100 and the sum of moisture, volatiles, and ash 550, according to UNI EN 1860-2: 2005), in the sludge (proximate analysis). Then, sludge C, H, N, and S content (ultimate analysis) were determined. In addition, the higher heating value (HHV) was analyzed, calculating the lower heating value (LHV) by means of the HHV, H, and moisture content (following UNI EN ISO 18125: 2018 and UNI EN ISO 16948: 2015). Finally, the concentration of metals P and Si was determined by microwave plasma atomic emission spectroscopy (MP-AES).
Char from slow pyrolysis in the pilot unit was characterized by the same analysis and instruments as the feedstock, while for the chars from laboratory-scale pyrolysis, ash 550 and 710, volatiles, and C, H, N, and S content were determined. The biocoal was characterized by its ash (550 and 710), volatiles (C, H, and N), and its composition in metals (P and Si). Activated carbons were characterized by their ash 550 and 710, C, H, and N contents, surface area, and pore volume. In addition, Fourier-transform infrared spectroscopy (FT-IR) analysis of the activated carbons was performed. The surface area of the pilot unit char, of the biocoal, and of the activated carbons was determined via the Brunauer-Emmett-Teller (BET) method, while the plots of the pore volume and of the surface area of the activated carbons were determined by the density functional theory (DFT) method. The micropore volume of the activated carbons was also derived by the alpha-S method.
The HOP and the LOP of the condensed pyrolysis oil were analyzed, determining their water content, C, H, and N content, HHV, and LHV.
The characterization of the leachate and of the precipitated compound consisted of the analysis of metal, P, and Si concentrations. The precipitated compound composition was also expressed as oxides, starting from the analytical elemental composition and considering the different molar masses.
Moisture was determined by drying the feedstock sample at 105 °C until a constant weight was reached, according to UNI EN ISO 18134-2: 2017. Ash and volatiles were determined by a thermogravimetric analyzer (LECO TGA701), following UNI EN ISO 18122: 2016 and UNI EN 1860-2: 2005 (for ash 550 and ash 710, respectively, by heating around 1 g sample up to 550 °C or 710 °C under constant air flow until constant weight was reached) and UNI EN ISO 18123: 2016 (for volatiles, by heating around 1 g sample under nitrogen flow up to 900 °C).
C, H, N, and S were determined by a CHN-S analyzer (LECO TruSpec CHN-S). C, H, and N analysis was performed according to ASTM D5291-10 for liquid samples (HOP and LOP) and to UNI EN ISO 16948: 2015 for solid materials. ASTM D 4239 was followed for S content determination. For C, H, N, and S analysis, about 60–80 mg of sample material were combusted at high temperatures (950 °C for C, H, and N and 1350 °C for S) and converted by catalysts to carbon dioxide, water vapor, elemental nitrogen, and sulfur dioxide. Lower S content values (<0.1% w/w) were instead determined by an ion chromatography system (Metrohm 883 Basic IC Plus) after combustion by a bomb calorimeter (LECO AC500). A higher heating value (HHV) was obtained by a bomb calorimeter (LECO AC500), according to UNI EN ISO 18125:2018 for solid materials and DIN 51900-1:2000 and DIN 51900-3:2005 for liquid samples (HOP and LOP). The sample (around 0.30–1.0 g) was weighted and placed in the combustion bomb, then flushed with oxygen reaching 30 bar. Before ignition, the temperature was stabilized for 3 min; then the ignition was started by an electrical ignition device, and the temperature was monitored for 5 min. HOP and LOP water content determination was performed by Karl Fischer titration (Metrohm 848 Titrino Plus, Herisau, Switzerland), following ASTM E203-08, titrating about 0.1–0.5 g with iodine-based Karl Fischer reagents. The concentration of metals, P, and Si (inorganic elements) was determined using microwave plasma atomic emission spectroscopy (MP-AES, by Agilent 4200 MP-AES), which uses nitrogen plasma, in compliance with UNI EN ISO 16967: 2015 and UNI EN ISO 16968: 2015. For liquid samples, around 500 mg were analyzed. For the determination of inorganic elements in the organic solids (sludge, chars, and biocoal), the samples (around 30 mg each) have been previously digested with 3 mL of hydrogen peroxide and 5 mL of nitric acid in a Milestone Start D microwave digestion system to be completely solubilized in a liquid sample, then analyzed by MP-AES. Pore volume and surface area were analyzed via a surface area and pore size analyzer (Quantachrome NOVA 2200E), following ASTM D6556-10. This instrument analyzes the pore size of the samples in the range of 2–52 nm.
Before the analysis, samples were dried at 200 °C for 48 h in oven. Then they were degassed in the analyzer at 200 °C for 24 h under vacuum to remove moisture and volatile compounds. Degassed samples were then weighed in a bulb cell and analyzed after immersion in liquid nitrogen to determine the correspondent adsorption isotherm, and then the sample surface area (by BET and DFT methods) and the pore volume (by DFT method) were derived. The cumulated surface area, cumulated pore volume, and alpha-S plot of the activated carbons were generated by the software NovaWinTM by Quantachrome (version 11.02).
FT-IR analysis was performed through a FT-IR Shimadzu IR Tracer 100 in attenuated total reflectance (ATR) mode with an optical resolution of 4 cm−1 and a spectral range from 600 to 4000 cm−1 averaged on 45 scans.
4. Discussion
Carbon, hydrogen, and nitrogen decreased with the increase in slow pyrolysis temperature (
Table 5), as temperature affects the devolatilization of organic compounds. About 6.1% nitrogen was found in the solid produced at 400 °C, while less than 4% was found in that produced at 650 °C (
Table 5). Nitrogen is estimated to be present in the sludge, mainly in the form of protein N (P-N). In fact, according to [
43], despite the fact that pyridine N (N-6), pyrrole N (N-5), and nitrogen oxides have been found in the sludge, P-N and N-6 often represent more than 80% of the total N. On the contrary, sulfur increases (
Table 5), probably due to the high stability of sulfur compounds and the presence of inorganic sulfur [
44].
As shown in
Figure 3, the DD achieved by the chemical leaching tests on the char produced on a laboratory scale slightly increases for the char produced at 450 °C in comparison with that obtained at 400 °C. This can be due to the solubility of the char organic matter produced at 400 °C, which reacts under leaching conditions, reducing the efficacy of leaching on inorganic compounds. At temperatures higher than 450 °C, the DD gradually decreases with increasing pyrolysis temperatures. This result is probably due to the reduction of the solubility of P, Ca, and other heavy metal compounds with the increase in temperature. The reduction of both heavy metals and P-Ca compounds solubility at higher pyrolysis temperatures has been widely investigated in several studies. A higher Ca-P crystallinity was found in the chars produced at high pyrolysis temperatures, correlated with low water-extractable P [
45], and a reduced availability of phosphorus was identified above 600 °C pyrolysis [
46]. A reduced solubility was also found for other heavy metals when the pyrolysis temperature increased from 500 to 800 °C [
47]. The pyrolysis trials performed in the pilot plant showed a lower char mass yield compared to that obtained in the laboratory at the same temperature. However, the composition of the char obtained at the pilot scale was similar to that produced in the laboratory. Therefore, the different mass balance is probably related to the losses or partial combustion of char powder, which could take place during the test in the pilot plant.
Despite the mass yield of the activated carbon compared to the initial processed feedstock resulted from 13.8 to 15.6%, the amount of stored carbon in the produced activated carbons was 28.0% for AC2 and 31.4% for AC1. Thus, it can be estimated that about two-thirds of the carbon retained in the feedstock was devolatilized during the process. Both the obtained activated carbons showed a high surface area, more than 20 times higher than that of the LBC
P. Moreover, carbon concentration increased from 59.4% in LBC
P to 72–73%, while hydrogen was reduced to less than 1 %. These results enabled us to consider both AC1 and AC2 as anthracitic carbons, according to the Van Krevelen coal classification (
Figure 10), and hence comparable to those used in the steelmaking sector.
Typical properties of the activated carbons are high carbon content, high specific surface areas, and a high level of porosity. The distribution of pores, in combination with the specific surface area, is dependent on the starting materials of activation and on the adopted process parameters [
49]. Therefore, specific absorption trials should be performed to better determine the quality of the produced activated carbons. However, the specific surface areas obtained by the tests were comparable to physically activated biomass [
49,
50].
Since activated carbon has a large specific surface area and has developed micropores, it has strong absorptivity and a large adsorption capacity. Since the physical adsorption properties of activated carbons are mainly related to the specific surface area, it is expected that both AC1 and AC2 could be used as adsorbents in wastewater treatment or in industrial emissions processes. Activated carbon can be used either alone or in combination with other water treatment technologies [
51].
The FT-IR spectra of both samples (
Figure 9) show a lack of functionalization typical of graphitic material [
52]. The spectra appear to be smooth in the region 4000–1700 cm
−1, while low signals are detected around 1600 cm
−1, from aromatic C=C stretching, and around 1100 cm
−1, probably related to inorganic oxides left from the feedstock. The spectra were also compared to a commercial steam-activated charcoal, NORIT B SUPRA EUR, which shows a similar smooth spectrum as the thermally activated samples, with even fewer functional groups detected. The FT-IR analysis enabled us to determine the high graphitization rate of the activated carbons (AC1 and AC2) and their similarity with commercial adsorbents available on the market [
53]. The precipitated salt showed a high concentration of P, Ca, and Fe, including a relevant content of K derived from the KOH dosed for the precipitation test. In particular, converting the P and K concentrations resulting from PS analysis into equivalent P
2O
5 and K
2O, the resulting concentrations are 16% P
2O
5 and 23% K
2O. The composition of the obtained precipitated salt was compared to that reported in the current Italian regulation on fertilizers [
54] (and following modifications). The obtained precipitate could be potentially framed as a couple of fertilizing products included by the decree, that is, mixed phosphate salts that shall have a minimum 10% P
2O
5 concentration, or PK fertilizers, named “concimi PK”, whose requirements include minimum P
2O
5 and K
2O concentrations of 5% each and a minimum overall concentration (P
2O
5 + K
2O) of 18% [
55]. The concentration of P, as P
2O
5, and K, as K
2O, resulted largely above the minimum values required by the decree.
Moreover, the concentration of P2O5 was in line with that required by the EU FPR for precipitated phosphate salts (recognized as Component Material Category 12). The high concentration of iron, above the 10% d.b. EU FPR limit for precipitated phosphate salts, could represent a limit for the marketability of the product in the fertilizer industry, despite the fact that iron is not considered a toxic element. However, due to the low precipitation pH of iron ions, a two-stage selective precipitation could be performed to separately collect iron and increase the quality of the inorganic fertilizer.