Energy production potential of phytoremediation plant biomass: Helianthus annuus and Silybum marianum

Industrial Crops & Products 135 (2019) 206–216

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Energy production potential of phytoremediation plant biomass: Helianthus annuus and Silybum marianum

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Selda Yiğit Hunce, Rafael Clemente, Maria Pilar Bernal

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Centro de Edafología y Biología Aplicada del Segura, CSIC. Campus Universitario de Espinardo, Apartado 164, 30100, Murcia, Spain

ARTICLE INFO

ABSTRACT

Keywords: Anaerobic digestion Biogas production Higher heating value Lignocellulosic biomass Phytoremediation Trace elements

The use of the biomass of two plant species (Helianthus annuus and Silybum marianum) generated during the phytoremediation of a trace element contaminated soil was studied for the production of clean and renewable energy in laboratory experiments. Biogas production potential, aerobic biodegradability and calorific values of the plants were determined and compared with the results of the plants grown in non-contaminated soil to evaluate the effect of trace elements contamination on their potential use as feedstock for bioenergy production. The biogas production potential of the aerial vegetative biomass of S. marianum (194 - 223 ml g−1) was found to be higher than that from H. annuus (134 - 154 ml g−1), with no significant differences between non-contaminated and contaminated soils (average CH4 concentration: 63.1% for S. marianum and 65.7% H. annuus, respectively). The greatest biogas production was obtained from the seeds of H. annuus (356 - 473 ml g−1; 70.7% of CH4) after 9 days of anaerobic digestion. Also, the aerial vegetative biomass of H. annuus had higher calorific values than that of S. marianum (17.1 and 14.3 MJ kg−1, respectively). Combustion was the most efficient method for energy production compared to anaerobic digestion, while the presence of low concentrations of trace elements in plant biomass did not limit the potential energy recovery. These findings show the suitability of bioenergy production as an added value alternative for the management of plant biomass coming from phytoremediation processes.

1. Introduction Soil pollution by trace elements (TE), which is mainly caused by industrial and human activities, leads to soil degradation and the deterioration of its physical and chemical properties. Phytoremediation is an environmentally friendly, effective, aesthetically appropriate, energy efficient and cost-effective technology for the restoration of contaminated soils (Sabir et al., 2015). The most appropriate phytoremediation techniques for TE contaminated soils have been found to be phytoextraction, phytoimmobilization and phytostabilization (Clemente et al., 2015). Phytoextraction uses plants to remove the pollutants from the soil, and generally can be only successfully applied in areas where the contamination is not too deep and concentrations are just above thresholds or guidance levels (Dickinson et al., 2009). Phytoimmobilization and phytostabilization techniques seek to reduce TE toxicity through chemical and physical changes in the soils and the rhizosphere that reduce TE solubility, mobility and bioavailability. For phytoimmobilization and phytostabilization, the plant species should be tolerant to metal(loid) toxicity by exclusion (retention of the contaminants in the roots and rhizosphere), with reduced transport from

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roots to shoots. However, the management of the plant biomass obtained during and at the end of the process remains one of the main challenges for the implementation of phytoremediation techniques, as it may contain varying concentrations of heavy metals and metalloids (Gomes, 2012). Drying, incineration, gasification, pyrolysis, acid extraction, composting and anaerobic digestion are the suggested methods for the recycling of phytoremediation plant residues (Gomes, 2012; McKendry, 2002a, b; Sas-Nowosielska et al., 2004). Also, obtaining valuable products from the biomass (such as bioenergy) may ensure the commercial success of the phytoremediation technologies for TE contaminated soils (Meers et al., 2010). Anaerobic digestion is considered one of the most efficient and practical technologies to obtain renewable energy from organic wastes and plant residues (Santarelli et al., 2012). The process consists of the biodegradation of the organic materials in the absence of oxygen with the generation of biogas (Sawatdeenarunat et al., 2015). Anaerobic digestion can also be defined as the direct conversion of organic matter into biogas, which involves mostly methane and carbon dioxide. During anaerobic digestion, complex organic substances such as carbohydrates, proteins and lipids are transformed into simple sugars, amino acids and

Corresponding author. E-mail address: [email protected] (M.P. Bernal).

https://doi.org/10.1016/j.indcrop.2019.04.029 Received 25 January 2019; Received in revised form 22 March 2019; Accepted 14 April 2019 Available online 30 April 2019 0926-6690/ © 2019 Elsevier B.V. All rights reserved.

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Nomenclature A A-C A-NC ANOVA B0 Bm BMP C/N C0 Cm EC HHV

k OM RMS S-C SE S-NC T TBMP TE TN TOC TS VFA VS

Aerial vegetative part Aerial vegetative part from contaminated soil Aerial vegetative part from non-contaminated soil Analysis of variance Biogas production potential Experimental biogas production Biomethane potential Carbon to nitrogen ratio Potentially mineralizable-C CO2-C mineralization Electrical conductivity Higher heating value

fatty acids by hydrolytic enzymes. These simple structured products are converted by acidogenesis into volatile fatty acids, CO2, and H2, which are further converted into methane gas by methanogenesis (Adekunle and Okolie, 2015). Methane-containing biogas produced by the anaerobic digestion of organic materials can provide clean, versatile and renewable energy, since methane can be used both as a substitute for fossil fuels for heat and energy production, and as a vehicle fuel (Balat and Balat, 2009; Weiland, 2010; Lanzini et al., 2017; Papurello and Lanzini, 2018). Characterization of solid organic substrates (moisture content, total and volatile solids and the characteristics of the biodegradable organic compounds) for anaerobic digestion is of great relevance in the control and operation of the process (Raposo et al., 2011). Anaerobic biodegradation of lignocellulosic agricultural wastes is a complex process, in which lignin and cellulose content of the substrate significantly affect the degradation process (Buffiere et al., 2006). In fact, the crystallinity of the cellulose and the lignin content limit the biodegradability of plant materials (Nizami et al., 2009; Triolo et al., 2012). Plants containing high proportion of soluble carbohydrates, lipids and proteins, and low amounts of hardly biodegradable polymers are considered the most appropriate species for the production of biogas (El Bassam, 1998). Then, the biogas production potential of plants may change depending on the different parts of the same plant used, due to their different composition (Gunaseelan, 2009). Numerous agricultural and crop residues are commonly used to produce compost that can be later used to improve soil fertility and for several agricultural uses (Al-Barakah et al., 2013; Sánchez-Rosales et al., 2017). Composting is an spontaneous exothermic process where the temperature increases during the microbial biodegradation of the organic material (Bernal et al., 2009 2017). Recent studies have revealed that the energy produced in composting piles can be recovered as a source of thermal energy for heating (Rodrigues et al., 2018; Smith and Aber, 2018; Walther et al., 2017). The suitability of an organic material for composting can be defined in terms of its aerobic biodegradability, determined by microbial respiration as O2 consumption or CO2 production (Barrena-Gómez et al., 2006). Also, several parameters of the aerobic degradation process have been used as indices of biological degradability or decomposition of organic residues, such as the potentially mineralizable carbon and the mineralizaton rate (Bernal et al., 1998; Nourbakhsh, 2006). The agricultural residues can also be converted to energy by combustion (Mckendry, 2002a, b). The heat energy obtained by combustion of agricultural and crop residues can be used by household warming, industrial processes, crop drying, etc., and it can also be used to generate electricity by producing steam (Broek et al., 1996; Hiloidhari and Baruah, 2011). The cultivation of energy crops may be unfeasible in the long term because of the production costs and also the unfortunate possibility to increase food prices due to the use of agricultural soils for non-food crops (Tilman et al., 2006; Triolo et al., 2012). Owing to this,

Rate constant Organic matter Residual mean square Seeds from contaminated soil Standard error Seeds from non-contaminated soil Temperature Theoretical biomethane potential Trace elements Total nitrogen Total organic carbon Total solids Volatile fatty acid Volatile solids

the production of bioenergy from plant biomass accumulated during soil phytoremediation can help to solve the problem associated with plant disposal and also have a potential benefit associated to the production of renewable energy using marginal non-productive soils. The objective of the study was to determine the potential of recycling two Mediterranean plant species used for phytostabilization of TE contaminated soil through energy production: biogas by anaerobic digestion, thermal energy by combustion and composting. Specifically, Helianthus annuus and Silybum marianum obtained from phytostabilization experiments of TE contaminated soils were studied as source of biomass for the production of biogas by anaerobic digestion, and thermal energy by combustion. Their suitability for composting was also assessed through the study of their aerobic degradation in a respiration test. 2. Materials and methods 2.1. Plant materials The plant species studied in this experiment were sunflower (Helianthus annuus L.) and milk thistle (Silybum marianum (L.) Gaertner). Sunflower (H. annuus) is grown for oil production and also considered a bioenergy crop due to the production of biodiesel by transesterification of the oil from the seeds (Santana et al., 2018). Commonly called milk thistle (S. marianum) is a native plant species grown in the Murcia region of Spain, which has been previously used for the phytostabilization of TE contaminated soils and has proved its potential as an energy crop (Clemente et al., 2019; Domínguez et al., 2017a; Martínez-Fernández et al., 2014). The plant materials were collected from a phytostabilization experiment previously run in a TE contaminated mining soil. The soil was taken from the left margin of Portmán gully (N 37° 35′ 44″, W 0° 51′ 36″), near North-West end of Portmán village within the Sierra Minera of Cartagena-La Unión (Murcia, SE Spain), which was clearly affected by former mining activities carried out in the area (Conesa and Schulin, 2010). The concentration of TE in the soil were: As: 1976, Cd: 12.0, Cu: 230, Pb: 19129, and Zn: 2257, all in mg kg−1. The soil had been treated with a mixture of amendments (solid fraction of pig slurry and paper mill sludge) to provide nutrients and organic matter and to increase the pH. Plants cultivated concomitantly in a non-contaminated agricultural soil were used as controls. Two planter boxes were used for each soil. The plants were harvested before senescence (79–80 BBCH scale) and composite samples were taken from each planter box for each species and divided into aerial vegetative parts (shoots of S. marianum and stems and leaves for H. annuus) and seeds (only ripe seeds were considered). Samples were dried in an oven at 60 °C during 48 h until constant weight and ground to 0.5 mm in a stainless steel laboratory mill (A10 IKA Labortechnik, Staufen, Germany) prior to experiments and analysis. Plant samples were analyzed for volatile solids (VS) by 207

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ashing at 550 °C using a muffle furnace (Carbolite AAF 11/3, Hope Valley, England) according to EPA method 1684 (EPA, 2001). Concentration of total organic carbon (TOC) and nitrogen (N) were measured in an elemental microanalyzer (EuroEA, EuroVector, Milan, Italy). Macro- and micro-nutrients (Ca, Mg, K, P, S, Fe) and TE concentrations (As, Cd, Cu, Pb, Zn) were determined by ICP-OES (ICAP 6500 DUO + ONE FAST, Thermo Scientific, Waltham, USA) after H2O2 - HNO3 (1:4 v/v) microwave (ETHOS1, Milestone, Sorisole, Italy) assisted digestion. Lignin, cellulose and hemicellulose concentrations were measured only in the aerial vegetative part of the plants (as the yield of seeds was limited and not enough for all analyses) using the American National Standard method (ANSI and ASTM, 1977a, b), and soluble carbohydrates were determined by the anthrone method. All chemical analyses were done in duplicate.

measured during the study and converted to volume (ml) of gas produced. The results were expressed as volume of biogas per gram of plant dry weight. The volume of produced biogas was calculated using the “ideal” gas law and the Avogadro’s law, as follows: Biogas volume (mL/g) = [VH × (Ps-Pi) × R × T × 22.4 × 1000] / m where VH is the headspace volume (L), Ps and Pi are the pressures (kPa) obtained from sample and inoculum experiments, R is the gas constant (8.314472 L kPa K−1 mol−1), T is the temperature in Kelvin (K) and m is the weight of sample (g). The experiment was considered finished when the biogas production reached almost the maximum in most samples, established in 9 days for all samples (by comparison). At the end of each experiment, gas samples (duplicate) were taken from each bottle through the septa port using a 10 ml glass syringe, and samples were kept in vacuum containers prior to analysis (Hansen et al., 2004). These gas samples were considered representative of biogas composition as the generated gases accumulated in the bottles throughout the incubation period. The percentage of methane (CH4) of the biogas was analyzed by using a Gas Chromatography system (Agilent 490 Micro GC, Santa Clara, CA, USA). Then, the biomethane potential (BMP) was calculated multiplying the data of the biogas production potential by the CH4 concentration in the biogas, and expressed as volume (ml) of CH4 per gram of VS. The inoculum and the digested samples were analyzed for total solids (TS) by drying at 105 °C, volatile solids (VS) by ashing at 550 °C, pH and electrical conductivity (EC) of samples directly using a pH meter with glass electrode (Crison Basic 20, Crison Instruments, S.A., Barcelona, Spain) and with a conductimeter cell (Crison GLP 31, Crison Instruments, S.A., Barcelona, Spain), respectively. Adjustments of pH were not required before the incubations, as the pH of the mixtures did not differ significantly from the values of the inoculum (data not shown).

2.2. Aerobic degradation experiments The aerobic biodegradability of the plants defines their suitability for composting, which was estimated measuring the CO2-C released during a respiration test. For this, 5 g (dry weight) of plant material were introduced into 500 ml hermetic glass vessels, moisture content was adjusted to 70% by adding distilled water and incubated in the dark at 25 °C for 35 days in an incubator. The CO2 evolved was captured in 10 ml of 0.5 mol L−1 NaOH in vials placed inside the hermetic glass vessels. Empty vessels were used as blanks. Each sample was replicated twice and blank was replicated three times. After 1, 3, 4, 7, 10, 14 days, and weekly until 35 days, the vessels were opened for air exchange and for replacing the vials containing NaOH. The evolved CO2 was measured by titrating the unreacted NaOH with 0.5 mol L−1 HCl, using phenolphthalein indicator after the precipitation of carbonates with a few drops of saturated BaCl2 solution. The results were expressed as the percentage of plants TOC content that had degraded during the incubation ([accumulated CO2-C (g) / TOC (g)] × 100). The seeds of S. marianum from non-contaminated soils could not be tested due to insufficient yield.

2.4. Higher heating values

2.3. Biogas production experiments

The higher heating value (HHV) or higher calorific value of the plant samples was measured by calorimetry. The method consists of the complete combustion of a known quantity of material inside an oxygen bomb calorimeter under controlled conditions. The calorific value is then determined from the temperature measured before, during and after combustion, taking into account the necessary corrections. Briefly, 1 g of dry material was pressed and introduced into the calorimetric bomb (IKA-WERKE, IKA®-Werke GmbH & Co. KG, Staufen, Germany), and then filled with excess oxygen under pressure to provide complete combustion. The bomb was completely immersed in a water filled container (2150 ml), and initial water temperature was kept at 18 ± 0.5 °C. The temperature of the water bath was measured with a temperature probe with 0.001 °C accuracy at 10 s time intervals. The variation of temperature during combustion was used to determine the amount of heat released. The results were obtained in kcal per g dry sample and then expressed in MJ per kg VS.

Biogas production was determined for the aerial vegetative parts and the seeds of both plant species produced in TE contaminated and non-contaminated soils in anaerobic experiments. An anaerobic inoculum was obtained from an urban wastewater treatment plant (reactor capacity 7612 m3 and hydraulic retention time of 29.7 days at mesophilic temperature, 37 °C). The ANKOM gas production system (ANKOMRF, ANKOM Technology, Macedon, NY, USA) was used for measuring the gas produced during anaerobic digestion. The system consists of glass bottles (containing the samples), modules (connected to each bottle as a screw tap) to measure pressure and temperature of each bottle, a reference module recording ambient pressure and a wireless base coordinator to connect to the modules signal by radio frequency. The pressure and temperature in each sample module and also the reference module are monitored and recorded automatically at selected intervals (15 min for this study) by the GPM software integrated in the system. The microbial activity of the inoculum was controlled for 24 h before the start of the experiments by measuring the gas produced under anaerobic conditions. After that, plant samples (0.5 g dry weight) were mixed with 150 ml of the anaerobic inoculum (duplicated samples per soil, plant species and part of the plant) in 305 ml incubation bottles (average substrate: inoculum ratio 1:4.5, referred to VS content), and the head space was flushed with N2 before sealing the bottles and running the system. A control with inoculum and without plant material and a positive control with cellulose (0.25 g in 150 ml inoculum) were also run. Bottles were kept in an incubator at 35 °C during 9 days. The substrate to inoculum ratio was optimized in a previous experiment using different ratios. The gas pressure produced during the anaerobic digestion was

2.5. Computational and statistical analysis The data of anaerobic and aerobic degradation studies were fitted to the following first-order kinetic models by the non-linear least-square technique (Marquardt-Levenberg algorithm) using SigmaPlot v.13.0 computer programme (SPSS Inc.): In the aerobic degradation: Cm = C0 (1-e−kt); In the anaerobic degradation: Bm = B0 (1-e−kt); where Cm and Bm are the production of CO2-C (% of TOC) and biogas (ml g−1) at time t, respectively, C0 and B0 indicate the potentiallymineralizable organic-C under aerobic conditions (% of TOC) and the 208

209

n.s. *** ** n.s. n.s. ** n.s. n.s. ** * n.s. n.s. n.s. n.s. n.s. ** *** n.s. n.s. ** n.s. n.s. n.s. n.s. n.s. n.s. * n.s. * ** n.s. n.s. n.s. ** n.s.

0.9 ± 0.21 175 ± 37 0.1 ± 0.07 2.2 ± 0.07 34 ± 0.92 212 ± 49 41 ± 5.5 61 ± 15 78 ± 5.8 375 ± 181 73 ± 30.3 70 ± 19.7 5.5 ± 0.67 10.8 ± 4.00 19.1 ± 3.04 24.7 ± 4.60

* n.s. * n.s. n.s. n.s. n.s. n.s. n.s. n.s. n.s. n.s. n.s. n.s. *** *** *** *** *** *** *** *** n.s. * n.s. n.s. * n.s. * *** n.s. *** *** n.s. n.s. * *** n.s. ** n.s. n.s. n.s. n.s. ** *** n.s. n.s. *** n.s. n.s. *** * n.s. n.s. * n.s. ** *** ** n.s. ** n.s. n.s.

A-NC A-C S-NC S-C ANOVA Sp Pp S Sp × Pp Sp × S Pp × S Sp × Pp × S

A: aerial vegetative part, S: seed, NC: non-contaminated soil, C: contaminated soil, TOC: total organic carbon, TN: total nitrogen. Sp: Species; P: Part of the plant; S: Soil; *, **: significant at P < 0.05 and 0.01, respectively. n.s.: not significant.

0.01 0.21 0.28 0.99 ± ± ± ± 0.1 0.7 0.2 0.9 0.1 ± 0.07 4.9 ± 0.14 < 0.01 0.6 ± 0.00 1.67 0.78 2.40 0.85 ± ± ± ± 52.9 14.3 15.9 12.8 0.19 0.29 0.85 0.75 ± ± ± ± 3.3 3.2 3.7 3.8 16.6 ± 0.93 18.1 ± 1.73 2.1 ± 0.21 3.1 ± 1.06 0.97 ± 0.08 0.87 ± 0.09 6.7 ± 0.64 6.9 ± 1.63 1.45 0.75 1.25 3.55 ± ± ± ± 18.4 37.3 20.6 20.4 2.2 0.5 1.8 7.4 ± ± ± ±

119 ± 0.3 248 ± 89 40 ± 1.3 69 ± 10 < 0.01 117 ± 95 2.1 ± 1.74 3.6 ± 2.33 40 ± 4.1 161 ± 107 33 ± 2.1 61 ± 2.3 33 ± 5.4 463 ± 442 143 ± 70.8 103 ± 21.8 14.1 ± 2.26 6.7 ± 0.21 17.6 ± 2.26 11.2 ± 0.57 < 0.01 2.8 ± 1.70 < 0.01 < 0.01 < 0.01 6.4 ± 7.21 < 0.01 < 0.01 2.05 4.10 1.27 0.78 ± ± ± ± 7.4 5.6 8.3 5.6 0.28 0.42 0.42 0.28 ± ± ± ± 3.8 1.7 2.7 2.5 18.2 ± 1.72 12.6 ± 1.98 8.8 ± 0.14 5.2 ± 1.36 0.14 0.05 0.07 0.78 ± ± ± ± A-NC A-C S-NC S-C

S. marianum 34 ± 0.1 19 40 ± 2.8 10 46 ± 3.8 24 55 ± 1.5 22 H. annuus 38 ± 0.3 16 41 ± 0.6 10 54 ± 0.9 26 54 ± 1.6 27

± ± ± ±

0.1 2.2 4.3 3.3

18.2 41.7 19.9 25.8

± ± ± ±

0.01 1.40 3.75 2.25

1.7 1.2 4.5 4.4

Fe (mg kg−1) Cu (mg kg−1) Cd (mg kg−1) As (mg kg−1) K (g kg−1) Mg (g kg−1) Ca (g kg−1) P (g kg−1) C/N

The concentrations of TE in the plants grown in the contaminated soil were higher than those in the plants from non-contaminated soils (Table 1), as it could be expected, but still below the maximum concentrations considered excessive or toxic for plants (As: 20; Cd: 30; Cu: 100; Pb: 300; Zn: 400 mg kg−1; Kabata-Pendias, 2011). Carbon concentrations (TOC) were slightly higher in plants from the contaminated soil than in those from non-contaminated one, while the opposite was true for N concentration (Table 1). This provides differing C/N ratios to the plants from the two soils and points out a differential nutritional status of the plants grown in each soil, but similar for both plant species (Table 1). Conversely, lignin content of S. marianum aerial vegetative part (12.5 ± 0.64% and 13.8 ± 2.69% in plants from non-contaminated and contaminated soil, respectively) was lower than that of H. annuus (20.9 ± 0.14% and 25.3 ± 0.42% from non-contaminated and contaminated soil, respectively), but without significant differences between soils. The concentrations of cellulose and hemicellulose in the aerial vegetative parts averaged 31.3 ± 0.21 and 19.9 ± 0.16% for H. annuus and 25.2 ± 0.10, and 10.3 ± 0.09% for S. marianum, respectively, and showed no significant differences between soils. Concerning water soluble carbohydrates, S. marianum showed higher concentration (6.8 ± 0.29%) than H. annuus (4.2 ± 0.46 g kg−1). These different characteristics of the plants may affect their aerobic and anaerobic degradation, as well as their heat capacity. According to the results of the aerobic degradation (respiration test), the mineralization process was very intense during the first 14 days for both plant species and plant parts (Fig. 1), while the process was much slower during the rest of the incubation period. However, the aerial vegetative part of S. marianum from the contaminated soil showed a 4-day delay in the start of the microbial degradation, which could delay the temperature increase in the initial phase of the composting process. Net C mineralization (Cm, % TOC) after 35 days of aerobic incubation ranged from 13.2–21.3 % for S. marianum and was between 11.4 and 25.7% for H. annuus (Table 2), with no significant differences between the different plant growing conditions

TN (g kg−1)

3.1. Characteristics and aerobic biodegradability of the plants

TOC (%)

3. Results and discussion

Sample

Table 1 Characteristics of the plant samples used in the different experiments. Average values ( ± SE) of two plant samples grown in the same soil (two different planter boxes).

Mn (mg kg−1)

Pb (mg kg−1)

Zn (mg kg−1)

ultimate biogas production potential under anaerobic conditions (ml g−1), respectively, and k is the degradation rate constant. For the statistical significance of the curve-fitting, residual mean square (RMS) and the F value of the ANOVA of the non-linear curve fitting were also calculated for each plant sample. The effect of the plant species, part of the plant and soil on plant composition and aerobic degradation was determined by 3-way ANOVA, and differences between means were determined using Tukey’s test at P < 0.05. Before the statistical analysis, data were tested for normality using the Kolmogorov–Smirnov test. Pearson’s correlation coefficients were determined between degradation parameters (aerobic and anaerobic) and TE concentrations in the plants (IBM SPSS Statistics 22.0 software). The theoretical BMP (TBMP) of the aerial vegetative parts was calculated using a stoichiometric equation (Triolo et al., 2012): TBMP = ([lipid] × 1014 + [protein] × 496 + [carbohydrate] × 415 + [lignin] × 727) × 0.001. Plant composition data (expressed in g kg−1 VS): determined lignin, cellulose, hemicellulose, soluble carbohydrates and proteins (calculated multiplying N concentration by 6.25) concentrations, and the estimated lipid concentration of the aerial vegetative parts (3.09 g kg−1 dry matter; Triolo et al., 2012) were used in the equation. The ratio BMP/TBMP was used as the anaerobic biodegradability. Principal component analysis was run with all the parameters to show the relationships between the aerobic degradation, the energy production potential (BMP and HHV) and the plant characteristics (TOC, TN, macro- and micro-nutrients, and TE concentrations).

77 ± 1.0 314 ± 92 76 ± 7.2 103 ± 52

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Fig. 1. Respiration test for aerobic degradation of S. marianum (a) aerial vegetative parts and (b) seeds, and H. annuus (c) aerial vegetative part and (d) seeds. A-C: Aerial vegetative part from contaminated soil, A-NC: Aerial vegetative part from non-contaminated soil, S-C: Seeds from contaminated soil, S-NC: Seeds from noncontaminated soil. The symbols represent the mean values and standard error for each species, part of the plant and soil replicate. The lines represent the predicted degradation by first-order kinetic model for each sample. The numbers in the legend indicate the soil/planter box replicate.

(contaminated or non-contaminated soils). The results of the two tested samples of H. annuus from the two different planter boxes with contaminated soil were very variable (11.4–20.3 %), which could be associated to the different leaves/stem proportions of the plant samples from each replicated planter box (0.30 and 0.52 dry matter), and the consequent different concentration of TE in each replicate sample (large variability (SE) in Fe and Zn concentrations; Table 1). The leaves/stem ratio (0.35 ± 0.04) and TE concentrations (Table 1) were more homogeneous in the samples from the non-contaminated soil replicates. The presence of certain TE in the plant biomass, even at low concentrations, may affect the aerobic and/or anaerobic degradation processes as the microbial activity could be partly inhibited. For instance, the richness and diversity of the bacterial community during the composting of agricultural wastes has been shown to be reduced under Pb stress (Huang et al., 2017). However, the microorganisms responsible for the aerobic transformation of waste materials during composting have shown tolerance to TE. Pseudomonas and Ochrobactrum isolated from sewage sludge tolerated up to 1250 mg L−1 of Pb2+ and more than 2900 mg L−1 of Zn2+ (Heck et al., 2015), and exposure to very high concentrations of Cu (3000 mg kg−1) were required to cause partial inhibition of the activity of the urease and hydrogenase enzymes (by 59 and 65%, respectively) during pig manure composting (Li et al., 2015). Therefore, it is unlikely that the concentration of TE in the plants from contaminated soil could limit the aerobic biodegradability. Net C mineralization in the present study reached up to 25.7% of TOC for H. annuus grown in a non-contaminated soil. These results were larger than those found by Clemente et al. (2014) for pruning wastes of

different shrubs/trees from public gardens (Cm 1–5 % of TOC after 56 days), which can be due to their high lignin content (> 35%). Lignin is the fraction most resistant to microbial degradation (Bernal et al., 2009; Nizami et al., 2009), and its abundance in the plants may be indicative of a reduced biodegradability of the plant material for both aerobic and anaerobic degradation processes. The experimental results fitted a first order kinetic model at highly significant level (P < 0.001; Table 2). There were no statistically significant differences between plants nor between the aerial vegetative part and seeds in both species concerning potentially mineralizable-C (C0), with larger variability in the results of plants from contaminated soils than in those from non-contaminated soils. S. marianum, despite having lower concentrations of lignin than H. annuus, showed similar biodegradability (measured by C0; Table 2), but the process was slower as shown by the lower k values for S. marianum than for H. annuus. Plant composition in terms of cellulose and hemicellulose fractions and cellulose crystallization, have been found to affect biodegradability (Pauly and Keegstra, 2008; Park et al., 2010; Triolo et al., 2012). During aerobic degradation (composting), lignin usually shows scarce biodegradation, with accumulation in the final compost. However, the degradation of cellulose and hemicellulose can induce the formation of soluble OM fractions, due to their hydrolysis into simple compounds (Sánchez-Monedero et al., 1999; Doublet et al., 2011; Bernal et al., 2017), and other components of the organic material can control the degradability under aerobic conditions. Bernal et al. (2019, unpublished results) suggested that the organic compounds forming the silymarin complex (a mixture of polyphenolic flavonoids) in S. 210

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Table 2 Results of the aerobic degradation (Cm) of H. annuus and S. marianum in both the aerial vegetative part (A) and seeds (S) from contaminated soil (C) and noncontaminated soil (NC), and first-order kinetic parameters (potentially mineralizable C (C0) and rate constant (k), including the significance of each parameter in the curve fitting). Samples S. marianum A-NC1 A-NC2 A-C1 A-C2 S-C1 S-C2 H. annuus A-NC1 A-NC2 A-C1 A-C2 S-NC1 S-NC2 S-C1 S-C2 ANOVA Species Part of the plant Soil Sp × Pp Sp × S Pp × S Sp × Pp × S

Cm (%)

k (d−1)

C0 (%)

RMS

F

21.3 21.3 13.2 16.8 17.2 19.7

± ± ± ± ± ±

1.43 0.25 1.00 1.22 0.05 1.15

27.3 28.3 18.9 26.3 21.5 23.5

± ± ± ± ± ±

1.69*** 1.92*** 3.37*** 7.29** 1.66*** 1.22***

0.044 0.043 0.037 0.032 0.051 0.058

± ± ± ± ± ±

0.0048*** 0.0049*** 0.0101** 0.0101* 0.0072*** 0.0058***

0.269 0.317 0.534 1.614 0.431 0.350

1461*** 1434*** 343*** 201*** 723*** 1131***

23.5 25.7 20.3 11.4 22.4 20.1 25.3 22.3

± ± ± ± ± ± ± ±

1.90 0.10 0.67 0.00 0.02 0.23 0.15 0.20

24.8 29.7 22.6 12.3 24.1 20.7 31.3 22.9

± ± ± ± ± ± ± ±

1.11*** 1.60*** 1.22*** 0.43*** 1.40*** 1.13*** 5.18*** 1.05***

0.090 0.062 0.066 0.074 0.070 0.083 0.052 0.089

± ± ± ± ± ± ± ±

0.0098*** 0.0069*** 0.0075*** 0.0057*** 0.0087*** 0.0107*** 0.0160* 0.0101***

1.127 0.890 0.619 0.103 0.940 0.979 5.226 0.998

528*** 820*** 710*** 1286*** 498*** 390*** 154*** 492***

n.s. n.s. n.s. n.s. n.s. * n.s.

n.s. n.s. n.s. n.s. n.s. n.s. n.s.

* n.s. n.s. n.s. n.s. n.s. n.s.

Cm: CO2-C mineralization; C0: potentially mineralizable-C; k: rate constant; RMS, residual mean square; F, factor of the ANOVA for the curve fitting; Numbers (1, 2): planter box per soil. Seeds from S. marianum in non-contaminated soils were not studied due to a low yield. *, **, ***: significant at P < 0.05, 0.01 and 0.001, respectively. n.s.: not significant.

significant differences in lignin concentrations were observed for the aerial vegetative parts of the plants, or by their differing TE concentrations (high SE for Pb and Zn; Table 1). Also, the plant sample having low biogas production (C2) showed low degradability under aerobic conditions (Fig. 1). In the case of the aerobic degradation, the results of the plant sample from the contaminated soil with the high mineralization data (C1) were similar to the results obtained in the plant samples from the non-contaminated soil (Fig. 1c), while in the anaerobic digestion the plant from the contaminated soil with the lower values of Bm (C2) was close to the samples from the non-contaminated soil (Fig. 2c). This can indicate that the concentrations of TE in the plants affect differently the microorganisms responsible for the aerobic and the anaerobic degradation (Sandrin and

marianum could be responsible for the delayed and slower aerobic degradation of the plant tissue. Such behavior can result in poor performance of the composting process with the aerial biomass of S. marianum. 3.2. Biogas and thermal energy production potential According to the characteristics of the inoculum and its mixtures with the plants, the pH values shown at the end of the anaerobic digestion (Table 3) generally remained in the optimum range for anaerobic digestion (6.8 and 7.2) and between the pH tolerance of the process (6.5–8.0; Cioabla et al., 2012). This indicated that the anaerobic digestion took place without significant accumulation of volatile fatty acids (VFA; Chen et al., 2015; Franke-Whittle et al., 2014). The TS content of the mixtures at the beginning of the experiment ranged between 2.0% and 3.2% (data not shown) and a large proportion of it consisted of VS. At the end of the experiment, the VS concentration of the samples with the aerial vegetative part of S. marianum showed similar values to those of the inoculum control samples at the beginning of the experiment (Table 3), and lower values were observed for H. annuus. The degradation of the VS of the plants ranged from 50.0% to 61.1% for the aerial vegetative parts, and reached up to 73% for the seeds of both plants after the 9 days of anaerobic digestion (Table 3). In relation to the anaerobic digestion results (Fig. 2), biogas production was significantly affected by the plant species and part of the plant (Table 4). S. marianum plants (aerial vegetative parts and seeds) grown in contaminated soil gave greater results of biogas production than plants grown in non-contaminated soil (Table 4). The seeds of H. annuus showed the highest biogas production (Table 4), and, as occurred for the aerobic biodegradability test, high variability in the results of H. annuus replicates was observed in the aerial vegetative part samples from the contaminated soil (Fig. 2, Table 4); those results were confirmed in two different runs (average data of the two runs are shown in Fig. 2). These differences could be explained by the leaves and stem proportion of plant sample (as indicated previously), although no

Table 3 Characteristics of the inoculum at the beginning and at the end of the anaerobic digestion test, and the mixtures with the plant materials at the end of the experiment, and the calculated VS removal from plant material (A: aerial vegetative part; S: seeds from both contaminated (C) and non-contaminated (NC) soils). Average values ( ± SE) of two plant samples grown in the same soil (two planter boxes per soil). Sample

pH

TS (%)

VS (%)

Inoculum-Initiala Inoculum-Finala S. marianum A-NC A-C S-NC S-C H. annuus A-NC A-C S-NC S-C

7.91 ± 0.44 7.44 ± 0.35

2.91 ± 0.44 2.65 ± 0.45

1.83 ± 0.25 1.73 ± 0.36

6.79 6.71 7.41 7.41

± ± ± ±

0.05 0.01 0.04 0.11

2.93 2.88 3.11 3.01

± ± ± ±

0.01 0.06 0.02 0.01

1.81 1.80 2.16 2.09

± ± ± ±

0.02 0.01 0.02 0.01

51.4 61.1 51.6 73.2

± ± ± ±

8.96 2.61 6.62 2.23

7.25 7.24 7.33 7.44

± ± ± ±

0.08 0.06 0.04 0.03

2.20 2.19 3.08 2.97

± ± ± ±

0.01 0.01 0.03 0.02

1.39 1.40 2.13 2.04

± ± ± ±

0.01 0.00 0.02 0.01

51.7 50.0 58.9 66.2

± ± ± ±

2.36 5.77 5.31 5.65

a

211

n = 4.

VS plant removal (%)

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Fig. 2. Biogas production from S. marianum aerial vegetative part (a) and seeds (b), H. annuus aerial vegetative part (c) and seeds (d) for both contaminated (C) and non-contaminated soils (NC). Symbols are the experimental data (single replicate) and lines represent the predicted values by the first-order kinetic model for each sample. The numbers in the legend indicate the soil/planter box replicate.

Maier, 2003). Some TE (Cr, Co, Cu, Mn, Mo, Ni, Se, W, Zn) are essential to the activity of enzymes involved in the anaerobic digestion (Drosg, 2013; Raposo et al., 2011; Thanh et al., 2016). However, they become toxic or inhibitory above certain concentrations (Cd 20, Cr 100, Cu 40, Ni 10, Pb 340 and Zn 150 mg L−1; Drosg, 2013), their degree of toxicity depending on their chemical forms in solution (e.g., as free ions or as carbonates; Deublein and Steinhauser, 2008). Iron is essential for the growth of almost all microorganisms, as are Mn and Mo, and can have some stimulatory effects, although these are rarely observed (Mudhoo and Kumar, 2013). Zinc was considered one of the most toxic heavy metals to VFA producing organisms (acidogens) in the anaerobic digestion process (Lin, 1993; Lin and Chen, 1999). The inhibition effect observed in one of the contaminated samples of H. annuus (C2) may possibly be related with its relatively high concentrations of Zn (448 mg kg−1), which provided 1.5 mg L−1 in the mixture with the inoculum for the anaerobic digestion. Zhang et al. (2003) reported that Zn enhanced methane fermentation productivity up to a limit of 1.13 mg L−1, while according to Bozym et al. (2015) and Matheri et al. (2016) Zn inhibits methane formation at a concentration of 1 mg L−1. Then, in the case of the enhanced biogas production of the aerial vegetative part of the replicate 1 of H. annuus from the contaminated soil (C1; Fig. 2) may be due to the Zn concentration (293 mg kg−1), which provided about 1 mg L−1 to the mixture for anaerobic digestion (lower than the C2, previously commented) sufficiently high to have a stimulatory effect on microbial activity, but not large enough to cause metal toxicity. The results of biogas production from the aerial vegetative part of H. annuus and S. marianum and the seeds of the latter fitted to a first-order

kinetic model (Table 4). As occurred for the experimental results of biogas production (Bm), B0 results were affected by the part of the plant, the soil type and at a lower level the plant species, with the interaction part of the plant and soil being statistically significant (Pp × S; Table 4). The TOC concentrations in both species (aerial vegetative part and seeds) were very similar (Table 1); however S. marianum resulted in larger biogas production. This can be explained considering the concentration of lignin in the plants, which was considerably lower in S. marianum aerial vegetative parts than in H. annuus and therefore easier to biodegrade. Lignin is known to affect the biogas production in anaerobic digestion of lignocellulosic substrates as agricultural crops, because of its low biodegradability and/or its complex structure (Buffiere et al., 2006). The seeds of both plant species had greater biogas production (Bm and B0) than the aerial vegetative parts (in H. annuus only seeds obtained from one replicate planter box of each soil was assessed due to the limited seed yield in the other replicate box; Fig. 2d, Table 4), with higher biogas production (Bm) for H. annuus than for S. marianum seeds. The methane production potential of fats and protein is greater than that of carbohydrates (Jingura and Kamusoko, 2017; Victorin, 2016), then the higher content of fats in the seeds may be the main reason for the higher biogas yield. The results of H. annuus seeds fitted the firstorder kinetic model (Fig. 2d), but the high values of RMS (Table 4) indicate large variations between the experimental and calculated data. Then, the results for B0 and k should be taken with care. The BMP of S. marianum averaged 0.174 ± 0.001 L CH4 g−1 VS for the aerial vegetative parts, and ranged 0.144 - 0.284 L CH4 g-1 VS for the seeds, showing lower values in non-contaminated than in 212

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Table 4 The experimental biogas production (Bm) and first-order kinetic parameters (biogas production potential (B0) and rate constant (k), including the significance of each parameter in the curve fitting) of S. marianum and H. annuus anaerobic digestion tests. Two plant samples grown in the same soil (two planter boxes). Samples S. marianum A-NC1 A-NC2 A-C1 A-C2 S-NC1 S-NC2 S-C1 S-C2 H. annuus A-NC1 A-NC2 A-C1 A-C2 S-NC2a S-C1a ANOVA Specie Part of the plant Soil Sp × Pp Sp × S Pp × S Sp × Pp × S

Bm (ml g−1)

B0 (ml g−1)

145 158 204 186 192 209 344 312

± ± ± ± ± ± ± ±

0.41 0.41 0.35 0.41 0.31 0.32 0.32 0.49

183 206 229 216 194 212 406 357

± ± ± ± ± ± ± ±

0.95*** 0.88*** 0.68*** 1.21*** 0.27*** 0.34*** 5.51*** 4.95***

0.0224 0.0252 0.0233 0.0199 0.8359 0.6847 0.2492 0.2409

± ± ± ± ± ± ± ±

117 130 186 122 473 356

± ± ± ± ± ±

0.35 0.36 1.28 0.38 0.33 0.32

128 140 189 133 618 397

± ± ± ± ± ±

1.20*** 1.03*** 0.23*** 1.29*** 9.65*** 3.62***

0.0164 0.0183 0.0275 0.0143 0.1927 0.4926

± ± ± ± ± ±

* *** n.s. *** * n.s. *

k (h−1)

** *** n.s. *** *** n.s. ***

RMS

F

0.0004*** 0.0004*** 0.0003*** 0.0004*** 0.0057*** 0.0047*** 0.0072*** 0.0070***

55.95 58.06 30.43 72.19 7.81 9.73 248.1 179.4

9162*** 10372*** 23013*** 9866*** 51900*** 53643*** 9805*** 10310***

0.0004*** 0.0004*** 0.0001*** 0.0004*** 0.0055*** 0.0154***

43.31 41.74 4.22 35.09 324.8 655.8

6377*** 7490*** 112207*** 7815*** 14828*** 3803***

* *** n.s. * n.s. n.s. ***

A: aerial vegetative part; S: seeds; C: Contaminated soil; NC: Non-contaminated soil. RMS, residual mean square; F, factor of the ANOVA for the curve fitting; Numbers (1, 2): planter box per soil. *, ** and ***: significant at P < 0.05, 0.01 and 0.001, respectively. n.s.: not significant. a Only seeds from one planter box tested for H. annuus due to low yield. Fig. 3. Energy production: a) Biomethane production potentials (BMP) of S. marianum (SM) and H. annuus (HA) aerial vegetative parts (A) and seeds (S) from non-contaminated (NC) and contaminated (C) soils; b) Comparison of energy production methods by anaerobic digestion (methane) and by combustion (HHV) of the aerial vegetative parts of S. marianum and H. annuus from non-contaminated and contaminated soils. BMP for seeds of H. annuus was calculated from the experimental values, as the curve did not fit the first-order kinetic model.

and 0.241 L CH4 g−1VS, respectively). The theoretical BMP (TBMP) of the plants (calculated only for the aerial vegetative parts biomass as not enough yield was obtained for the seeds) was found to be 0.456 L CH4 g−1 VS for S. marianum and 0.542 L CH4 g−1 VS for H. annuus, giving an average anaerobic biodegradability of 31.6 ± 1.88% for S. marianum and of 22.3 ± 1.80% for H. annuus. These results are lower than previous results for wood and hedge cuttings and wild plants (32.7–44.9 %) as well as crops (Triolo et al., 2012). Considering the low biodegradability of the lignin fraction of the plant biomass, the TBMP of the degradable (not containing lignin) fractions were calculated. Based on the biodegradable TBMP, the anaerobic biodegradability was re-calculated obtaining 42.6 ± 2.54% and 31.5 ± 2.76% for S. marianum and H. annuus, respectively. As occurred for wild species (Triolo et al., 2012), the BMP of both plant species was lower than the TBMP (without lignin fraction), which could be due to the crystallinity of the cellulose (which is also low degradable), the proportion of lignocellulose, the composition of the hemicellulose fraction and the lignin-hemicellulose linkages (Pauly and Keegstra, 2008). For S. marianum, the low degradability can be also

contaminated soil (Fig. 3). For H. annuus, BMP was 0.119 ± 0.012 L CH4 g-1 VS for the aerial vegetative parts, while the seeds had the highest methane yield ranging 0.201 - 0.260 L CH4 g-1 VS for both contaminated and non-contaminated soil. Raposo et al. (2008) determined the BMP from solid waste derived from the H. annuus oil extraction (de-oiled cake), ranging 0.107 - 0.227 L CH4 g-1 VS, which is similar to the BMP values obtained from the aerial vegetative parts of the plants in the present study. Other researchers have previously observed differences between the BMP for the seeds and aerial vegetative parts of the same plant. Bauer et al. (2010) determined the BMP of whole plant silage of H. annuus, wheat straw and wheat seeds obtaining 0.345, 0.276 and 0.419 L CH4 g−1 VS, respectively, which are close to the results obtained for seeds of H. annuus. Hesami et al. (2015) obtained a BMP of 0.12 L CH4 g−1 VS from untreated H. annuus stalks after 45 days anaerobic digestion, consistent with the data obtained in the present study for the aerial vegetative parts of H. annuus. Kalamaras and Kotsopoulos (2014) found that BMP of the co-digestion of cattle manure was lower with S. marianum - 0.192 L CH4 g−1 VS, close to our results - than with maize, cardoon and sorghum silage (0.267, 0.308, 213

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In addition, the energy produced from the anaerobic digestion was compared with that recovered by combustion. For this, the results of BMP of the plants were multiplied by the calorific value of methane (39.8 MJ m−3) (Fig. 3b). The results reveal that the energy that can be obtained through biogas production for S. marianum averaged 6.91 ± 0.021 MJ kg-1 VS and 4.70 ± 0.403 MJ kg-1 VS for H. annuus, for plants grown in both contaminated and non-contaminated soil. From these results, it is clear that the energy production by combustion of the aerial vegetative parts of those plants can be more efficient than by anaerobic digestion. In order to determine the main factors affecting the degradability and the production of energy by anaerobic digestion or combustion, the physicochemical properties of the plants and the energy production potentials by anaerobic digestion and by combustion were evaluated together using PCA (Fig. 4). The PCA of the data resulted in 4 principal components, with the first two components constituting 74.8% of total variance. In PC1 (44.8% of the variance), As, Cd, Fe, Mn, Pb and Zn concentrations were on the opposite side of P concentrations, showing a negative effect of TE on plant nutritional status. In PC2 (30.0% of the variance), BMP was on the opposite side of lignin, HHV, TN, Ca, Mg and K, while the HHV of plants was on the same side as lignin concentrations. Then, the lignin content of the plants may condition the method for energy production, while the concentration of TE in the plants may not significantly affect the energy produced but in the plant nutrition.

Fig. 4. Principal component analysis (PCA) of the data of bioenergy production (BMP and HHV) and chemical characteristics of the plants (aerial vegetative part).

4. Conclusions

associated to the presence of polyphenolic flavonoids forming the silymarin complex (Vaknin et al., 2008; Lucini et al., 2016), as suggested for the aerobic degradation (Bernal et al., unpublished). The thermal energy potential by combustion of the plants was determined through the calculation of the higher heating values (HHV). The experiments were only conducted with the aerial vegetative part samples of both plant species, as they are the feasible biomass for combustion. H. annuus had the highest energy values of HHV 17.1 ± 0.1 MJ kg−1 TS (21.3 ± 0.1 MJ kg−1 VS) without differences between plants from contaminated and non-contaminated soils (Fig. 3). Similar results were found by Mckendry (2002a) for different biomasses (17–21 MJ kg−1 TS) and by Boundy et al. (2011) for herbaceous biomass (17.2 MJ kg−1 TS). However, the combustion of S. marianum gave lower values of HHV: 12.3 ± 0.1 MJ kg−1 TS (15.3 and 17.1 MJ kg−1 VS for contaminated and non-contaminated soil, respectively). Besides, Domínguez et al. (2017a) determined the energy production of S. marianum biomass by thermo-chemical conversion (pyrolysis), obtaining a HHV of 16.3 MJ kg−1, and reported a gross calorific value (calorimetry) of 16.2 ± 0.7 MJ kg−1 (Domínguez et al., 2017b), within the range found in the present experiment. Domínguez et al. (2017b) suggested that S. marianum can be a potential bioenergy crop in Mediterranean soils due to its ability to colonize highly contaminated soils. In agreement with the results found in the present experiment, the calorific value of the biomass was not affected by the presence of moderate TE concentrations. The HHV of S. marianum was lower than the values of herbaceous biomass (17.2 MJ kg−1 TS; Boundy et al., 2011), but the HHV of H. annuus was found to be quite similar. As expected, the HHV of the plants was lower than the values of coal (22.7 MJ kg−1 on a wet basis; Boundy et al., 2011), but close to the HHV of forest residues and farmed trees (15.4 and 19.5 MJ kg−1, respectively; Boundy et al., 2011), especially for H. annuus. According to Saidur et al. (2011), the lignin content of the lignocellulosic biomass is generally strongly correlated with the heating value. They found increased HHV (18.0, 18.5 and 19.5 MJ kg−1) with increases in the lignin contents of the following respective biomasses: tobacco leaf (15.0% lignin), wheat straw (20.9% lignin) and spruce wood (31.6% lignin). The data found for S. marianum in the present work (with low HHV and lignin content) were consistent with the afore-mentioned case.

The aerobic biodegradability revealed that in both plant species aerial vegetative parts and seeds showed similar behavior. But excessively high concentrations of TE in the plants can affect their degradability and therefore their composting, compromising the final quality of the compost. The seeds from both plant species produced more biogas than the aerial vegetative part of the plants. The presence of TE in the plants grown in contaminated soil did not affect negatively the production of energy by anaerobic digestion or combustion, lignin being the main factor controlling the energy production. Therefore, the plant biomass generated during soil phytoremediation can be successfully used as energy crops for biogas production and thermal energy, promoting the economic sustainability of the phytostabilization technology. For S. marianum and H. annuus, the combustion of the aerial vegetative parts (shoots and leaves) can be more efficient for energy recovery than biogas production by anaerobic digestion, the latter being more convenient for the seeds, especially from H. annuus. Although the highest energy gain was achieved by combustion, the selection of the energy production method will depend on the local needs. Acknowledgements The authors wish to thank Dr. Claudia Montiel from the Department of Chemical Engineering, Faculty of Chemistry, University of Murcia for her help in the analysis of HHV. The Scientific and Technological Research Council of Turkey (TUBITAK) is acknowledged for supporting S. Yiğit Hunce's postdoctoral study under the TUBITAK-BIDEB 2219International Postdoctoral Research Scholarship Programme. This work was financed by the Spanish Ministry of Economy and Competitiveness and EU FEDER Funds through the project CTM2013-48697-C2-1-R. References Adekunle, K.F., Okolie, J.A., 2015. A review of biochemical process of anaerobic digestion. Adv. Biosci. Biotechnol. 6, 205–212. https://doi.org/10.4236/abb.2015.63020. Al-Barakah, F.N., Radwan, S.M., Abdel-Aziz, R.A., 2013. Using biotechnology in recycling agricultural waste for sustainable agriculture and environmental protection. Int. J. Curr. Microbiol. Appl. Sci. 2, 446–459. American National Standards Institute (ANSI), American Society for Testing and Materials (ASTM), 1977a. Standard Test Methods for Lignin in Wood D 1106-56. American

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