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Effects of sewage sludge biochar on plant metal availability after application to a Mediterranean soil
Chemosphere 89 (2012) 1354–1359
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Effects of sewage sludge biochar on plant metal availability after application to a Mediterranean soil A. Méndez a, A. Gómez b, J. Paz-Ferreiro b, G. Gascó b,⇑ a b
Departamento de Ingeniería de Materiales, E.T.S.I. Minas, Universidad Politécnica de Madrid, C/Ríos Rosas No. 21, 28003 Madrid, Spain Departamento de Edafología, E.T.S.I. Agrónomos, Universidad Politécnica de Madrid, Ciudad Universitaria, 28004 Madrid, Spain
h i g h l i g h t s " Biochar can reduce the leaching of heavy metals present in raw sewage sludge, in particular nickel and zinc. " Biochar amended samples also reduced plant heavy metal availability when compared to sewage sludge amended samples. " Biochar and sewage sludge increased soil respiration with respect to the control soil. " The increment of soil respiration was lower in the case of biochar than when sewage sludge was added.
a r t i c l e
i n f o
Article history: Received 28 February 2012 Received in revised form 9 May 2012 Accepted 23 May 2012 Available online 23 June 2012 Keywords: Biochar Heavy metals Sewage sludge Soil
a b s t r a c t Pyrolytic conversion of sewage sludge into biochar could be a sustainable management option for Mediterranean agricultural soils. The aim of this work is to evaluate the effects of biochar from sewage sludge pyrolysis on soil properties; heavy metals solubility and bioavailability in a Mediterranean agricultural soil and compared with those of raw sewage sludge. Biochar (B) was prepared by pyrolysis of selected sewage sludge (SL) at 500 °C. The pyrolysis process decreased the plant-available of Cu, Ni, Zn and Pb, the mobile forms of Cu, Ni, Zn, Cd and Pb and also the risk of leaching of Cu, Ni, Zn and Cd. A selected Mediterranean soil was amended with SL and B at two different rates in mass: 4% and 8%. The incubation experiment (200 d) was conducted in order to study carbon mineralization and trace metal solubility and bioavailability of these treatments. Both types of amendments increased soil respiration with respect to the control soil. The increase was lower in the case of B than when SL was directly added. Metals mobility was studied in soil after the incubation and it can be established that the risk of leaching of Cu, Ni and Zn were lower in the soil treated with biochar that in sewage sludge treatment. Biochar amended samples also reduced plant availability of Ni, Zn, Cd and Pb when compared to sewage sludge amended samples. Ó 2012 Elsevier Ltd. All rights reserved.
1. Introduction Industrial development and economic growth imply an increase in waste generation. For example, the 6 million people in the region of Madrid (Spain) generate more than 600 000 tons of sewage sludge annually (Comunidad de Madrid, 2006). The most conventional uses of sewage sludges include industrial utilization, landfill, combustion and composting for farmland utilization (Sánchez Monedero and Mondini, 2004). The last one is considered the most sustainable use for sewage sludge when there is suitable land available for application. Sludges provide organic matter to soils,
⇑ Corresponding author. Address: Departamento de Edafología, E.T.S.I. Agrónomos, Universidad Politécnica de Madrid, Avda Complutense s.n., 28040 Madrid, Spain. Tel.: +34 913365681. E-mail address: [email protected] (G. Gascó). 0045-6535/$ - see front matter Ó 2012 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.chemosphere.2012.05.092
which is favorable due to an important fraction of the agricultural soils presenting soil organic matter values under the recommended levels for an adequate productive function. In particular, semiarid climatic conditions in the Mediterranean, coupled with intensive tillage systems have resulted in soil organic matter losses and consequently low organic matter stocks (Romanyà and Rovira, 2011). Sewage sludge application can also improve physical properties such as soil structure, infiltration rate and water holding capacity (Sort and Alcañiz, 1999) or soil respiration (HernándezApaolaza et al., 2000). However, sewage sludges are characterized by high metal content and appreciable amounts of pathogenic organisms that could constitute a contamination source for soil (Wang et al., 2008) and, due to potential leaching of the contaminant, to groundwater (Gascó et al., 2005a). As a consequence, soil amended with sewage sludge can result in heavy metal accumulation in plant tissues (Gascó and Lobo, 2007; Jamali et al., 2009).
A. Méndez et al. / Chemosphere 89 (2012) 1354–1359
Most of the research on pyrolysis of sewages sludge has focused on energy and fuel quality (Fonts et al., 2009), pyrolysis behavior (Gascó et al., 2005b; Sánchez et al., 2009) and the use of the solid fraction (biochar) as adsorbents to remove water pollutants (Méndez et al., 2005; Wang et al., 2011) while few studies have quantified the effect of adding biochar from sewage sludge to soil, their stability and even fewer in the leaching of heavy metals. Thermal processing of sewage sludges provides advantages as it reduces the waste volume and transport costs. In this way, sludges can be transformed into biochar by pyrolysis and then applied to the land. This procedure has mainly two advantages, firstly, it removes pathogens from sewage sludge and, secondly, biochar can improve the structure of soil, increase agricultural output and at the same time contribute to carbon sequestration due to carbon stability of biochar materials (Lehmann, 2007; Laird et al., 2010) and to improve soil biological properties and soil quality (PazFerreiro et al., 2012). Recently, biochar from pyrolysis of different wastes has been used in multiple ways in soil remediation due to its adsorption of pesticides (Zheng et al., 2010) or metals, both in laboratory (Hossain et al., 2010; Karami et al., 2011) and field (Fellet et al., 2011) studies. However, as far as we know, in previous studies there is no comparison between the sewage sludge biochar and the original sewage sludge. Thus, it is mostly unknown whether transforming a sewage sludge into biochar would result in increased or decreased levels of heavy metal availability. In addition, the effects of biochar addition to soil has been tested on a variety of soils, however, to date there are no studies on how biochar can affect Mediterranean soils. Taking into account the above considerations, we performed an experiment in a typical Mediterranean agricultural soil. Our aim was to evaluate if pyrolysis could provide a more sustainable way of disposing from sewage sludges. 2. Materials and methods 2.1. Soil characterization The selected soil (S) was sampled in central Spain area. The soil is classified as a Haplic Cambisol, it has a sandy-loam texture, basic pH (8.63) and low organic carbon content (0.66%). This agricultural soil is representative of many parts of the Mediterranean landscape, as it has an organic matter value in the average value of an extensive work recently conducted (Romanyà and Rovira, 2011). Other properties, in our soil such as clay and pH values are also typical of Mediterranean agricultural soils. Soil was airdried, crushed and sieved through a 2 mm mesh prior to analysis. Initial pH and electrical conductivity (EC) were determined in a ratio soil: water 1:2.5 (g mL1) by a Crison micro-pH 2000 (Thomas, 1996) in the case of pH and with a Crison 222 conductivimeter (Rhoades, 1996) in the case of EC. Cation exchange capacity (CEC) was determined with NH4OAc/AcOH pH 7.0 (Sumner and Miller, 1996). Soil texture was determined following Bouyoucos methodology (Bouyoucos, 1962). Total organic matter (TOM) and total organic carbon (TOC) were measured by the dry combustion method at 540 °C (Nelson and Sommers, 1996). Total humic substances (THSs) of soil sample after incubation were extracted with a mixture of 1 M Na4P2O7 and 0.1 M NaOH, centrifuged at 3000 rpm, and filtered. An aliquot of this extract was acidified to pH = 1 using concentrated H2SO4 and centrifuged to separate coagulated humic acids (HAs). The non-coagulated fraction with H2SO4 is referred as fulvic acids (FAs) (Schnitzer, 1982). Soil metal content was determined using a Perkin Elmer 2280 atomic absorption spectrophotometer and a Varian 10/20 atomic absorption spectrophotometer with graphite furnace after sample
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extraction by digestion with 3:1 (v/v) concentrated HCl/HNO3 following USEPA-3051a method (USEPA, 1997). Field capacity (FC) and wilting point (WP) were determined as the soil moisture content at 33 kPa (FC) and 1500 kPa (WP) (Richards, 1954). Available water (AW) was calculated as the difference between FC and WP. The percentage of soil aggregates over 2 mm of diameter was determined by dry sieving. All the analyses were carried out in triplicate. 2.2. Organic waste characterization Sewage sludge (SL) was obtained after aerobic treatment of sludge from wastewater treatment plant in Madrid region. SL was air-dried, crushed, and passed through a 2-mm sieve. pH, EC, CEC, TOM, TOC, THS and heavy metal content in the sludge were determined as described for the soil (Section 2.1). 2.3. Preparation and characterization of biochar Sewage sludge (200 g) was placed in a ceramic cup placed in an electrical tubular furnace Carbolite. Samples were pyrolyzed by increasing the temperature to 500 °C at a rate of 3 °C min1 using a nitrogen flow rate of 150 mL min1. The final temperature was maintained for 2 h. pH, EC, CEC, TOC and heavy metal content in the biochar were determined as described for the soil. Biochar nitrogen adsorption analysis to determined BET surface was carried out at 77 K in a Micromeritics Tristar 3000 (Instituto de Catálisis y Petroquímica, CSIC, Spain). Fixed C content, ash and volatile fraction for biochar was determined by thermogravimetry. Briefly, samples were heated up to a temperature of 900 °C under an N2 atmosphere at a flux of 40 mL min1. At 900 °C, air was introduced until a constant weight was reached. Humidity was calculated as the weight loss from the initial temperature to 150 °C. Volatile fraction was determined as the weight loss from 150 °C to 900 °C under N2 atmosphere and fixed C as the weight produced when the final sample was burnt under an air atmosphere. Ashes were determined as the final weight of the samples. 2.4. Treatments Selected soil (T) was amended with the sewage sludge (SL) and the biochar (B) at two different rates in mass: 4% and 8%, leading to SL4, SL8, B4 and B8 treatments. All the treatments were replicated three times. 2.5. Incubation procedure The biological activity of non-amended and amended soils was evaluated by soil respiration (cumulative CO2 evolution) and total mineralization coefficient (TMC). Samples (200 g) were introduced in 1 L glass vessel and the CO2 evolved was evaluated during 200 d at a temperature of 28 ± 2 °C. Soils were incubated at 60% FC and this water content was maintained over the 200 d of evaluation. Three replicates per treatment were used. The decomposition rate was determined by passing CO2 and NH3 free air through the respiration vessels, trapping the evolving CO2 in 50 mL of 1 M NaOH. Titration of trapped CO2 was performed with 1 M HCl after precipitation of carbonates using BaCl2. CO2 measurements of each soil were taken periodically, 2, 3, 4, 5, 7, 9, 11, 16, 18, 22, 50, 73, 120, 139 and 200 d after the beginning of the incubation. Three vessels without soil sample were used as blanks. Total mineralization coefficient (TMC) was calculated according to Díez et al. (1992) as follows:
TMCðmgC CO2 =g CÞ ¼ C CO2 evolved=initial TOC
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where C–CO2 evolved is expressed as mg C–CO2 100 g1 soil and initial TOC is expressed as g C 100 g1 soil. Plant-available metals were extracted from treated soil using diethylenetriaminepentaacetic acid–CaCl2–triethanolamine (DT PA) as described elsewhere (Lindsay and Norvell, 1969). The mobile forms of heavy metals were extracted using 0.1 M CaCl2 (Ure et al., 1995). To evaluate metal leached from each treatment, soil and water were mixed in a 1:2.5 ratio and agitated during 2 h. After this, solutions were centrifuged at 3000 rpm during 10 min to remove suspension particles. The content in Cu, Ni, Zn, Cd and Pb was evaluated by atomic absorption using a Perkin Elmer 2280 atomic absorption spectrophotometer and a Varian 10/20 atomic absorption spectrophotometer with graphite furnace. The same procedures were followed to measure plant-available, mobile forms and leached metals in the sewage sludge and biochar. 2.6. Statistical analysis The statistical analyses (calculation of means and standard deviations, differences of means) were performed using SPSS 15.0 package. Differences of means were tested using an analysis of variance (ANOVA). Means were considered to be different when P < 0.05 using the Tukey’s test.
pyrolysis of sewage sludge as a result of polymerization/condensation reactions. BET area of biochar was 32.24 m2 g1. Proximate analysis of biochar show an elevated ash content (67.5 wt%). Previous data (Gascó et al., 2012) have shown that after pyrolysis, carbonates and other inorganic compounds were concentrated in the biochar. 3.2. Effects of biochar on soil carbon mineralization Soil respiration followed the order T < B8 < B4 < SL4 < SL8. (Table 2). Soil respiration increases after biochar addition (Gascó et al., 2012) have been attributed to biochar being not as inert as previously believed but providing significant amounts of labile C (Smith et al., 2010), used as soil as an energy source by soil microorganisms. Respiration per unit of TOC (TMC) followed the order SL4 = SL8 > B4 > T > B8. High values of TMC, as shown by the sludges and the treatment B4, result in a more fragile humus and thus in a lower quality soil. The treatment B8 presents lower TMC compared to the control soil, meaning that organic matter is conserved more efficiently and maintaining the activity of the microorganisms responsible for soil organic matter biodegradation. 3.3. Effects of biochar and sewage sludge amendments on soil properties
3. Results and discussion 3.1. Initial amendment characterization As expected, pyrolysis affected the properties of the sludge biochar (Table 1). Pyrolysis increased the pH of the sewage sludge by almost 2.5 units, from 6.98 to 9.54, while decreasing EC from 1.055 to 0.554 dS m1. CEC was almost four times higher in the sewage sludge than in the biochar (9.16 and 2.36 cmol kg1, respectively). Both, sewage sludge and biochar exhibited high contents for all the heavy metals studied. Copper content increased by 80%, when comparing the biochar with the sewage sludge, and by 40% for other heavy metals. On the contrary, TOM decreased by 51% after
Table 1 General properties for the soil, sewage sludge and biochar used in the study.
Pyrolysis yield (wt%) pH EC (1:2.5) (lS cm1, 25 °C) TOM (%) C (total humic substances) (%) C (humic acids) (%) C (fulvic acids) (%) CEC (cmol(+) kg1) Soil moisture at 33 kPa (%) Soil moisture at 1500 kPa (%) Ca (g kg1) Mg (g kg1) Na (g kg1) K (g kg1) Cu (mg kg1) Ni (mg kg1) Cd (mg kg1) Zn (mg kg1) Pb (mg kg1) Clay (<0.002 mm) (%) Silt (0.002–0.05 mm) (%) Sand (0.05–2 mm) (%) Texture BET (m2 g1) Volatile organic content (%) Ash content (%) Fixed carbon content (%)
Soil
Sewage sludge
Biochar
– 8.63 71.6 1.16 0.56 0.21 0.35 6.6 7.62 4.55
– 6.98 1055 60.4 2.25 1.82 0.43 9.16
45 9.54 554 29.85
2.36
20.24 3.94 0.74 6.47 151 25 1.28 900 136
222 35 1.79 1250 196
7.6 13.35 4.35 48.45 26.8 10.0 12.5 77.5 Sandy-loam
6.10
32.24 24.1 67.5 5.7
Table 3 shows the changes of the chemical properties of soils after the application of biochar and sludge at two different rates. The largest increase in both TOC and TOM at the end of the incubation happened when adding SL8 which resulted in a 3.5 fold increase in both TOC and TOM compared to the control soil. The order of both TOC and TOM at the end of the incubation was SL8 > B8 > B4 = SL4 > T. Biochar increased soil pH by 0.2 units when a high dose (B8) was added (Table 3). This result is expected as the biochar used in our experiment had a much higher pH (9.5) than the soil (8.6). However, sewage sludge addition with the equivalent amount of sewage sludge (SL8) decreased pH to 7.44. Sewage sludge amendment multiplied EC value by 15 in the soil, while biochar amendment quadruplicated the value of EC in the soil compared to the control (Table 3). Hossain et al. (2010) reported an increase on soil electrical conductivity after biochar addition, similar as found in our study. Electrolytes added to soil affects its flocculation (Brady and Weil, 2002). Thus, we could expect this to affect crops sensitive to increased salt concentrations or soils with unstable structure. Table 3 shows that sewage sludge increased the value of CEC, but biochar amendment did not have any effect on soil CEC. It has been found that Anthrosols from the Brazilian Amazon (ages between 600 and 8700 years before present) with high biomass derived black carbon content had greater potential cation exchange capacity per unit of organic carbon than adjacent oxisols (Liang et al., 2006), but generally it is admitted that the effect of biochar addition on CEC is dependent on the type of biomass Table 2 CO2–C evolved (mg C/100 g dry weight) and total mineralization coefficient (TMC) for control and amended soils during incubation experiment. Soil sample
T SL4 SL8 B4 B8
CO2 evolved
TMC
(mg C–CO2/100 g)
(mg C–CO2 g C1)
67 ± 8a 477 ± 24d 691 ± 13e 163 ± 13c 116 ± 6b
100 ± 12b 276 ± 14d 264 ± 5d 132 ± 11c 64 ± 4a
Values in a given column followed by the same letter are not significantly different (P = 0.05) using Tukey test.
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A. Méndez et al. / Chemosphere 89 (2012) 1354–1359 Table 3 Total organic carbon (TOC), cation exchange capacity (CEC), pH and electrical conductivity (EC) of control and amended soils after the incubation experiment. Soil sample
CEC (cmol(c) kg1)
TOC (%) a
T SL4 SL8 B4 B8
0.66 ± 0.05 1.30 ± 0.04b 2.42 ± 0.05d 1.28 ± 0.07b 1.74 ± 0.08c
EC (1:2.5) (lS cm1, 25 °C)
pH (1:2.5)
a
b
6.76 ± 0.16 7.94 ± 0.28b 9.97 ± 0.13c 7.11 ± 0.04a 7.16 ± 0.06a
73 ± 7a 978 ± 8c 1124 ± 9c 293 ± 7b 304 ± 6b
7.84 ± 0.04 7.80 ± 0.08b 7.44 ± 0.05a 7.83 ± 0.02b 8.01 ± 0.02c
Values in a given column followed by the same letter are not significantly different (P = 0.05) using Tukey test.
Table 4 Field capacity (FC), wilting point (WP), available water (AW) and percentage of soil aggregates (diameter > 2 mm) of control and amended soils after the incubation experiment. Soil sample
FC (%)
WP (%)
AW (%)
Soil aggregates (%)
T SL4 SL8 B4 B8
10.65 ± 0.07a 10.85 ± 0.08a 12.68 ± 0.12c 12.32 ± 0.12b 13.83 ± 0.10d
5.05 ± 0.04a 5.48 ± 0.04c 7.76 ± 0.06e 5.25 ± 0.07b 6.01 ± 0.07d
5.60 ± 0.09b 5.37 ± 0.06b 4.92 ± 0.15a 7.07 ± 0.06c 7.83 ± 0.14d
22.2 ± 3.1ab 27.2 ± 2.1bc 29.0 ± 1.9c 20.4 ± 2.27a 18.3 ± 1.50a
Values in a given column followed by the same letter are not significantly different (P = 0.05) using Tukey test.
Table 5 Metal content in sewage sludge and biochar after leaching experiment (mg L1) and after extraction with CaCl2 or DTPA (mg kg1). Sample
Cu
Ni
Cd
Pb
0.071 ± 0.005a 0.034 ± 0.001b
0.008 ± 0.001a 0.002 ± 0.001b
0.0022 ± 0.0006a 0.0013 ± 0.0004a
31.04 ± 1.21a 0.63 ± 0.12b
90.05 ± 2.13a 29.60 ± 1.27b
0.503 ± 0.175a 0.321 ± 0.112b
68.60 ± 3.78a 8.82 ± 0.07b
5.45 ± 0.42a 4.10 ± 0.31b
247.54 ± 7.60a 24.52 ± 2.33b
0.350 ± 0.107a 0.302 ± 0.101a
12.540 ± 0.107a 2.452 ± 0.072b
Metal content after leaching experiment (mg L1) L 0.300 ± 0.085a B 0.012 ± 0.006b
0.081 ± 0.004a 0.029 ± 0.015b
Metal content after extraction with DTPA (mg kg1) L 85.18 ± 4.10a B 43.55 ± 2.47b Metal content after extraction with CaCl2 (mg kg1) L 34.24 ± 4.31a B 14.8 ± 2.55b
Zn
Values in a given column followed by the same letter are not significantly different (P = 0.05) using Tukey test.
pyrolyzed (Jha et al., 2010). In addition, CEC of biochar is scarce after low pyrolysis temperatures, as the ones in our study, increasing at higher temperatures (Lehmann, 2007). Our results agreed with other studies reporting biochar to have a minimal CEC when compared to soil organic matter (Cheng et al., 2006, 2008; Lehmann, 2007). Amendments affected the value of FC in the soil, following the order T = SL4 < B4 < SL8 < B8 (see Table 4). In a similar way, the value of WP followed the order T < B4 < SL4 < B8 < SL8. On the other hand, available water (AW) was lower in the SL8 treatment compared with the soil without amendments, while biochar amendment significantly increased AW of soil. Coarse-textured soils have little water holding capacity and this allows for rapid movement of infiltrating water to depth. The use of biochar could provide a medium where there is less drainage and thus has the potential for a better plant growth. The improvement of water retention after biochar addition has previously been observed by Beck et al. (2011). However, biochar did not improve soil aggregation in our samples (Table 4) similarly to the results reported by Peng et al. (2011). Water-soluble inorganic pollutants may constitute an environmental toxicity problem when they are plant-available or when their movement through soils and potential transfer to groundwater is not impeded. Table 1 shows that metal content of biochar sample was significantly higher than metal content in sewage sludge due to ash concentration during pyrolysis. Nevertheless, plant-available metals (DTPA extraction) and the mobile forms of Cu, Ni, Zn and Pb decreased after pyrolysis process; also the risk of leaching of Cu, Ni, Zn and Cd decreased in biochar sample (Table
5). Previous works carried out by our research group have showed that metals leaching decreases after sewage sludge pyrolysis (Méndez and Gascó, 2005; Gascó et al., 2004) and in some cases no metal lixiviation was observed after pyrolysis treatment (Méndez et al., 2005). Also, Hwang et al. (2007) have shown that metals leached from sewage sludge decreased in the pyrolyzed product and Martín et al. (2003) observed that the risk of metals lixiviation in pyrolysis ashes is lower than in the incineration ones. With respect to leaching experiment in soil (Table 6), Cd and Pb contents did not vary for the different treatments considered while sewage sludge amendments increased the Cu, Ni and Zn leaching with respect to the control, especially as higher is the sewage sludge ratio. This fact is according to Gascó et al. (2005a,b) that observed an increment in Ni and Cu leaching in an acid soil amended with sewage sludge in a soil column experiment increasing the risk of harmful effects in soils with less clay content due to the influence of clay in metal retention. Indeed, Toribio and Romanyà (2006) found that concentration of metals and organic matter in the leachates depended on the soil characteristics and on the type of sewage sludge added to the soil. Basic soils with a high amount of clay showed the highest metal retention capacity, while acid soils with low clay content showed the lowest. So, soil pH and clay content can influence the metal lixiviation in soil. Also, Ahlberg et al. (2006) have showed that high metal content of sewage sludges lead to metal leaching after soil addition being one of the sewage sludge amendments restrictions. The leaching of Cu, Ni and Zn were higher in the soil treated by sewage sludge than in soil treated with biochar according to the lower leaching of metals in the biochar (Table 5). Previous works
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Table 6 Metals content (mg L1) in final solutions after leaching experiment of control and amended soils after the incubation experiment. Soil sample T SL4 SL8 B4 B8
Cu
Ni a
0.153 ± 0.008 0.268 ± 0.051b 0.345 ± 0.019c 0.142 ± 0.007a 0.155 ± 0.017a
Zn b
0.041 ± 0.008 0.062 ± 0.021bc 0.092 ± 0.015c 0.027 ± 0.005d 0.041 ± 0.002b
Cd bc
0.026 ± 0.05 0.034 ± 0.003cd 0.038 ± 0.003d 0.013 ± 0.004a 0.019 ± 0.006ab
Pb a
0.003 ± 0.002 0.004 ± 0.006a 0.005 ± 0.001a 0.002 ± 0.002a 0.002 ± 0.001a
0.0040 ± 0.0003a 0.0041 ± 0.0005a 0.0053 ± 0.0010a 0.0041 ± 0.0003a 0.0043 ± 0.0005a
Values in a given column followed by the same letter are not significantly different (P = 0.05) using Tukey test.
Table 7 Metal content (mg kg1) in samples extracted with CaCl2 or DTPA. Soil sample
Ni
Zn
Cd
Pb
T(CaCl2) SL4(CaCl2) SL8(CaCl2) B4(CaCl2) B8(CaCl2)
Cu 0.82 ± 0.05a 1.85 ± 0.05c 2.41 ± 0.10b 0.97 ± 0.10a 2.42 ± 0.09b
4.361 ± 1.930a 7.293 ± 0.260b 6.627 ± 0.297ab 5.837 ± 0.366ab 5.423 ± 0.137ab
0.419 ± 0.002a 1.160 ± 0.061c 1.257 ± 0.118c 1.213 ± 0.133c 0.790 ± 0.105b
0.100 ± 0.001a 0.493 ± 0.086bc 0.550 ± 0.044c 0.510 ± 0.115bc 0.357 ± 0.055b
0.850 ± 0.050a 4.477 ± 1.208c 3.850 ± 0.401bc 2.147 ± 0.165b 2.860 ± 0.678bc
T(DTPA) SL4(DTPA) SL8(DTPA) B4(DTPA) B8(DTPA)
4.433 ± 0.183a 7.046 ± 0.187b 7.315 ± 0.362b 6.838 ± 0.553b 7.000 ± 0.110b
0.933 ± 0.291a 5.023 ± 0.427c 4.880 ± 0.689c 3.303 ± 0.208b 2.773 ± 0.485b
0.761 ± 0.136a 2.997 ± 0.915b 3.753 ± 0.690b 0.930 ± 0.046a 1.133 ± 0.093a
0.190 ± 0.09a 0.580 ± 0.181b 0.507 ± 0.332b 0.177 ± 0.133ab 0.168 ± 0.08a
0.740 ± 0.200a 2.457 ± 0.072bc 4.233 ± 1.703c 1.193 ± 0.175ab 1.313 ± 0.471ab
Values in a given column followed by the same letter are not significantly different (P = 0.05) using Tukey test.
have proved that biochar is an effective material for reducing high concentrations of soluble metals like Cd and Zn from a contaminated soil (Beesley and Marmiroli, 2011). These authors concluded that sorption and not pH increased is the more important mechanism by which those metals are retained. In our work, the leaching of Ni and Zn decreased for lower biochar additions (B4) with respect to control soil though only high addition of biochar (B8) increased the soil pH. So, according to Beesley and Marmiroli (2011) sorption mechanism by biochar seems to be the more important factor controlling the metal leaching decrement in biochar amended soils. Finally, there were not differences in leached metals between control soil and B8 treatments which is quite interesting to control the risk of contamination of groundwater by trace metals. Table 7 shows the result for the bioavailable extracted heavy metals and for the mobile forms. In general, sewage sludge increased soil heavy metal bioavailability and mobility with respect to control soil, while biochar addition resulted in less increased heavy metal bioavailability and mobility according to both bioavailable and mobile metal content of raw materials (Table 5). Indeed, Gascó and Lobo (2007) found that sewage sludge significantly increased the level of trace elements in plants. If the highest treatments are compared, the plant bioavailability (DTPA) of Ni, Zn, Cd and Pb in B8 treatment were 57%, 30%, 29% and 31% of those in SL8. Our results agrees with Hossain et al. (2010) who measured that the contents of heavy metals in cherry tomato plants amended with sewage sludge biochar were below the Australian maximum permitted concentrations for food products. Also, Hua et al. (2009) reported a lower DTPA extracted Cd in soils amended with biochar. The results with respect to Ni are particularly important as in sewage sludge, Ni occurs in organic chelated forms that are readily available to plants and thus it can be highly phytotoxic. In addition, high Ni concentrations reduce the plant’s uptake of most other nutrients due to damaging effect on the root (Mengel, 1978). Comparing the mobile forms (CaCl2 extracted metals), there were not differences between the soils amended by sludge and biochar for Cu, Ni and Pb in the high rates. While for Zn–CaCl2 and Cd–CaCl2, the concentrations in B8 were 63% and 65% of those in SL8.
In summary, even through the amount of metals added by biochar is higher than total metal added by sewage sludge, metals seems to be retained stronger by biochar. Sewage sludge biochar can be used as a valuable soil amendment as it increases the organic matter content and can hinder the leaching of heavy metals present in raw sewage sludge. 4. Conclusions The sewage sludge pyrolysis decreased the plant-available of Cu, Ni, Zn and Pb, the mobile forms of Cu, Ni, Zn, Cd and Pb and also the risk of leaching of Cu, Ni, Zn and Cd. Sewage sludge and biochar amendments increased soil respiration with respect to the control soil. The increase was lower in the case of biochar than when sewage sludge was directly added. The leaching of Cu, Ni and Zn were lower in the soil treated with biochar that in sewage sludge treatment. Biochar amended samples also reduced plant availability of Ni, Zn, Cd and Pb when compared to sewage sludge amended samples. The amendment of the sandy loam soil with biochar from sewage sludge pyrolysis significantly increased the available water and field capacity of soil. However, only sewage sludge addition improved the percentage of soil aggregates. Acknowledgments Authors wish to thank Madrid Community-Universidad Politécnica de Madrid for the support through Project CCG10UPM/AMB-5683. References Ahlberg, G., Gustafsson, O., Wedel, P., 2006. Leaching of metals from sewage sludge during one year and their relationship to particle size. Environ. Pollut. 144, 545– 553. Beck, D.A., Johnson, G.R., Spolek, G.A., 2011. Amending greenroof soil with biochar to affect runoff water quantity and quality. Environ. Pollut. 159, 2111–2118. Beesley, L., Marmiroli, M., 2011. The immobilisation and retention of soluble arsenic, cadmium and zinc by biochar. Environ. Pollut. 159, 474–480. Bouyoucos, G.J., 1962. Hydrometer method improved for making particle size analyses of soil. Agron. J. 54, 464–465.
A. Méndez et al. / Chemosphere 89 (2012) 1354–1359 Brady, N.C., Weil, R.R., 2002. The Nature and Properties of Soils, 13th ed. Prentice Hall, Upper Saddle River, New York. Cheng, C.H., Lehmann, J., Engelhard, M., 2008. Natural oxidation of black carbon in soils: changes in molecular form and surface charge along a climosequence. Geochim. Cosmochim. Acta 72, 1598–1610. Cheng, C.H., Lehmann, J., Thies, J.E., Burton, S.D., Engelhard, M.H., 2006. Oxidation of black carbon by biotic and abiotic processes. Org. Geochem. 37, 1477–1488. Comunidad de Madrid, 2006. Plan Regional de Lodos de Depuradora (2006–2016), Comunidad de Madrid, Spain. Díez, J.A., Polo, A., Guerrero, F., 1992. Effect of sewage sludge on nitrogen availability in peat. Biol. Fert. Soils 13, 248–251. Fellet, G., Marchiol, L., Delle Vedove, G., Peressotti, A., 2011. Application of biochar on mine tailings: effects and perspectives for land reclamation. Chemosphere 83, 1262–1267. Fonts, I., Azuara, M., Gea, G., Murillo, M.B., 2009. Study of the pyrolysis liquids obtained from different sewage sludge. J. Anal. Appl. Pyrol. 85, 184–191. Gascó, G., Paz-Ferreiro, J., Méndez, A., 2012. Thermal analysis of soil amended with sewage sludge and biochar from sewage sludge pyrolysis. J. Therm. Anal. Calorim. 108, 769–775. Gascó, G., Lobo, M.C., Guerrero, F., 2005a. Land application of sewage sludge: a soil columns study. Water SA 31 (3), 309–318. Gascó, G., Blanco, C.G., Guerrero, F., Méndez, A., 2005b. The influence of organic matter on sewage sludge pyrolysis. J. Anal. Appl. Pyrol. 74, 413–420. Gascó, G., Lobo, M.C., 2007. Composition of a Spanish sewage sludge and effects on treated soil and olive trees. Waste Manage. 27, 1494–1500. Gascó, G., Méndez, A., Gascó, J.M., 2004. Metals adsorption by carbon adsorbent prepared from sewage sludges pyrolysis. In: XXIX Reuniao Ibérica de Adsorcao, Porto, Portugal. Hernández-Apaolaza, L., Gascó, J.M., Guerrero, F., 2000. Initial organic matter transformation of soil amended with composted sewage sludge. Biol. Fert. Soils 32, 421–426. Hossain, M.K., Strezov, A., Chan, K.Y., Nelson, P.F., 2010. Agronomic properties of wastewater sludge biochar and bioavailability of metals in production of cherry tomato (Lycopersicon esculentum). Chemosphere 78, 1167–1171. Hua, L., Wu, W., Liu, Y., McBride, M.B., Chen, Y., 2009. Reduction of nitrogen loss and Cu and Zn mobility during sludge composting with bamboo charcoal amendment. Environ. Sci. Pollut. Res. 16, 1–9. Hwang, I.H., Ouchi, Y., Matsuto, T., 2007. Characteristics of leachate from pyrolysis residue of sewage sludge. Chemosphere 68, 1913–1919. Jamali, M.K., Kazi, T.G., Arain, M.B., Afridi, H.I., Jalbani, N., Kandhro, G.A., Shah, A.Q., Baig, J.A., 2009. Heavy metal accumulation in different varieties of wheat (Triticum aestivum L.) grown in soil amended with domestic sewage sludge. J. Hazard. Mater. 164, 1386–1391. Jha, P., Biswas, A.K., Lakaria, B.L., Subba Rao, A., 2010. Biochar in agriculture prospects and related implications. Curr. Sci. India 99, 1218–1225. Karami, N., Clemente, R., Moreno-Jiménez, E., Lepp, N.W., Beesley, L., 2011. Efficiency of green waste compost and biochar soil amendments for reducing lead and copper mobility and uptake to ryegrass. J. Hazard. Mater. 191, 41–48. Laird, D., Fleming, P., Wang, B., Horton, R., Karlen, D., 2010. Biochar impact on nutrient leaching from a midwestern agricultural soil. Geoderma 158, 436–442. Lehmann, J., 2007. Bio-energy in the black. Front. Ecol. Environ. 5, 381–387. Liang, B., Lehmann, J., Solomon, D., Kinyangi, J., Grossman, J., O’Neill, B., Skjemstad, J.O., Thies, J., Luizao, F.J., Petersen, J., Neves, E.G., 2006. Black carbon increases cation exchange capacity in soils. Soil Sci. Soc. Am. J. 70, 1719–1730. Lindsay, W.L., Norvell, W.A., 1969. Development of a DTPA micronutrients soil test. Agron. Abs. 69, 84. Martín, M.J., Artola, A., Balaguer, M.D., Rigola, M., 2003. Activated carbons developed from surplus sewage sludge for the removal of dyes from dilute aqueous solutions. Chem. Eng. J. 94, 231–239.
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Effects of sewage sludge biochar on plant metal availability after application to a Mediterranean soil A. Méndez a, A. Gómez b, J. Paz-Ferreiro b, G. Gascó b,⇑ a b
Departamento de Ingeniería de Materiales, E.T.S.I. Minas, Universidad Politécnica de Madrid, C/Ríos Rosas No. 21, 28003 Madrid, Spain Departamento de Edafología, E.T.S.I. Agrónomos, Universidad Politécnica de Madrid, Ciudad Universitaria, 28004 Madrid, Spain
h i g h l i g h t s " Biochar can reduce the leaching of heavy metals present in raw sewage sludge, in particular nickel and zinc. " Biochar amended samples also reduced plant heavy metal availability when compared to sewage sludge amended samples. " Biochar and sewage sludge increased soil respiration with respect to the control soil. " The increment of soil respiration was lower in the case of biochar than when sewage sludge was added.
a r t i c l e
i n f o
Article history: Received 28 February 2012 Received in revised form 9 May 2012 Accepted 23 May 2012 Available online 23 June 2012 Keywords: Biochar Heavy metals Sewage sludge Soil
a b s t r a c t Pyrolytic conversion of sewage sludge into biochar could be a sustainable management option for Mediterranean agricultural soils. The aim of this work is to evaluate the effects of biochar from sewage sludge pyrolysis on soil properties; heavy metals solubility and bioavailability in a Mediterranean agricultural soil and compared with those of raw sewage sludge. Biochar (B) was prepared by pyrolysis of selected sewage sludge (SL) at 500 °C. The pyrolysis process decreased the plant-available of Cu, Ni, Zn and Pb, the mobile forms of Cu, Ni, Zn, Cd and Pb and also the risk of leaching of Cu, Ni, Zn and Cd. A selected Mediterranean soil was amended with SL and B at two different rates in mass: 4% and 8%. The incubation experiment (200 d) was conducted in order to study carbon mineralization and trace metal solubility and bioavailability of these treatments. Both types of amendments increased soil respiration with respect to the control soil. The increase was lower in the case of B than when SL was directly added. Metals mobility was studied in soil after the incubation and it can be established that the risk of leaching of Cu, Ni and Zn were lower in the soil treated with biochar that in sewage sludge treatment. Biochar amended samples also reduced plant availability of Ni, Zn, Cd and Pb when compared to sewage sludge amended samples. Ó 2012 Elsevier Ltd. All rights reserved.
1. Introduction Industrial development and economic growth imply an increase in waste generation. For example, the 6 million people in the region of Madrid (Spain) generate more than 600 000 tons of sewage sludge annually (Comunidad de Madrid, 2006). The most conventional uses of sewage sludges include industrial utilization, landfill, combustion and composting for farmland utilization (Sánchez Monedero and Mondini, 2004). The last one is considered the most sustainable use for sewage sludge when there is suitable land available for application. Sludges provide organic matter to soils,
⇑ Corresponding author. Address: Departamento de Edafología, E.T.S.I. Agrónomos, Universidad Politécnica de Madrid, Avda Complutense s.n., 28040 Madrid, Spain. Tel.: +34 913365681. E-mail address: [email protected] (G. Gascó). 0045-6535/$ - see front matter Ó 2012 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.chemosphere.2012.05.092
which is favorable due to an important fraction of the agricultural soils presenting soil organic matter values under the recommended levels for an adequate productive function. In particular, semiarid climatic conditions in the Mediterranean, coupled with intensive tillage systems have resulted in soil organic matter losses and consequently low organic matter stocks (Romanyà and Rovira, 2011). Sewage sludge application can also improve physical properties such as soil structure, infiltration rate and water holding capacity (Sort and Alcañiz, 1999) or soil respiration (HernándezApaolaza et al., 2000). However, sewage sludges are characterized by high metal content and appreciable amounts of pathogenic organisms that could constitute a contamination source for soil (Wang et al., 2008) and, due to potential leaching of the contaminant, to groundwater (Gascó et al., 2005a). As a consequence, soil amended with sewage sludge can result in heavy metal accumulation in plant tissues (Gascó and Lobo, 2007; Jamali et al., 2009).
A. Méndez et al. / Chemosphere 89 (2012) 1354–1359
Most of the research on pyrolysis of sewages sludge has focused on energy and fuel quality (Fonts et al., 2009), pyrolysis behavior (Gascó et al., 2005b; Sánchez et al., 2009) and the use of the solid fraction (biochar) as adsorbents to remove water pollutants (Méndez et al., 2005; Wang et al., 2011) while few studies have quantified the effect of adding biochar from sewage sludge to soil, their stability and even fewer in the leaching of heavy metals. Thermal processing of sewage sludges provides advantages as it reduces the waste volume and transport costs. In this way, sludges can be transformed into biochar by pyrolysis and then applied to the land. This procedure has mainly two advantages, firstly, it removes pathogens from sewage sludge and, secondly, biochar can improve the structure of soil, increase agricultural output and at the same time contribute to carbon sequestration due to carbon stability of biochar materials (Lehmann, 2007; Laird et al., 2010) and to improve soil biological properties and soil quality (PazFerreiro et al., 2012). Recently, biochar from pyrolysis of different wastes has been used in multiple ways in soil remediation due to its adsorption of pesticides (Zheng et al., 2010) or metals, both in laboratory (Hossain et al., 2010; Karami et al., 2011) and field (Fellet et al., 2011) studies. However, as far as we know, in previous studies there is no comparison between the sewage sludge biochar and the original sewage sludge. Thus, it is mostly unknown whether transforming a sewage sludge into biochar would result in increased or decreased levels of heavy metal availability. In addition, the effects of biochar addition to soil has been tested on a variety of soils, however, to date there are no studies on how biochar can affect Mediterranean soils. Taking into account the above considerations, we performed an experiment in a typical Mediterranean agricultural soil. Our aim was to evaluate if pyrolysis could provide a more sustainable way of disposing from sewage sludges. 2. Materials and methods 2.1. Soil characterization The selected soil (S) was sampled in central Spain area. The soil is classified as a Haplic Cambisol, it has a sandy-loam texture, basic pH (8.63) and low organic carbon content (0.66%). This agricultural soil is representative of many parts of the Mediterranean landscape, as it has an organic matter value in the average value of an extensive work recently conducted (Romanyà and Rovira, 2011). Other properties, in our soil such as clay and pH values are also typical of Mediterranean agricultural soils. Soil was airdried, crushed and sieved through a 2 mm mesh prior to analysis. Initial pH and electrical conductivity (EC) were determined in a ratio soil: water 1:2.5 (g mL1) by a Crison micro-pH 2000 (Thomas, 1996) in the case of pH and with a Crison 222 conductivimeter (Rhoades, 1996) in the case of EC. Cation exchange capacity (CEC) was determined with NH4OAc/AcOH pH 7.0 (Sumner and Miller, 1996). Soil texture was determined following Bouyoucos methodology (Bouyoucos, 1962). Total organic matter (TOM) and total organic carbon (TOC) were measured by the dry combustion method at 540 °C (Nelson and Sommers, 1996). Total humic substances (THSs) of soil sample after incubation were extracted with a mixture of 1 M Na4P2O7 and 0.1 M NaOH, centrifuged at 3000 rpm, and filtered. An aliquot of this extract was acidified to pH = 1 using concentrated H2SO4 and centrifuged to separate coagulated humic acids (HAs). The non-coagulated fraction with H2SO4 is referred as fulvic acids (FAs) (Schnitzer, 1982). Soil metal content was determined using a Perkin Elmer 2280 atomic absorption spectrophotometer and a Varian 10/20 atomic absorption spectrophotometer with graphite furnace after sample
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extraction by digestion with 3:1 (v/v) concentrated HCl/HNO3 following USEPA-3051a method (USEPA, 1997). Field capacity (FC) and wilting point (WP) were determined as the soil moisture content at 33 kPa (FC) and 1500 kPa (WP) (Richards, 1954). Available water (AW) was calculated as the difference between FC and WP. The percentage of soil aggregates over 2 mm of diameter was determined by dry sieving. All the analyses were carried out in triplicate. 2.2. Organic waste characterization Sewage sludge (SL) was obtained after aerobic treatment of sludge from wastewater treatment plant in Madrid region. SL was air-dried, crushed, and passed through a 2-mm sieve. pH, EC, CEC, TOM, TOC, THS and heavy metal content in the sludge were determined as described for the soil (Section 2.1). 2.3. Preparation and characterization of biochar Sewage sludge (200 g) was placed in a ceramic cup placed in an electrical tubular furnace Carbolite. Samples were pyrolyzed by increasing the temperature to 500 °C at a rate of 3 °C min1 using a nitrogen flow rate of 150 mL min1. The final temperature was maintained for 2 h. pH, EC, CEC, TOC and heavy metal content in the biochar were determined as described for the soil. Biochar nitrogen adsorption analysis to determined BET surface was carried out at 77 K in a Micromeritics Tristar 3000 (Instituto de Catálisis y Petroquímica, CSIC, Spain). Fixed C content, ash and volatile fraction for biochar was determined by thermogravimetry. Briefly, samples were heated up to a temperature of 900 °C under an N2 atmosphere at a flux of 40 mL min1. At 900 °C, air was introduced until a constant weight was reached. Humidity was calculated as the weight loss from the initial temperature to 150 °C. Volatile fraction was determined as the weight loss from 150 °C to 900 °C under N2 atmosphere and fixed C as the weight produced when the final sample was burnt under an air atmosphere. Ashes were determined as the final weight of the samples. 2.4. Treatments Selected soil (T) was amended with the sewage sludge (SL) and the biochar (B) at two different rates in mass: 4% and 8%, leading to SL4, SL8, B4 and B8 treatments. All the treatments were replicated three times. 2.5. Incubation procedure The biological activity of non-amended and amended soils was evaluated by soil respiration (cumulative CO2 evolution) and total mineralization coefficient (TMC). Samples (200 g) were introduced in 1 L glass vessel and the CO2 evolved was evaluated during 200 d at a temperature of 28 ± 2 °C. Soils were incubated at 60% FC and this water content was maintained over the 200 d of evaluation. Three replicates per treatment were used. The decomposition rate was determined by passing CO2 and NH3 free air through the respiration vessels, trapping the evolving CO2 in 50 mL of 1 M NaOH. Titration of trapped CO2 was performed with 1 M HCl after precipitation of carbonates using BaCl2. CO2 measurements of each soil were taken periodically, 2, 3, 4, 5, 7, 9, 11, 16, 18, 22, 50, 73, 120, 139 and 200 d after the beginning of the incubation. Three vessels without soil sample were used as blanks. Total mineralization coefficient (TMC) was calculated according to Díez et al. (1992) as follows:
TMCðmgC CO2 =g CÞ ¼ C CO2 evolved=initial TOC
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where C–CO2 evolved is expressed as mg C–CO2 100 g1 soil and initial TOC is expressed as g C 100 g1 soil. Plant-available metals were extracted from treated soil using diethylenetriaminepentaacetic acid–CaCl2–triethanolamine (DT PA) as described elsewhere (Lindsay and Norvell, 1969). The mobile forms of heavy metals were extracted using 0.1 M CaCl2 (Ure et al., 1995). To evaluate metal leached from each treatment, soil and water were mixed in a 1:2.5 ratio and agitated during 2 h. After this, solutions were centrifuged at 3000 rpm during 10 min to remove suspension particles. The content in Cu, Ni, Zn, Cd and Pb was evaluated by atomic absorption using a Perkin Elmer 2280 atomic absorption spectrophotometer and a Varian 10/20 atomic absorption spectrophotometer with graphite furnace. The same procedures were followed to measure plant-available, mobile forms and leached metals in the sewage sludge and biochar. 2.6. Statistical analysis The statistical analyses (calculation of means and standard deviations, differences of means) were performed using SPSS 15.0 package. Differences of means were tested using an analysis of variance (ANOVA). Means were considered to be different when P < 0.05 using the Tukey’s test.
pyrolysis of sewage sludge as a result of polymerization/condensation reactions. BET area of biochar was 32.24 m2 g1. Proximate analysis of biochar show an elevated ash content (67.5 wt%). Previous data (Gascó et al., 2012) have shown that after pyrolysis, carbonates and other inorganic compounds were concentrated in the biochar. 3.2. Effects of biochar on soil carbon mineralization Soil respiration followed the order T < B8 < B4 < SL4 < SL8. (Table 2). Soil respiration increases after biochar addition (Gascó et al., 2012) have been attributed to biochar being not as inert as previously believed but providing significant amounts of labile C (Smith et al., 2010), used as soil as an energy source by soil microorganisms. Respiration per unit of TOC (TMC) followed the order SL4 = SL8 > B4 > T > B8. High values of TMC, as shown by the sludges and the treatment B4, result in a more fragile humus and thus in a lower quality soil. The treatment B8 presents lower TMC compared to the control soil, meaning that organic matter is conserved more efficiently and maintaining the activity of the microorganisms responsible for soil organic matter biodegradation. 3.3. Effects of biochar and sewage sludge amendments on soil properties
3. Results and discussion 3.1. Initial amendment characterization As expected, pyrolysis affected the properties of the sludge biochar (Table 1). Pyrolysis increased the pH of the sewage sludge by almost 2.5 units, from 6.98 to 9.54, while decreasing EC from 1.055 to 0.554 dS m1. CEC was almost four times higher in the sewage sludge than in the biochar (9.16 and 2.36 cmol kg1, respectively). Both, sewage sludge and biochar exhibited high contents for all the heavy metals studied. Copper content increased by 80%, when comparing the biochar with the sewage sludge, and by 40% for other heavy metals. On the contrary, TOM decreased by 51% after
Table 1 General properties for the soil, sewage sludge and biochar used in the study.
Pyrolysis yield (wt%) pH EC (1:2.5) (lS cm1, 25 °C) TOM (%) C (total humic substances) (%) C (humic acids) (%) C (fulvic acids) (%) CEC (cmol(+) kg1) Soil moisture at 33 kPa (%) Soil moisture at 1500 kPa (%) Ca (g kg1) Mg (g kg1) Na (g kg1) K (g kg1) Cu (mg kg1) Ni (mg kg1) Cd (mg kg1) Zn (mg kg1) Pb (mg kg1) Clay (<0.002 mm) (%) Silt (0.002–0.05 mm) (%) Sand (0.05–2 mm) (%) Texture BET (m2 g1) Volatile organic content (%) Ash content (%) Fixed carbon content (%)
Soil
Sewage sludge
Biochar
– 8.63 71.6 1.16 0.56 0.21 0.35 6.6 7.62 4.55
– 6.98 1055 60.4 2.25 1.82 0.43 9.16
45 9.54 554 29.85
2.36
20.24 3.94 0.74 6.47 151 25 1.28 900 136
222 35 1.79 1250 196
7.6 13.35 4.35 48.45 26.8 10.0 12.5 77.5 Sandy-loam
6.10
32.24 24.1 67.5 5.7
Table 3 shows the changes of the chemical properties of soils after the application of biochar and sludge at two different rates. The largest increase in both TOC and TOM at the end of the incubation happened when adding SL8 which resulted in a 3.5 fold increase in both TOC and TOM compared to the control soil. The order of both TOC and TOM at the end of the incubation was SL8 > B8 > B4 = SL4 > T. Biochar increased soil pH by 0.2 units when a high dose (B8) was added (Table 3). This result is expected as the biochar used in our experiment had a much higher pH (9.5) than the soil (8.6). However, sewage sludge addition with the equivalent amount of sewage sludge (SL8) decreased pH to 7.44. Sewage sludge amendment multiplied EC value by 15 in the soil, while biochar amendment quadruplicated the value of EC in the soil compared to the control (Table 3). Hossain et al. (2010) reported an increase on soil electrical conductivity after biochar addition, similar as found in our study. Electrolytes added to soil affects its flocculation (Brady and Weil, 2002). Thus, we could expect this to affect crops sensitive to increased salt concentrations or soils with unstable structure. Table 3 shows that sewage sludge increased the value of CEC, but biochar amendment did not have any effect on soil CEC. It has been found that Anthrosols from the Brazilian Amazon (ages between 600 and 8700 years before present) with high biomass derived black carbon content had greater potential cation exchange capacity per unit of organic carbon than adjacent oxisols (Liang et al., 2006), but generally it is admitted that the effect of biochar addition on CEC is dependent on the type of biomass Table 2 CO2–C evolved (mg C/100 g dry weight) and total mineralization coefficient (TMC) for control and amended soils during incubation experiment. Soil sample
T SL4 SL8 B4 B8
CO2 evolved
TMC
(mg C–CO2/100 g)
(mg C–CO2 g C1)
67 ± 8a 477 ± 24d 691 ± 13e 163 ± 13c 116 ± 6b
100 ± 12b 276 ± 14d 264 ± 5d 132 ± 11c 64 ± 4a
Values in a given column followed by the same letter are not significantly different (P = 0.05) using Tukey test.
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A. Méndez et al. / Chemosphere 89 (2012) 1354–1359 Table 3 Total organic carbon (TOC), cation exchange capacity (CEC), pH and electrical conductivity (EC) of control and amended soils after the incubation experiment. Soil sample
CEC (cmol(c) kg1)
TOC (%) a
T SL4 SL8 B4 B8
0.66 ± 0.05 1.30 ± 0.04b 2.42 ± 0.05d 1.28 ± 0.07b 1.74 ± 0.08c
EC (1:2.5) (lS cm1, 25 °C)
pH (1:2.5)
a
b
6.76 ± 0.16 7.94 ± 0.28b 9.97 ± 0.13c 7.11 ± 0.04a 7.16 ± 0.06a
73 ± 7a 978 ± 8c 1124 ± 9c 293 ± 7b 304 ± 6b
7.84 ± 0.04 7.80 ± 0.08b 7.44 ± 0.05a 7.83 ± 0.02b 8.01 ± 0.02c
Values in a given column followed by the same letter are not significantly different (P = 0.05) using Tukey test.
Table 4 Field capacity (FC), wilting point (WP), available water (AW) and percentage of soil aggregates (diameter > 2 mm) of control and amended soils after the incubation experiment. Soil sample
FC (%)
WP (%)
AW (%)
Soil aggregates (%)
T SL4 SL8 B4 B8
10.65 ± 0.07a 10.85 ± 0.08a 12.68 ± 0.12c 12.32 ± 0.12b 13.83 ± 0.10d
5.05 ± 0.04a 5.48 ± 0.04c 7.76 ± 0.06e 5.25 ± 0.07b 6.01 ± 0.07d
5.60 ± 0.09b 5.37 ± 0.06b 4.92 ± 0.15a 7.07 ± 0.06c 7.83 ± 0.14d
22.2 ± 3.1ab 27.2 ± 2.1bc 29.0 ± 1.9c 20.4 ± 2.27a 18.3 ± 1.50a
Values in a given column followed by the same letter are not significantly different (P = 0.05) using Tukey test.
Table 5 Metal content in sewage sludge and biochar after leaching experiment (mg L1) and after extraction with CaCl2 or DTPA (mg kg1). Sample
Cu
Ni
Cd
Pb
0.071 ± 0.005a 0.034 ± 0.001b
0.008 ± 0.001a 0.002 ± 0.001b
0.0022 ± 0.0006a 0.0013 ± 0.0004a
31.04 ± 1.21a 0.63 ± 0.12b
90.05 ± 2.13a 29.60 ± 1.27b
0.503 ± 0.175a 0.321 ± 0.112b
68.60 ± 3.78a 8.82 ± 0.07b
5.45 ± 0.42a 4.10 ± 0.31b
247.54 ± 7.60a 24.52 ± 2.33b
0.350 ± 0.107a 0.302 ± 0.101a
12.540 ± 0.107a 2.452 ± 0.072b
Metal content after leaching experiment (mg L1) L 0.300 ± 0.085a B 0.012 ± 0.006b
0.081 ± 0.004a 0.029 ± 0.015b
Metal content after extraction with DTPA (mg kg1) L 85.18 ± 4.10a B 43.55 ± 2.47b Metal content after extraction with CaCl2 (mg kg1) L 34.24 ± 4.31a B 14.8 ± 2.55b
Zn
Values in a given column followed by the same letter are not significantly different (P = 0.05) using Tukey test.
pyrolyzed (Jha et al., 2010). In addition, CEC of biochar is scarce after low pyrolysis temperatures, as the ones in our study, increasing at higher temperatures (Lehmann, 2007). Our results agreed with other studies reporting biochar to have a minimal CEC when compared to soil organic matter (Cheng et al., 2006, 2008; Lehmann, 2007). Amendments affected the value of FC in the soil, following the order T = SL4 < B4 < SL8 < B8 (see Table 4). In a similar way, the value of WP followed the order T < B4 < SL4 < B8 < SL8. On the other hand, available water (AW) was lower in the SL8 treatment compared with the soil without amendments, while biochar amendment significantly increased AW of soil. Coarse-textured soils have little water holding capacity and this allows for rapid movement of infiltrating water to depth. The use of biochar could provide a medium where there is less drainage and thus has the potential for a better plant growth. The improvement of water retention after biochar addition has previously been observed by Beck et al. (2011). However, biochar did not improve soil aggregation in our samples (Table 4) similarly to the results reported by Peng et al. (2011). Water-soluble inorganic pollutants may constitute an environmental toxicity problem when they are plant-available or when their movement through soils and potential transfer to groundwater is not impeded. Table 1 shows that metal content of biochar sample was significantly higher than metal content in sewage sludge due to ash concentration during pyrolysis. Nevertheless, plant-available metals (DTPA extraction) and the mobile forms of Cu, Ni, Zn and Pb decreased after pyrolysis process; also the risk of leaching of Cu, Ni, Zn and Cd decreased in biochar sample (Table
5). Previous works carried out by our research group have showed that metals leaching decreases after sewage sludge pyrolysis (Méndez and Gascó, 2005; Gascó et al., 2004) and in some cases no metal lixiviation was observed after pyrolysis treatment (Méndez et al., 2005). Also, Hwang et al. (2007) have shown that metals leached from sewage sludge decreased in the pyrolyzed product and Martín et al. (2003) observed that the risk of metals lixiviation in pyrolysis ashes is lower than in the incineration ones. With respect to leaching experiment in soil (Table 6), Cd and Pb contents did not vary for the different treatments considered while sewage sludge amendments increased the Cu, Ni and Zn leaching with respect to the control, especially as higher is the sewage sludge ratio. This fact is according to Gascó et al. (2005a,b) that observed an increment in Ni and Cu leaching in an acid soil amended with sewage sludge in a soil column experiment increasing the risk of harmful effects in soils with less clay content due to the influence of clay in metal retention. Indeed, Toribio and Romanyà (2006) found that concentration of metals and organic matter in the leachates depended on the soil characteristics and on the type of sewage sludge added to the soil. Basic soils with a high amount of clay showed the highest metal retention capacity, while acid soils with low clay content showed the lowest. So, soil pH and clay content can influence the metal lixiviation in soil. Also, Ahlberg et al. (2006) have showed that high metal content of sewage sludges lead to metal leaching after soil addition being one of the sewage sludge amendments restrictions. The leaching of Cu, Ni and Zn were higher in the soil treated by sewage sludge than in soil treated with biochar according to the lower leaching of metals in the biochar (Table 5). Previous works
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Table 6 Metals content (mg L1) in final solutions after leaching experiment of control and amended soils after the incubation experiment. Soil sample T SL4 SL8 B4 B8
Cu
Ni a
0.153 ± 0.008 0.268 ± 0.051b 0.345 ± 0.019c 0.142 ± 0.007a 0.155 ± 0.017a
Zn b
0.041 ± 0.008 0.062 ± 0.021bc 0.092 ± 0.015c 0.027 ± 0.005d 0.041 ± 0.002b
Cd bc
0.026 ± 0.05 0.034 ± 0.003cd 0.038 ± 0.003d 0.013 ± 0.004a 0.019 ± 0.006ab
Pb a
0.003 ± 0.002 0.004 ± 0.006a 0.005 ± 0.001a 0.002 ± 0.002a 0.002 ± 0.001a
0.0040 ± 0.0003a 0.0041 ± 0.0005a 0.0053 ± 0.0010a 0.0041 ± 0.0003a 0.0043 ± 0.0005a
Values in a given column followed by the same letter are not significantly different (P = 0.05) using Tukey test.
Table 7 Metal content (mg kg1) in samples extracted with CaCl2 or DTPA. Soil sample
Ni
Zn
Cd
Pb
T(CaCl2) SL4(CaCl2) SL8(CaCl2) B4(CaCl2) B8(CaCl2)
Cu 0.82 ± 0.05a 1.85 ± 0.05c 2.41 ± 0.10b 0.97 ± 0.10a 2.42 ± 0.09b
4.361 ± 1.930a 7.293 ± 0.260b 6.627 ± 0.297ab 5.837 ± 0.366ab 5.423 ± 0.137ab
0.419 ± 0.002a 1.160 ± 0.061c 1.257 ± 0.118c 1.213 ± 0.133c 0.790 ± 0.105b
0.100 ± 0.001a 0.493 ± 0.086bc 0.550 ± 0.044c 0.510 ± 0.115bc 0.357 ± 0.055b
0.850 ± 0.050a 4.477 ± 1.208c 3.850 ± 0.401bc 2.147 ± 0.165b 2.860 ± 0.678bc
T(DTPA) SL4(DTPA) SL8(DTPA) B4(DTPA) B8(DTPA)
4.433 ± 0.183a 7.046 ± 0.187b 7.315 ± 0.362b 6.838 ± 0.553b 7.000 ± 0.110b
0.933 ± 0.291a 5.023 ± 0.427c 4.880 ± 0.689c 3.303 ± 0.208b 2.773 ± 0.485b
0.761 ± 0.136a 2.997 ± 0.915b 3.753 ± 0.690b 0.930 ± 0.046a 1.133 ± 0.093a
0.190 ± 0.09a 0.580 ± 0.181b 0.507 ± 0.332b 0.177 ± 0.133ab 0.168 ± 0.08a
0.740 ± 0.200a 2.457 ± 0.072bc 4.233 ± 1.703c 1.193 ± 0.175ab 1.313 ± 0.471ab
Values in a given column followed by the same letter are not significantly different (P = 0.05) using Tukey test.
have proved that biochar is an effective material for reducing high concentrations of soluble metals like Cd and Zn from a contaminated soil (Beesley and Marmiroli, 2011). These authors concluded that sorption and not pH increased is the more important mechanism by which those metals are retained. In our work, the leaching of Ni and Zn decreased for lower biochar additions (B4) with respect to control soil though only high addition of biochar (B8) increased the soil pH. So, according to Beesley and Marmiroli (2011) sorption mechanism by biochar seems to be the more important factor controlling the metal leaching decrement in biochar amended soils. Finally, there were not differences in leached metals between control soil and B8 treatments which is quite interesting to control the risk of contamination of groundwater by trace metals. Table 7 shows the result for the bioavailable extracted heavy metals and for the mobile forms. In general, sewage sludge increased soil heavy metal bioavailability and mobility with respect to control soil, while biochar addition resulted in less increased heavy metal bioavailability and mobility according to both bioavailable and mobile metal content of raw materials (Table 5). Indeed, Gascó and Lobo (2007) found that sewage sludge significantly increased the level of trace elements in plants. If the highest treatments are compared, the plant bioavailability (DTPA) of Ni, Zn, Cd and Pb in B8 treatment were 57%, 30%, 29% and 31% of those in SL8. Our results agrees with Hossain et al. (2010) who measured that the contents of heavy metals in cherry tomato plants amended with sewage sludge biochar were below the Australian maximum permitted concentrations for food products. Also, Hua et al. (2009) reported a lower DTPA extracted Cd in soils amended with biochar. The results with respect to Ni are particularly important as in sewage sludge, Ni occurs in organic chelated forms that are readily available to plants and thus it can be highly phytotoxic. In addition, high Ni concentrations reduce the plant’s uptake of most other nutrients due to damaging effect on the root (Mengel, 1978). Comparing the mobile forms (CaCl2 extracted metals), there were not differences between the soils amended by sludge and biochar for Cu, Ni and Pb in the high rates. While for Zn–CaCl2 and Cd–CaCl2, the concentrations in B8 were 63% and 65% of those in SL8.
In summary, even through the amount of metals added by biochar is higher than total metal added by sewage sludge, metals seems to be retained stronger by biochar. Sewage sludge biochar can be used as a valuable soil amendment as it increases the organic matter content and can hinder the leaching of heavy metals present in raw sewage sludge. 4. Conclusions The sewage sludge pyrolysis decreased the plant-available of Cu, Ni, Zn and Pb, the mobile forms of Cu, Ni, Zn, Cd and Pb and also the risk of leaching of Cu, Ni, Zn and Cd. Sewage sludge and biochar amendments increased soil respiration with respect to the control soil. The increase was lower in the case of biochar than when sewage sludge was directly added. The leaching of Cu, Ni and Zn were lower in the soil treated with biochar that in sewage sludge treatment. Biochar amended samples also reduced plant availability of Ni, Zn, Cd and Pb when compared to sewage sludge amended samples. The amendment of the sandy loam soil with biochar from sewage sludge pyrolysis significantly increased the available water and field capacity of soil. However, only sewage sludge addition improved the percentage of soil aggregates. Acknowledgments Authors wish to thank Madrid Community-Universidad Politécnica de Madrid for the support through Project CCG10UPM/AMB-5683. References Ahlberg, G., Gustafsson, O., Wedel, P., 2006. Leaching of metals from sewage sludge during one year and their relationship to particle size. Environ. Pollut. 144, 545– 553. Beck, D.A., Johnson, G.R., Spolek, G.A., 2011. Amending greenroof soil with biochar to affect runoff water quantity and quality. Environ. Pollut. 159, 2111–2118. Beesley, L., Marmiroli, M., 2011. The immobilisation and retention of soluble arsenic, cadmium and zinc by biochar. Environ. Pollut. 159, 474–480. Bouyoucos, G.J., 1962. Hydrometer method improved for making particle size analyses of soil. Agron. J. 54, 464–465.
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