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Arsenic and chromium removal from water using biochars derived from rice husk, organic solid wastes and sewage sludge
Journal of Environmental Management 133 (2014) 309e314
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Journal of Environmental Management journal homepage: www.elsevier.com/locate/jenvman
Arsenic and chromium removal from water using biochars derived from rice husk, organic solid wastes and sewage sludge Evita Agrafioti a, Dimitrios Kalderis b, Evan Diamadopoulos a, * a b
Department of Environmental Engineering, Technical University of Crete, 73100 Chania, Greece Department of Natural Resources and the Environment, Technological Educational Institute of Crete, 73100 Chania, Greece
a r t i c l e i n f o
a b s t r a c t
Article history: Received 25 June 2013 Received in revised form 3 December 2013 Accepted 9 December 2013 Available online 8 January 2014
Biochars derived from rice husk, the organic fraction of municipal solid wastes and sewage sludge, as well as a sandy loam soil, were used as adsorbents for As(V), Cr(III) and Cr(VI) removal from aqueous solutions. The kinetic study showed that sorption can be well described by the pseudo-second order kinetic model, while simulation of sorption isotherms gave better fit for the Freundlich model. The materials examined removed more than 95% of the initial Cr(III). However, removal rates for As(V) and Cr(VI) anions were significantly lower. Biochar derived from sewage sludge was efficient in removing 89% of Cr(VI) and 53% of As(V). Its ash high Fe2O3 content may have enhanced metal adsorption via precipitation. Soil was the most effective material for the removal of As(V), yet it could not strongly retain metal anions compared to biochars, as a significant amount of the adsorbed metal was released during desorption experiments. Ó 2013 Elsevier Ltd. All rights reserved.
Keywords: Biochar Chromium Arsenic Adsorption Desorption
1. Introduction The presence of heavy metals in the environment is of global concern due to possible adverse effects on human health, as well as on aquatic flora and fauna. Heavy metals direct or secondary disposal to soils and waters poses significant environmental risks, as they are non-degraded and, in high concentrations, are also toxic. Among the heavy metals, arsenic (As) is a rather important environmental pollutant with severe carcinogenic impacts on human beings (Choong et al., 2007). The Environmental Protection Agency (EPA), as well as the European Commission guidelines set up a limit of 10 mg/L for As concentration in drinking water (Council Directive 98/83/EC, 1998; US EPA, 2009). Inorganic As exists predominately in the þ3 and þ5 oxidation states. Its presence in the environment is related not only to volcanic deposits, geothermal sources and sedimentary rocks, but also to several anthropogenic activities including pesticide manufacturing, wood preservatives production, glass industry, semiconductor production and pigmentation (Alvarado et al., 2008). Apart from As, chromium (Cr) is also widely used in various industries and is released to the environment through wastewater of industrial leather tanning, electroplating, photography, pigmentation and metal cleaning
* Corresponding author. Tel.: þ30 28210 37795. E-mail address: [email protected] (E. Diamadopoulos). 0301-4797/$ e see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.jenvman.2013.12.007
(Ucun et al., 2002). The most important Cr oxidation states in the environment are þ3 and þ6. Cr(III) is an essential trace element for human health that plays an important role in metabolic disorders (Staniek et al., 2010), reducing blood glucose and cholesterol levels while controlling diabetes. On the other hand, Cr(VI) is hazardous and toxic, having adverse effects on humans (Costa, 2003). Based on drinking water guidelines recommended by the EPA, total chromium concentration should not exceed 100 mg/l (US EPA, 2009). The corresponding limit set by the European Commission is 50 mg/l (Council Directive 98/83/EC, 1998). The conventional methods used for heavy metal removal from water include adsorption on activated carbons, precipitation, use of ion exchange resins and membrane filtration (Srivastava and Thakur, 2006). In the case of a heavy metal contaminated soil, soil removal and landfilling, physicochemical extraction, stabilization/solidification, soil washing, phytoremediation and bioremediation are usually applied (Jeyansingh and Phillip, 2005; Polti et al., 2011). However, lately there is an intense interest regarding heavy metal immobilization using biochars in waters and soils. Biochar is a carbon rich, solid by-product resulting from the pyrolysis of biomass under oxygen-free and low temperature conditions. Biochar’s proven ability to remain stable against chemical and biological degradation, when applied to soils, makes it a pioneer means of mitigating climate change (Lehmann, 2007). In addition, biochar can improve soil productivity, not only because it may be a valuable nitrogen and phosphorous source, but also it
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affects soil cation exchange capacity, pH and retention of water and nutrients (Steiner et al., 2008; Glaser et al., 2002). Finally, biochar has the potential to restore and remediate contaminated soils as it can adsorb both organic and inorganic pollutants (Beesley and Marmiroli, 2011). Biochar’s ability to adsorb heavy metals is possibly attributed to electrostatic interactions between carbon negative surface charge and metal cations, as well as to exchange of ions between biochar surface protons and metallic cations. In addition, the presence of mineral impurities (e.g. ash and metal oxides), acidic oxygen groups (e.g. carboxylic and lactonic groups) and basic nitrogen groups could further enhance the adsorption capacity of carbonaceous materials (Machida et al., 2006). The use of biochar, as a cost effective sorbent for heavy metal removal from contaminated water and soils, has already been reported by many researchers. The majority of studies are focused on the immobilization of metal cations, such as Pb, Cu, Ni and Cd (Inyang et al., 2012; Jiang et al., 2012; Liu and Zhang, 2009; Park et al., 2011; Regmi et al., 2012; Uchimiya et al., 2010), while limited research has been conducted on As(V) and Cr(VI) removal by biochar. For instance, Mohan et al. (2007) studied As(III) removal from water by woody biomass-derived biochars, showing that oak bark char has a significant potential for As(III) adsorption, whereas Beesley et al. (2011) and Gomez-Eyles et al. (2011) dealt with As immobilization in multi-element contaminated soils. In addition, Shen et al. (2012), Dong et al. (2011) and Mohan et al. (2011) studied the use of biochars for Cr(VI) removal from water and found that biochar can efficiently adsorb chromate, while its maximum sorption capacity was, in some cases, 123 mg/g. The purpose of the present study is to investigate the feasibility of using biochars for the removal of both anionic and cationic metals from water. For this reason, batch kinetic, as well as equilibrium sorption and desorption experiments were conducted using As(V), Cr(III) and Cr(VI) as adsorbates. The biochars used were derived from rice husk, the organic fraction of municipal solid wastes and sewage sludge. The first was chosen as one of the most abundant types of biomass worldwide, while the other two in order to find alternative innovative uses of these wastes. In addition, soil was also used as an adsorbent for As(V), Cr(III) and Cr(VI) removal, in order to compare its adsorption efficiency with those of biochars. Knowing the separate behavior of biochars and soil towards metal sorption, it could be the first step in explaining the fate of heavy metals in biochar-amended soil. The concentrations of heavy metals examined were rather low, in the range of mg/L, in order to simulate a moderate water contamination (based on ground/ drinking water standards). 2. Materials and methods 2.1. Feedstocks Rice husk, the organic fraction of municipal solid wastes and sewage sludge were used for biochar production. Rice husk was collected from a rice mill located in northern Greece. The organic fraction of solid wastes was collected from the Chania municipal solid waste materials recovery facility. Finally, sewage sludge was obtained from the Chania municipal wastewater treatment plant. In this plant municipal wastewater receives secondary treatment by activated sludge system, whereas sludge treatment is practiced via anaerobic digestion and belt-filter-press dewatering. This sludge sample has already been used for biochar production in a previous study (Agrafioti et al., 2013). All feedstocks were dried at 103 C for 24 h, they were then milled to <0.5 mm and stored in airtight plastic bags. The physical properties of raw samples are presented in Table 1. In addition, the inorganic constituents of the three types of biomass were determined by Energy Dispersive X-ray
Table 1 Physical properties of the three feedstocks (proximate analysis).
Rice husk Sewage sludge Solid waste a b c d e
Moisture (%)a
Ash (%)b,c
Volatile matter (%)b,d
Fixed carbon (%)b,e
10.1 84.5 58.3
17.4 25.9 32.0
81.6 73.7 65.8
1.0 0.4 2.2
On as-received basis (ASTM D2216-98). On dry basis. (ASTM E1534-93). (APHA, AWWA, WPCF, 1992). Determined by difference.
Fluorescence Spectroscopy (EDS-XRF), using a Bruker S2 Ranger XRF instrument. The soil used in the present study was collected from an agricultural area in Chania. The soil sample was dried in the sun, then sieved to 2.0 mm particle size and stored in airtight plastic bags until further use. Soil mineralogical composition and properties have already been reported (Giannis et al., 2008). It is a sandy loam soil, with pH 7.02, hydraulic conductivity 4.6 104 m/s and CEC 0.62 meq/g. 2.2. Biochar preparation Pyrolysis took place in a muffle furnace (Linn High Therm) at 300 C. To maintain an oxygen-free atmosphere during the process, nitrogen was supplied to the system at a flowrate of 200 mL/min. The temperature increase rate was set at 17 C/min. After reaching the target temperature, the sample was kept in the operating furnace for 60 min (residence time). The biochars were then removed from the furnace, cooled in a desiccator, weighted and stored in airtight plastic containers. Henceforth, the biochar produced from rice husk would be referred to as BC-RH, whereas biochars originated from sewage sludge and solid wastes would be referred to as BC-SS and BC-SW respectively. 2.3. As(V), Cr(III) and Cr(VI) kinetic and equilibrium experiments Three stock solutions of 1000 mg/L As(V), 100 mg/L Cr(III) and 100 mg/L Cr(VI) were prepared by dissolving Na2HAsO4$7H2O, Cr(NO3)3$9H2O and K2CrO4 in distilled water, respectively. These stock solutions were further diluted in distilled water in order to obtain the three working solutions. In the present study three sets of experiments were conducted: adsorption kinetic, equilibrium and desorption experiments. For the kinetic experiments, the concentrations of the three working solutions were 90 mg/L As(V), 170 mg/L Cr(III) and 160 mg/L Cr(VI). Each kinetic experiment was carried out by mixing 150 mL of the working solution with 4 g/L of each adsorbent (the three biochars and the soil) and agitating on an orbital shaker at 200 rpm at room temperature for specific time intervals. At the end of each time interval, samples were withdrawn and subsequently filtered through Whatman GF/C filters and syringe membrane filters (0.45 mm). The filtrates were then analyzed for their specific metal content with an atomic absorption spectrophotometer equipped with graphite furnace (Analytik Jena, AAS6 Vario). The adsorbed heavy metal concentration was calculated as the difference of initial and final metal concentration of the liquid phase. Equilibrium experiments took place for varying concentrations of metals (90e850 mg/L) and adsorbents (1e16 g/L). 150 mL of working solutions were mixed with each adsorbent and the mixture was agitated at 200 rpm for the contact time determined by the kinetic experiments. By the end of contact time, samples
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were filtered in the same way described in the kinetic experiments, and analyzed for their metal content. The residues (cake) on the membrane filters were used for the desorption experiments. The Whatman GF/C filters used for the adsorption experiments were placed in bottles containing 150 mL deionised water and the mixture was agitated at 200 rpm at room temperature. It was assumed that desorption equilibrium was reached at the same time as adsorption equilibrium. After reaching equilibrium time, the samples were filtered with the filters described above and the filtrate metal content was measured by means of the atomic adsorption spectrophotometer. The pH values of the solutions were measured with a pH-meter (Crison 2002), before and after adsorption as well as after desorption experiments, in order to record any significant changes. 2.4. Chromium analysis Cr(VI) is possible to reduce to Cr(III) under acid conditions and Cr(III) can be oxidized to Cr(VI) under alkaline conditions, due to Cr high redox value. Considering that the atomic adsorption spectrophotometer measures the total Cr concentration in the filtrates, in order to investigate the potential reduction or oxidation of Cr(VI) and Cr(III), respectively, a set of experiments was conducted. Each biochar (BC-RH, BC-SS, BC-SW) and the soil were mixed with 200 mg/l Cr(III) and 200 mg/l Cr(VI), separately, for 20 h. The mixture was then filtered through Whatman GF/C filters and syringe membrane filters (0.45 mm) and the filtrate was further analyzed with the Chromate Cell Test (Merck, cat no. 114552). The Chromate Cell Test is a photometric method of measuring the Cr(VI) and total Cr concentration in samples. Cr(III) concentrations were calculated as the difference between total Cr and Cr(VI).
The pore volume and BET surface area were determined from nitrogen adsorption data by means of a Thermo Scientific Surfer gas sorption analyzer. 3. Results and discussion 3.1. Kinetic studies Cr(VI) kinetic experiments resulted in close to zero metal removal for all adsorbents examined and for contact times up to 96 h, and thus the results could not be further analyzed with kinetic models. However, the kinetic sorption of As(V) and Cr(III) were analyzed with the pseudo-first (Eq. (1)) and pseudo-second order kinetic model (Eq. (2)):
logðqe qt Þ ¼ log qe t 1 t ¼ þ qt k2 q2e qe
k1 t 2:303
shown). Pseudo-second order kinetic model constants are presented in Table 2. Based on pseudo-second order kinetic model assumptions, the reaction rate is proportional to the number of active sites on the adsorbent’s surface and the rate-limiting step may be a chemical sorption between the adsorbate and the adsorbent (Mohan et al., 2011). The effect of reaction time on As(V) and Cr(III) is presented in Fig. 1. As(V) equilibrium was achieved in 24 h for BC-RH, BC-SS and BC-SW, whereas for the case of soil the equilibrium time was 72 h. Cr(III) removal was faster compared to As(V) and was achieved in less than 3 h for all adsorbents examined. In addition, biochars were more effective in adsorbing Cr(III) than As(V) from aqueous solution. The maximum Cr(III) removal achieved was >99% for the case of BC-SW, whereas the corresponding percentage for As(V) was approximately 40% for the case of soil. This could be attributed to the higher Cr(III) concentration examined but is mostly related to interactions between biochar negative surface charge and Cr(III) cations. Kinetic equilibrium times determined contact times for the subsequent sorption isotherm studies. 3.2. Sorption isotherm studies The simulation of sorption isotherms of As(V), Cr(III) and Cr(VI) on the three biochars and soil was based on the Langmuir and Freundlich models. The Langmuir model reflects the standard equilibrium process behavior assuming that the adsorbent has a constant number of adsorption sites and sorption on adsorbent surface is monolayer. It also assumes that the adsorbent surface is homogeneous and there is no interaction between adsorbed molecules. The Langmuir model is described by Equation (3):
qe ¼
2.5. Surface area analysis
(1)
(2)
where qt and qe (mg/g) are the amount of metal adsorbed at time t and equilibrium time, respectively, and k1 (1/h) and k2 (1/h) are the rate constants for the pseudo-first and pseudo-second order kinetics, respectively. Linear plot of log (qe qt) against t gives k1 and qe values for the pseudo-first order model, whereas the plot of t/qt against t gives k2 and qe for the pseudo-second order model. For all metal and adsorbents examined the pseudo-second order kinetic model gave a better fit and provided the best correlation to the data (R2 > 0.999) compared to pseudo-first order. For the latter, correlation coefficients were rather low reaching even 0.518 (data not
311
QbCe 1 þ bCe
(3)
where qe (mg/g) is the amount of metal adsorbed per unit weight of adsorbent, Ce (mg/L) is the equilibrium solution concentration of the adsorbate, Q (mg/g) is the maximum amount of adsorbed metal ions needed to form a monolayer on adsorbent surface and b (L/mg) is the Langmuir adsorption constant related to binding energies. On the other hand, the Freundlich model assumes that the adsorbent surface is heterogeneous and sorption on its surface is multilayer (Eq. (4)). 1=n
qe ¼ KCe
(4)
where qe (mg/g) is the amount of metal adsorbed per unit weight of adsorbent, Ce (mg/L) is the equilibrium solution concentration of the adsorbate, K ((mg/g)(L/mg)1/n) is a constant related to adsorbent maximum adsorption capacity and 1/n is a constant measuring the strength of adsorption. Plotting the linear forms of Langmuir and Freundlich equations, the corresponding constants and correlation coefficients were obtained (Table 3). The adsorption isotherms of As(V), Cr(III) and Table 2 Parameters of pseudo-second order kinetic model for As(V) and Cr(III) adsorption onto BC-RH, BC-SS, BC-SW and soil. Sample
Heavy metal
qe (mg/g)
k2 (1/h)
R2
BC-RH
As(V) Cr(III) As(V) Cr(III) As(V) Cr(III) As(V) Cr(III)
2.59 15.13 4.25 30.12 3.54 42.37 10.46 39.53
0.17 1.56 0.22 1.00 0.09 1.11 0.02 0.71
0.99 0.99 0.99 0.99 0.99 0.99 0.99 0.99
BC-SS BC-SW Soil
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Fig. 1. Effect of adsorption time on (a) As(V) and (b) Cr(III) removal by biochars and soil (initial As(V) concentration: 90 mg/l, initial Cr(III) concentration: 170 mg/l, adsorbent dose: 4 g/l).
Cr(VI) are presented in Fig. 2. Based on the correlation coefficients, the best-fit model for all metals and adsorbents examined was the Freundlich model. The R2 values for the Langmuir model were significantly lower and in some cases (such as in the case of As(V) adsorption on BC-RH) did not fit to experimental data at all, even though the Langmuir isotherm plot (Fig. 2) was very close to Freundlich isotherm and experimental data. The R2 values for the Freundlich model varied from 0.655 to 0.983. The failure of the Langmuir model was possibly attributed to the heterogeneous surface of the initial feedstocks and thus of the produced biochars. Besides, soil was a mixture of fine aggregates and this further supports its heterogeneity. The analysis of experimental data for Cr(III) adsorption on BC-SS could not fit either the Freundlich or the Table 3 Langmuir and Freundlich constants and correlation coefficients. Sample
Heavy metal
Langmuir model Q
BC-RH
BC-SS
BC-SW
Soil
a
As(V) Cr(III) Cr(VI) As(V) Cr(III) Cr(VI) As(V) Cr(III) Cr(VI) As(V) Cr(III) Cr(VI)
a
NF NF NF 13.42 94.34 64.10 NF NF 44.05 NF NF 17.06
Freundlich model
b
R2
K
1/n
R2
NF NF NF 0.01 0.01 0.01 NF NF 0.001 NF NF 0.002
NF NF NF 0.318 0.202 0.873 NF NF 0.564 NF NF 0.543
0.01 6 1011 5 104 0.21 5.21 1.95 0.001 NF 0.14 9 104 0.01 0.26
1.3 6.57 1.56 0.71 0.50 0.61 2.07 NF 0.75 2.52 2.95 0.54
0.983 0.894 0.842 0.747 0.700 0.914 0.940 NF 0.937 0.783 0.665 0.655
NF: No satisfactory fit by any model.
Langmuir isotherm model as even low doses of BC-SS (1 g/L) were able to adsorb approximately all the amount of the adsorbate. Based on 1/n Freundlich constants values, the use of BC-RH for the adsorption of heavy metals resulted in unfavorable isotherms (1/n > 1), BC-SS in favorable isotherms (1/n < 1), BC-SW in unfavorable isotherms for the case of As(V) and in favorable for the case of Cr(VI), while isotherms of soil were unfavorable for As(V) and Cr(III) removal and favorable for Cr(VI) adsorption. Table 4 presents the maximum removal rates of the three heavy metals for each adsorbent examined. As far as As(V) is concerned the maximum removal observed was 65% in the case of soil, followed by 55% for BC-SW, 53% for BC-SS, and 25% for BC-RH. The high adsorption efficiency of BC-SW may be attributed to the interactions of arsenate anion with the oxides in biochar solid matrix. The XRF analysis of BC-SW ash showed that the CaO content was 49.8%, whereas the Fe2O3 and Al2O3 contents were negligible (<1.5%). Considering the alkaline nature of BC-SW, As(V) solution (pH w 9.5) and the high Ca content of BC-SW, it is rather possible that the calcium oxide removed As(V) by precipitation. It has already been reported that As(V) adsorption at pH 10 is due to CaO, whereas at lower pH values is mainly attributed to Fe2O3 and Al2O3 (Diamadopoulos et al., 1993), as the pHPZC of Fe2O3 and Al2O3 has been estimated at 6.7 and 8.5 (Bayat, 2002). In addition, As(V) higher adsorption on BC-SW may be related to its higher ash content compared to the other adsorbents examined (Table 1), as carbons with high ash content have been found to be more effective in As(IV) adsorption (Lorenzen et al., 1995). BC-SS sample exhibited a similar As(V) removal efficiency to BC-SW. BC-SS ash XRF analysis showed that it contained 17.4%, 12.5 and 5.3 w/w of dry matter CaO, Fe2O3 and Al2O3, respectively. It is likely that ferric oxides enhanced As(V) adsorption onto BC-SS (the pH value of the solution varied from 6.7 to 7 during the kinetic and adsorption experiments). The low ash content of BC-RH (w17%), as well as its high ash content in SiO2 (81.3% w/w of dry matter) and low content in Ca, Fe, and Al oxides (<1.3% w/w) could explain its inefficiency in removing As(V) from the aqueous solution. However, the highest As(V) adsorption removal was observed in the case of soil. Soil ash was found to contain 5% w/w Al2O3 and <1.5% w/w CaO and Fe2O3. The low Al, Fe and Ca oxide content could not explain its efficiency in removing As(V). A possible explanation is that As(V) could be retained through ion exchange, and not precipitation, on the soil surface. As far as Cr(VI) is concerned, its potential reduction to Cr(III) was also investigated. Based on Saha and Orvig (2010), there are four potential mechanisms explaining Cr(VI) sorption by biosorbents: a) anionic adsorption, b) adsorption-coupled reduction, c) anionic and cationic adsorption and d) reduction and cationic adsorption. Our experiments with the chromate Cell Test indicated that the Cr(VI) concentration was equal to the total Cr concentration in the filtrates and thus Cr(VI) did not reduce to Cr(III). Considering that Cr(VI) reduction takes place at low pH values and the pH of the solutions used for the sorption experiments varied between 7 and 9.5, it was rather expected that Cr(VI) would not be reduced. Based on the results presented in Table 4, BC-SS was found to be the most efficient in adsorbing Cr(VI) compared to the other biochars and soil, removing approximately 89% of the total amount (10 mg/g). The adsorption may be attributed to electrostatic interactions between the Cr(VI) anion and the positively charged functional groups on the surface of BC-SS. The high Fe content of BC-SS (compared to the other biochars) may have favored Cr(VI) adsorption. BC-SW and BCRH showed the same tendency in removing Cr(VI) compared to As(V) removal, with slightly lower adsorption rates. However soil removed only 35% of the initial Cr(VI) content. This percentage was low compared to the 65% adsorption observed for the case of As(V). It has been considered that the most dominant mechanism of Cr(VI) removal is the surface reduction of Cr(VI) to Cr(III) followed
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313
Fig. 2. Isotherms of As(V), Cr(III) and Cr(VI) adsorption on BC-RH, BC-SS, BC-SW and soil.
by adsorption of Cr(III) (Saha and Orvig, 2010). However, in the current work no such reduction took place resulting in low Cr(VI) removals. Finally, Cr(III) adsorption experiments resulted in significant metal removal for the majority of adsorbents examined. Apart from BC-RH whose removal efficiency was low, the rest of biochars adsorbed more than 99% of the Cr(III), while the corresponding removal for the soil was 95%. The high retention of Cr(III) is possibly related to the electrostatic interactions between adsorbents negative surface charge and Cr(III) cations. It should be mentioned, that for all heavy metals and adsorbents examined the pH values of the solutions during the experiments did not vary considerably prior to and after adsorption (pH 1), and thus it was assumed that solution pH variation did not have a significant impact on heavy metal adsorption mechanisms. The BET surface areas of BC-RH, BC-SS and BC-SW were calculated at 155, 51 and 5 m2/g respectively, while the corresponding values of their pore volumes were 0.153, 0.0584 and 0.0291 cm3/g. The high ash content of BC-SW (32% w/w), compared to the other biochars, did not favor the development of high surface area, as ash may have filled or blocked access to the micropores in the BC-SW (Song and Guo, 2011). Nevertheless, despite the low BET surface area of BC-SW (5 m2/g), its immobilization capacity of As(V) and Cr(VI) was moderate, while its Cr(III) removal reached 99%. In addition, the high surface area (155 m2/g) BC-RH exhibited lower metal adsorption capacity, as compared to the BC-SS and BC-SW
biochars concerning As(V), Cr(III) and Cr(VI) removal. Thus, it is suggested that the main mechanisms regarding heavy metal immobilization were electrostatic interactions between biochar negative surface charge and metal cation as well as metal precipitation for the case of anionic metals. Yet, high surface area and pore volume facilitate faster mass transfer of pollutants into the biochar pores and provide more opportunities for metal-active site binding. The aforementioned results could not be directly compared to those obtained in previous studies, because there is a lack of consistency in literature data for heavy metal remediation of waters using adsorbents. This inconsistency is attributed to the different pH values, adsorbent type and dosage and temperature examined in each case. In addition, literature studies concerning As(V), Cr(VI) and Cr(III) removal by biochars and other adsorbents are limited to high initial concentrations of these metals (in the range of mg/L) (Shen et al., 2012; Mohan et al., 2011). High initial concentrations are more common in industrial wastewaters, however in case of ground and surface waters, heavy metal concentrations are much lower. Lower metal concentrations (in the range of mg/L) have been reported in activated carbon studies. For instance, Chuang et al. (2005) used oat hull derived activated carbons for the adsorption of As(V) with an initial concentration 25e200 mg/L and found that activated carbon maximum capacity was 3.09 mg As/g AC. However, the AC dosages used to achieve a specific As (V) removal were lower than the adsorbent dosages used in the present study for the same level of removal.
Table 4 Maximum As(V), Cr(III) and Cr(VI) removal (%) achieved for the lowest initial metal concentrations examined during equilibrium experiments. Heavy metal
As(V) Cr(III) Cr(VI)
Initial metal concentration (mg/L)
90 185 190
Maximum removal (%)
Optimum adsorbent dose (g/L)
BC-RH
BC-SS
BC-SW
Soil
25 42 18
53 >99 89
55 >99 44
65 95 35
BC-RH 8 2 16
BC-SS 16 12 12
BC-SW 16 2 16
Soil 16 16 16
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3.3. Heavy metal desorption Batch desorption experiments were conducted in order to investigate whether heavy metal adsorption on biochars and soil is reversible or not. The maximum desorption rates were observed for the highest initial metal concentrations examined and thus for the cases where metal removal by the adsorbents was also maximum. The results indicated that desorption varied considerably not only among the metal examined but also among the adsorbents. As far as As(V) desorption is concerned, the maximum desorption rate was observed for the case of soil, as it desorbed approximately 24% of the total As(V) amount adsorbed. On the other hand, BC-SW, which was the most efficient adsorbent in removing As(V) from the aqueous solution (its removal rate reached 98%), it desorbed only 4% of the amount adsorbed. The corresponding percentage for BC-RH and BC-SS was close to zero and 10%, respectively, while they adsorbed 56 and 80% of As(V). These findings are of great importance since they suggest that although both biochars and soil have, in some cases, close As(V) removal capacities, biochars would be more effective in strongly retaining this metal, whereas soil would release a rather significant amount. From an environmental point of view, biochar application to soils could enhance soil ability to retain heavy metals, reducing their potential release to groundwater. Similar results were also found for Cr(VI) desorption. Also in this case, the soil released close to 20% of the Cr(VI) adsorbed. The corresponding percentages for biochars were all below 3.5%. The lowest desorption was observed for Cr(III), as desorption rates were below 4% for both biochars and soil. This is possibly related to the strong electrostatic interactions between negative surface charge of the adsorbents and the Cr(III) cations.
4. Conclusions The sorption process of Cr(III), Cr(VI) and As(V) onto biochars and soil was adequately described by the pseudo-second order kinetic model and Freundlich isotherm. Biochars and soil removed Cr(III) more efficiently than the two anions. BC-SS was the only material that managed to significantly remove Cr(VI) from the aqueous solution, with a removal percentage approximately equal to 90%, possibly due to its high (compared to the other biochars) Fe2O3 content. Removal rates for the other adsorbents were below 44%. Although soil was the most efficient material for As(V) removal, desorption experiments revealed that a significant amount of the adsorbed quantity could not be retained.
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Choong, T.S.Y., Chuan, T.G., Robiah, Y., Gregory Koay, F.L., Azni, I., 2007. Arsenic toxicity. Health hazards and removal techniques from water: an overview. Desalination 217, 139e166. Chuang, C.L., Fan, M., Xu, M., Brown, R.C., Sung, S., Saha, B., Huang, C.P., 2005. Adsorption of arsenic(V) by activated carbon prepared from oat hulls. Chemosphere 61, 478e483. Costa, M., 2003. Potential hazards of hexavalent chromate in our drinking water. Toxicol. Appl. Pharmacol. 188, 1e5. Council Directive 98/83/EC, 1998. Council Directive 98/83/EC of 3 November 1998 on the quality of water intended for human consumption. Official J. Eur. Community L330, 32e54. Diamadopoulos, E., Ioannidis, S., Sakellaropoulos, G.P., 1993. As(V) removal from aqueous solutions by fly ash. Water Res. 27, 1773e1777. Dong, X., Ma, L.Q., Li, Yuncong, 2011. Characteristics and mechanisms of hexavalent chromium removal by biochar from sugar beet tailing. J. Haz. Mat. 190, 909e 915. Giannis, A., Gidarakos, E., Skouta, A., 2008. Transport of cadmium and assessment of phytotoxicity after electrokinetic remediation. J. Environ. Manage. 86, 535e544. Glaser, B., Lehman, J., Zech, W., 2002. Ameliorating physical and chemical properties of highly weathered soils in the tropics with charcoal. Biol. Fert. Soils 35, 219e 230. Gomez-Eyles, J.L., Sizmur, T., Collins, C.D., Hodson, M.E., 2011. Effects of biochar and the earthworm Eisenia fetida on the bioavailability of polycyclic aromatic hydrocarbons and potentially toxic elements. Environ. Pollut. 159, 616e622. Inyang, M., Gao, B., Yao, Y., Xue, Y., Zimmerman, A.R., Pullammanappallil, P., Cao, X., 2012. Removal of heavy metals from aqueous solutions by biochars derived from anaerobically digested biomass. Biores. Technol. 110, 50e56. Jeyansingh, J., Phillip, L., 2005. Bioremediation of chromium contaminated soil: optimization of operating parameters under laboratory conditions. J. Haz. Mat. 118, 113e120. Jiang, J., Xu, R., Jiang, T., Li, Z., 2012. Immobilization of Cu(II), Pb(II) and Cd(II) by the addition of rice straw derived biochar to a simulated polluted Ultisol. J. Haz. Mat. 229e230, 145e150. Lehmann, J., 2007. A handful of carbon. Nature 447, 143e144. Liu, Z., Zhang, F.-S., 2009. Removal of lead from water using biochars prepared from hydrothermal liquefaction of biomass. J. Haz. Mat. 167, 933e939. Lorenzen, L., Deventer, J.S.J., Landi, W.M., 1995. Factors affecting the mechanism of the adsorption of arsenic species on activated carbon. Minerals Eng. 8, 557e 569. Machida, M., Mochimaru, T., Tatsumoto, H., 2006. Lead (II) adsorption onto the graphene layer of carbonaceous materials in aqueous solutions. Carbon 44, 2681e2688. Mohan, D., Pittman, C.U., Bricka, M., Smith, F., Yancey, B., Mohammad, J., Steele, P.H., Alexandre-Franco, M.F., Gómez-Serrano, V., Gong, H., 2007. Sorption of arsenic, cadmium, and lead by biochars produced from fast pyrolysis of wood and bark during bio-oil production. J. Colloid Inter. Sci. 310, 57e73. Mohan, D., Rajput, S., Singh, V.K., Steele, P.H., Pittman Jr., C.U., 2011. Modeling and evaluation of chromium remediation from water using low cost bio-char, a green adsorbent. J. Haz. Mat. 188, 319e333. Park, J.H., Choppala, G.K., Bolan, N.S., Chung, J.W., Chuasavathi, T., 2011. Biochar reduces the bioavailability and phytotoxicity of heavy metals. Plant Soil 348, 439e451. Polti, M.A., Atjián, M.C., Amoroso, M.J., Abate, C.M., 2011. Soil chromium bioremediation: synergic activity of actinobacteria and plants. Int. Biodeterior. Biodegradation 65, 1175e1181. Regmi, P., Moscoso, J.L.G., Kumar, S., Cao, X., Mao, J., Schafran, G., 2012. Removal of copper and cadmium from aqueous solution using switchgrass biochar produced via hydrothermal carbonization process. J. Environ. Manag. 109, 61e69. Saha, B., Orvig, C., 2010. Biosorbents for hexavalent chromium elimination for industrial and municipal effluents. Coord. Chem. Rev. 254, 2959e2972. Shen, Y.-S., Wang, S.-L., Tzou, Y.-M., Yan, Y.-Y., Kuan, W.-K., 2012. Removal of hexavalent Cr by coconut coir noir and derived chars e the effect of surface functionality. Biores. Technol. 104, 165e172. Song, W., Guo, M., 2011. Quality variations of poultry litter biochar generated at different pyrolysis temperatures. J. Anal. Appl. Pyrol. 94, 138e145. Srivastava, S., Thakur, I.S., 2006. Evaluation of bioremediation and detoxification potentiality of Aspergillus niger for removal of hexavalent chromium in soil microcosm. Soil Biol. Biochem. 7, 1904e1911. Staniek, H., Krejpcio, Z., Iwanik, K., 2010. Evaluation of the acute oral toxicity class of tricentric chromium(III) propionate complex in rat. Food Chem. Toxicol. 48, 859e864. Steiner, C., Glaser, B., Teixeira, W.G., Lehmann, J., Blum, W.E.H., Zech, W., 2008. 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Journal of Environmental Management journal homepage: www.elsevier.com/locate/jenvman
Arsenic and chromium removal from water using biochars derived from rice husk, organic solid wastes and sewage sludge Evita Agrafioti a, Dimitrios Kalderis b, Evan Diamadopoulos a, * a b
Department of Environmental Engineering, Technical University of Crete, 73100 Chania, Greece Department of Natural Resources and the Environment, Technological Educational Institute of Crete, 73100 Chania, Greece
a r t i c l e i n f o
a b s t r a c t
Article history: Received 25 June 2013 Received in revised form 3 December 2013 Accepted 9 December 2013 Available online 8 January 2014
Biochars derived from rice husk, the organic fraction of municipal solid wastes and sewage sludge, as well as a sandy loam soil, were used as adsorbents for As(V), Cr(III) and Cr(VI) removal from aqueous solutions. The kinetic study showed that sorption can be well described by the pseudo-second order kinetic model, while simulation of sorption isotherms gave better fit for the Freundlich model. The materials examined removed more than 95% of the initial Cr(III). However, removal rates for As(V) and Cr(VI) anions were significantly lower. Biochar derived from sewage sludge was efficient in removing 89% of Cr(VI) and 53% of As(V). Its ash high Fe2O3 content may have enhanced metal adsorption via precipitation. Soil was the most effective material for the removal of As(V), yet it could not strongly retain metal anions compared to biochars, as a significant amount of the adsorbed metal was released during desorption experiments. Ó 2013 Elsevier Ltd. All rights reserved.
Keywords: Biochar Chromium Arsenic Adsorption Desorption
1. Introduction The presence of heavy metals in the environment is of global concern due to possible adverse effects on human health, as well as on aquatic flora and fauna. Heavy metals direct or secondary disposal to soils and waters poses significant environmental risks, as they are non-degraded and, in high concentrations, are also toxic. Among the heavy metals, arsenic (As) is a rather important environmental pollutant with severe carcinogenic impacts on human beings (Choong et al., 2007). The Environmental Protection Agency (EPA), as well as the European Commission guidelines set up a limit of 10 mg/L for As concentration in drinking water (Council Directive 98/83/EC, 1998; US EPA, 2009). Inorganic As exists predominately in the þ3 and þ5 oxidation states. Its presence in the environment is related not only to volcanic deposits, geothermal sources and sedimentary rocks, but also to several anthropogenic activities including pesticide manufacturing, wood preservatives production, glass industry, semiconductor production and pigmentation (Alvarado et al., 2008). Apart from As, chromium (Cr) is also widely used in various industries and is released to the environment through wastewater of industrial leather tanning, electroplating, photography, pigmentation and metal cleaning
* Corresponding author. Tel.: þ30 28210 37795. E-mail address: [email protected] (E. Diamadopoulos). 0301-4797/$ e see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.jenvman.2013.12.007
(Ucun et al., 2002). The most important Cr oxidation states in the environment are þ3 and þ6. Cr(III) is an essential trace element for human health that plays an important role in metabolic disorders (Staniek et al., 2010), reducing blood glucose and cholesterol levels while controlling diabetes. On the other hand, Cr(VI) is hazardous and toxic, having adverse effects on humans (Costa, 2003). Based on drinking water guidelines recommended by the EPA, total chromium concentration should not exceed 100 mg/l (US EPA, 2009). The corresponding limit set by the European Commission is 50 mg/l (Council Directive 98/83/EC, 1998). The conventional methods used for heavy metal removal from water include adsorption on activated carbons, precipitation, use of ion exchange resins and membrane filtration (Srivastava and Thakur, 2006). In the case of a heavy metal contaminated soil, soil removal and landfilling, physicochemical extraction, stabilization/solidification, soil washing, phytoremediation and bioremediation are usually applied (Jeyansingh and Phillip, 2005; Polti et al., 2011). However, lately there is an intense interest regarding heavy metal immobilization using biochars in waters and soils. Biochar is a carbon rich, solid by-product resulting from the pyrolysis of biomass under oxygen-free and low temperature conditions. Biochar’s proven ability to remain stable against chemical and biological degradation, when applied to soils, makes it a pioneer means of mitigating climate change (Lehmann, 2007). In addition, biochar can improve soil productivity, not only because it may be a valuable nitrogen and phosphorous source, but also it
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affects soil cation exchange capacity, pH and retention of water and nutrients (Steiner et al., 2008; Glaser et al., 2002). Finally, biochar has the potential to restore and remediate contaminated soils as it can adsorb both organic and inorganic pollutants (Beesley and Marmiroli, 2011). Biochar’s ability to adsorb heavy metals is possibly attributed to electrostatic interactions between carbon negative surface charge and metal cations, as well as to exchange of ions between biochar surface protons and metallic cations. In addition, the presence of mineral impurities (e.g. ash and metal oxides), acidic oxygen groups (e.g. carboxylic and lactonic groups) and basic nitrogen groups could further enhance the adsorption capacity of carbonaceous materials (Machida et al., 2006). The use of biochar, as a cost effective sorbent for heavy metal removal from contaminated water and soils, has already been reported by many researchers. The majority of studies are focused on the immobilization of metal cations, such as Pb, Cu, Ni and Cd (Inyang et al., 2012; Jiang et al., 2012; Liu and Zhang, 2009; Park et al., 2011; Regmi et al., 2012; Uchimiya et al., 2010), while limited research has been conducted on As(V) and Cr(VI) removal by biochar. For instance, Mohan et al. (2007) studied As(III) removal from water by woody biomass-derived biochars, showing that oak bark char has a significant potential for As(III) adsorption, whereas Beesley et al. (2011) and Gomez-Eyles et al. (2011) dealt with As immobilization in multi-element contaminated soils. In addition, Shen et al. (2012), Dong et al. (2011) and Mohan et al. (2011) studied the use of biochars for Cr(VI) removal from water and found that biochar can efficiently adsorb chromate, while its maximum sorption capacity was, in some cases, 123 mg/g. The purpose of the present study is to investigate the feasibility of using biochars for the removal of both anionic and cationic metals from water. For this reason, batch kinetic, as well as equilibrium sorption and desorption experiments were conducted using As(V), Cr(III) and Cr(VI) as adsorbates. The biochars used were derived from rice husk, the organic fraction of municipal solid wastes and sewage sludge. The first was chosen as one of the most abundant types of biomass worldwide, while the other two in order to find alternative innovative uses of these wastes. In addition, soil was also used as an adsorbent for As(V), Cr(III) and Cr(VI) removal, in order to compare its adsorption efficiency with those of biochars. Knowing the separate behavior of biochars and soil towards metal sorption, it could be the first step in explaining the fate of heavy metals in biochar-amended soil. The concentrations of heavy metals examined were rather low, in the range of mg/L, in order to simulate a moderate water contamination (based on ground/ drinking water standards). 2. Materials and methods 2.1. Feedstocks Rice husk, the organic fraction of municipal solid wastes and sewage sludge were used for biochar production. Rice husk was collected from a rice mill located in northern Greece. The organic fraction of solid wastes was collected from the Chania municipal solid waste materials recovery facility. Finally, sewage sludge was obtained from the Chania municipal wastewater treatment plant. In this plant municipal wastewater receives secondary treatment by activated sludge system, whereas sludge treatment is practiced via anaerobic digestion and belt-filter-press dewatering. This sludge sample has already been used for biochar production in a previous study (Agrafioti et al., 2013). All feedstocks were dried at 103 C for 24 h, they were then milled to <0.5 mm and stored in airtight plastic bags. The physical properties of raw samples are presented in Table 1. In addition, the inorganic constituents of the three types of biomass were determined by Energy Dispersive X-ray
Table 1 Physical properties of the three feedstocks (proximate analysis).
Rice husk Sewage sludge Solid waste a b c d e
Moisture (%)a
Ash (%)b,c
Volatile matter (%)b,d
Fixed carbon (%)b,e
10.1 84.5 58.3
17.4 25.9 32.0
81.6 73.7 65.8
1.0 0.4 2.2
On as-received basis (ASTM D2216-98). On dry basis. (ASTM E1534-93). (APHA, AWWA, WPCF, 1992). Determined by difference.
Fluorescence Spectroscopy (EDS-XRF), using a Bruker S2 Ranger XRF instrument. The soil used in the present study was collected from an agricultural area in Chania. The soil sample was dried in the sun, then sieved to 2.0 mm particle size and stored in airtight plastic bags until further use. Soil mineralogical composition and properties have already been reported (Giannis et al., 2008). It is a sandy loam soil, with pH 7.02, hydraulic conductivity 4.6 104 m/s and CEC 0.62 meq/g. 2.2. Biochar preparation Pyrolysis took place in a muffle furnace (Linn High Therm) at 300 C. To maintain an oxygen-free atmosphere during the process, nitrogen was supplied to the system at a flowrate of 200 mL/min. The temperature increase rate was set at 17 C/min. After reaching the target temperature, the sample was kept in the operating furnace for 60 min (residence time). The biochars were then removed from the furnace, cooled in a desiccator, weighted and stored in airtight plastic containers. Henceforth, the biochar produced from rice husk would be referred to as BC-RH, whereas biochars originated from sewage sludge and solid wastes would be referred to as BC-SS and BC-SW respectively. 2.3. As(V), Cr(III) and Cr(VI) kinetic and equilibrium experiments Three stock solutions of 1000 mg/L As(V), 100 mg/L Cr(III) and 100 mg/L Cr(VI) were prepared by dissolving Na2HAsO4$7H2O, Cr(NO3)3$9H2O and K2CrO4 in distilled water, respectively. These stock solutions were further diluted in distilled water in order to obtain the three working solutions. In the present study three sets of experiments were conducted: adsorption kinetic, equilibrium and desorption experiments. For the kinetic experiments, the concentrations of the three working solutions were 90 mg/L As(V), 170 mg/L Cr(III) and 160 mg/L Cr(VI). Each kinetic experiment was carried out by mixing 150 mL of the working solution with 4 g/L of each adsorbent (the three biochars and the soil) and agitating on an orbital shaker at 200 rpm at room temperature for specific time intervals. At the end of each time interval, samples were withdrawn and subsequently filtered through Whatman GF/C filters and syringe membrane filters (0.45 mm). The filtrates were then analyzed for their specific metal content with an atomic absorption spectrophotometer equipped with graphite furnace (Analytik Jena, AAS6 Vario). The adsorbed heavy metal concentration was calculated as the difference of initial and final metal concentration of the liquid phase. Equilibrium experiments took place for varying concentrations of metals (90e850 mg/L) and adsorbents (1e16 g/L). 150 mL of working solutions were mixed with each adsorbent and the mixture was agitated at 200 rpm for the contact time determined by the kinetic experiments. By the end of contact time, samples
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were filtered in the same way described in the kinetic experiments, and analyzed for their metal content. The residues (cake) on the membrane filters were used for the desorption experiments. The Whatman GF/C filters used for the adsorption experiments were placed in bottles containing 150 mL deionised water and the mixture was agitated at 200 rpm at room temperature. It was assumed that desorption equilibrium was reached at the same time as adsorption equilibrium. After reaching equilibrium time, the samples were filtered with the filters described above and the filtrate metal content was measured by means of the atomic adsorption spectrophotometer. The pH values of the solutions were measured with a pH-meter (Crison 2002), before and after adsorption as well as after desorption experiments, in order to record any significant changes. 2.4. Chromium analysis Cr(VI) is possible to reduce to Cr(III) under acid conditions and Cr(III) can be oxidized to Cr(VI) under alkaline conditions, due to Cr high redox value. Considering that the atomic adsorption spectrophotometer measures the total Cr concentration in the filtrates, in order to investigate the potential reduction or oxidation of Cr(VI) and Cr(III), respectively, a set of experiments was conducted. Each biochar (BC-RH, BC-SS, BC-SW) and the soil were mixed with 200 mg/l Cr(III) and 200 mg/l Cr(VI), separately, for 20 h. The mixture was then filtered through Whatman GF/C filters and syringe membrane filters (0.45 mm) and the filtrate was further analyzed with the Chromate Cell Test (Merck, cat no. 114552). The Chromate Cell Test is a photometric method of measuring the Cr(VI) and total Cr concentration in samples. Cr(III) concentrations were calculated as the difference between total Cr and Cr(VI).
The pore volume and BET surface area were determined from nitrogen adsorption data by means of a Thermo Scientific Surfer gas sorption analyzer. 3. Results and discussion 3.1. Kinetic studies Cr(VI) kinetic experiments resulted in close to zero metal removal for all adsorbents examined and for contact times up to 96 h, and thus the results could not be further analyzed with kinetic models. However, the kinetic sorption of As(V) and Cr(III) were analyzed with the pseudo-first (Eq. (1)) and pseudo-second order kinetic model (Eq. (2)):
logðqe qt Þ ¼ log qe t 1 t ¼ þ qt k2 q2e qe
k1 t 2:303
shown). Pseudo-second order kinetic model constants are presented in Table 2. Based on pseudo-second order kinetic model assumptions, the reaction rate is proportional to the number of active sites on the adsorbent’s surface and the rate-limiting step may be a chemical sorption between the adsorbate and the adsorbent (Mohan et al., 2011). The effect of reaction time on As(V) and Cr(III) is presented in Fig. 1. As(V) equilibrium was achieved in 24 h for BC-RH, BC-SS and BC-SW, whereas for the case of soil the equilibrium time was 72 h. Cr(III) removal was faster compared to As(V) and was achieved in less than 3 h for all adsorbents examined. In addition, biochars were more effective in adsorbing Cr(III) than As(V) from aqueous solution. The maximum Cr(III) removal achieved was >99% for the case of BC-SW, whereas the corresponding percentage for As(V) was approximately 40% for the case of soil. This could be attributed to the higher Cr(III) concentration examined but is mostly related to interactions between biochar negative surface charge and Cr(III) cations. Kinetic equilibrium times determined contact times for the subsequent sorption isotherm studies. 3.2. Sorption isotherm studies The simulation of sorption isotherms of As(V), Cr(III) and Cr(VI) on the three biochars and soil was based on the Langmuir and Freundlich models. The Langmuir model reflects the standard equilibrium process behavior assuming that the adsorbent has a constant number of adsorption sites and sorption on adsorbent surface is monolayer. It also assumes that the adsorbent surface is homogeneous and there is no interaction between adsorbed molecules. The Langmuir model is described by Equation (3):
qe ¼
2.5. Surface area analysis
(1)
(2)
where qt and qe (mg/g) are the amount of metal adsorbed at time t and equilibrium time, respectively, and k1 (1/h) and k2 (1/h) are the rate constants for the pseudo-first and pseudo-second order kinetics, respectively. Linear plot of log (qe qt) against t gives k1 and qe values for the pseudo-first order model, whereas the plot of t/qt against t gives k2 and qe for the pseudo-second order model. For all metal and adsorbents examined the pseudo-second order kinetic model gave a better fit and provided the best correlation to the data (R2 > 0.999) compared to pseudo-first order. For the latter, correlation coefficients were rather low reaching even 0.518 (data not
311
QbCe 1 þ bCe
(3)
where qe (mg/g) is the amount of metal adsorbed per unit weight of adsorbent, Ce (mg/L) is the equilibrium solution concentration of the adsorbate, Q (mg/g) is the maximum amount of adsorbed metal ions needed to form a monolayer on adsorbent surface and b (L/mg) is the Langmuir adsorption constant related to binding energies. On the other hand, the Freundlich model assumes that the adsorbent surface is heterogeneous and sorption on its surface is multilayer (Eq. (4)). 1=n
qe ¼ KCe
(4)
where qe (mg/g) is the amount of metal adsorbed per unit weight of adsorbent, Ce (mg/L) is the equilibrium solution concentration of the adsorbate, K ((mg/g)(L/mg)1/n) is a constant related to adsorbent maximum adsorption capacity and 1/n is a constant measuring the strength of adsorption. Plotting the linear forms of Langmuir and Freundlich equations, the corresponding constants and correlation coefficients were obtained (Table 3). The adsorption isotherms of As(V), Cr(III) and Table 2 Parameters of pseudo-second order kinetic model for As(V) and Cr(III) adsorption onto BC-RH, BC-SS, BC-SW and soil. Sample
Heavy metal
qe (mg/g)
k2 (1/h)
R2
BC-RH
As(V) Cr(III) As(V) Cr(III) As(V) Cr(III) As(V) Cr(III)
2.59 15.13 4.25 30.12 3.54 42.37 10.46 39.53
0.17 1.56 0.22 1.00 0.09 1.11 0.02 0.71
0.99 0.99 0.99 0.99 0.99 0.99 0.99 0.99
BC-SS BC-SW Soil
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Fig. 1. Effect of adsorption time on (a) As(V) and (b) Cr(III) removal by biochars and soil (initial As(V) concentration: 90 mg/l, initial Cr(III) concentration: 170 mg/l, adsorbent dose: 4 g/l).
Cr(VI) are presented in Fig. 2. Based on the correlation coefficients, the best-fit model for all metals and adsorbents examined was the Freundlich model. The R2 values for the Langmuir model were significantly lower and in some cases (such as in the case of As(V) adsorption on BC-RH) did not fit to experimental data at all, even though the Langmuir isotherm plot (Fig. 2) was very close to Freundlich isotherm and experimental data. The R2 values for the Freundlich model varied from 0.655 to 0.983. The failure of the Langmuir model was possibly attributed to the heterogeneous surface of the initial feedstocks and thus of the produced biochars. Besides, soil was a mixture of fine aggregates and this further supports its heterogeneity. The analysis of experimental data for Cr(III) adsorption on BC-SS could not fit either the Freundlich or the Table 3 Langmuir and Freundlich constants and correlation coefficients. Sample
Heavy metal
Langmuir model Q
BC-RH
BC-SS
BC-SW
Soil
a
As(V) Cr(III) Cr(VI) As(V) Cr(III) Cr(VI) As(V) Cr(III) Cr(VI) As(V) Cr(III) Cr(VI)
a
NF NF NF 13.42 94.34 64.10 NF NF 44.05 NF NF 17.06
Freundlich model
b
R2
K
1/n
R2
NF NF NF 0.01 0.01 0.01 NF NF 0.001 NF NF 0.002
NF NF NF 0.318 0.202 0.873 NF NF 0.564 NF NF 0.543
0.01 6 1011 5 104 0.21 5.21 1.95 0.001 NF 0.14 9 104 0.01 0.26
1.3 6.57 1.56 0.71 0.50 0.61 2.07 NF 0.75 2.52 2.95 0.54
0.983 0.894 0.842 0.747 0.700 0.914 0.940 NF 0.937 0.783 0.665 0.655
NF: No satisfactory fit by any model.
Langmuir isotherm model as even low doses of BC-SS (1 g/L) were able to adsorb approximately all the amount of the adsorbate. Based on 1/n Freundlich constants values, the use of BC-RH for the adsorption of heavy metals resulted in unfavorable isotherms (1/n > 1), BC-SS in favorable isotherms (1/n < 1), BC-SW in unfavorable isotherms for the case of As(V) and in favorable for the case of Cr(VI), while isotherms of soil were unfavorable for As(V) and Cr(III) removal and favorable for Cr(VI) adsorption. Table 4 presents the maximum removal rates of the three heavy metals for each adsorbent examined. As far as As(V) is concerned the maximum removal observed was 65% in the case of soil, followed by 55% for BC-SW, 53% for BC-SS, and 25% for BC-RH. The high adsorption efficiency of BC-SW may be attributed to the interactions of arsenate anion with the oxides in biochar solid matrix. The XRF analysis of BC-SW ash showed that the CaO content was 49.8%, whereas the Fe2O3 and Al2O3 contents were negligible (<1.5%). Considering the alkaline nature of BC-SW, As(V) solution (pH w 9.5) and the high Ca content of BC-SW, it is rather possible that the calcium oxide removed As(V) by precipitation. It has already been reported that As(V) adsorption at pH 10 is due to CaO, whereas at lower pH values is mainly attributed to Fe2O3 and Al2O3 (Diamadopoulos et al., 1993), as the pHPZC of Fe2O3 and Al2O3 has been estimated at 6.7 and 8.5 (Bayat, 2002). In addition, As(V) higher adsorption on BC-SW may be related to its higher ash content compared to the other adsorbents examined (Table 1), as carbons with high ash content have been found to be more effective in As(IV) adsorption (Lorenzen et al., 1995). BC-SS sample exhibited a similar As(V) removal efficiency to BC-SW. BC-SS ash XRF analysis showed that it contained 17.4%, 12.5 and 5.3 w/w of dry matter CaO, Fe2O3 and Al2O3, respectively. It is likely that ferric oxides enhanced As(V) adsorption onto BC-SS (the pH value of the solution varied from 6.7 to 7 during the kinetic and adsorption experiments). The low ash content of BC-RH (w17%), as well as its high ash content in SiO2 (81.3% w/w of dry matter) and low content in Ca, Fe, and Al oxides (<1.3% w/w) could explain its inefficiency in removing As(V) from the aqueous solution. However, the highest As(V) adsorption removal was observed in the case of soil. Soil ash was found to contain 5% w/w Al2O3 and <1.5% w/w CaO and Fe2O3. The low Al, Fe and Ca oxide content could not explain its efficiency in removing As(V). A possible explanation is that As(V) could be retained through ion exchange, and not precipitation, on the soil surface. As far as Cr(VI) is concerned, its potential reduction to Cr(III) was also investigated. Based on Saha and Orvig (2010), there are four potential mechanisms explaining Cr(VI) sorption by biosorbents: a) anionic adsorption, b) adsorption-coupled reduction, c) anionic and cationic adsorption and d) reduction and cationic adsorption. Our experiments with the chromate Cell Test indicated that the Cr(VI) concentration was equal to the total Cr concentration in the filtrates and thus Cr(VI) did not reduce to Cr(III). Considering that Cr(VI) reduction takes place at low pH values and the pH of the solutions used for the sorption experiments varied between 7 and 9.5, it was rather expected that Cr(VI) would not be reduced. Based on the results presented in Table 4, BC-SS was found to be the most efficient in adsorbing Cr(VI) compared to the other biochars and soil, removing approximately 89% of the total amount (10 mg/g). The adsorption may be attributed to electrostatic interactions between the Cr(VI) anion and the positively charged functional groups on the surface of BC-SS. The high Fe content of BC-SS (compared to the other biochars) may have favored Cr(VI) adsorption. BC-SW and BCRH showed the same tendency in removing Cr(VI) compared to As(V) removal, with slightly lower adsorption rates. However soil removed only 35% of the initial Cr(VI) content. This percentage was low compared to the 65% adsorption observed for the case of As(V). It has been considered that the most dominant mechanism of Cr(VI) removal is the surface reduction of Cr(VI) to Cr(III) followed
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313
Fig. 2. Isotherms of As(V), Cr(III) and Cr(VI) adsorption on BC-RH, BC-SS, BC-SW and soil.
by adsorption of Cr(III) (Saha and Orvig, 2010). However, in the current work no such reduction took place resulting in low Cr(VI) removals. Finally, Cr(III) adsorption experiments resulted in significant metal removal for the majority of adsorbents examined. Apart from BC-RH whose removal efficiency was low, the rest of biochars adsorbed more than 99% of the Cr(III), while the corresponding removal for the soil was 95%. The high retention of Cr(III) is possibly related to the electrostatic interactions between adsorbents negative surface charge and Cr(III) cations. It should be mentioned, that for all heavy metals and adsorbents examined the pH values of the solutions during the experiments did not vary considerably prior to and after adsorption (pH 1), and thus it was assumed that solution pH variation did not have a significant impact on heavy metal adsorption mechanisms. The BET surface areas of BC-RH, BC-SS and BC-SW were calculated at 155, 51 and 5 m2/g respectively, while the corresponding values of their pore volumes were 0.153, 0.0584 and 0.0291 cm3/g. The high ash content of BC-SW (32% w/w), compared to the other biochars, did not favor the development of high surface area, as ash may have filled or blocked access to the micropores in the BC-SW (Song and Guo, 2011). Nevertheless, despite the low BET surface area of BC-SW (5 m2/g), its immobilization capacity of As(V) and Cr(VI) was moderate, while its Cr(III) removal reached 99%. In addition, the high surface area (155 m2/g) BC-RH exhibited lower metal adsorption capacity, as compared to the BC-SS and BC-SW
biochars concerning As(V), Cr(III) and Cr(VI) removal. Thus, it is suggested that the main mechanisms regarding heavy metal immobilization were electrostatic interactions between biochar negative surface charge and metal cation as well as metal precipitation for the case of anionic metals. Yet, high surface area and pore volume facilitate faster mass transfer of pollutants into the biochar pores and provide more opportunities for metal-active site binding. The aforementioned results could not be directly compared to those obtained in previous studies, because there is a lack of consistency in literature data for heavy metal remediation of waters using adsorbents. This inconsistency is attributed to the different pH values, adsorbent type and dosage and temperature examined in each case. In addition, literature studies concerning As(V), Cr(VI) and Cr(III) removal by biochars and other adsorbents are limited to high initial concentrations of these metals (in the range of mg/L) (Shen et al., 2012; Mohan et al., 2011). High initial concentrations are more common in industrial wastewaters, however in case of ground and surface waters, heavy metal concentrations are much lower. Lower metal concentrations (in the range of mg/L) have been reported in activated carbon studies. For instance, Chuang et al. (2005) used oat hull derived activated carbons for the adsorption of As(V) with an initial concentration 25e200 mg/L and found that activated carbon maximum capacity was 3.09 mg As/g AC. However, the AC dosages used to achieve a specific As (V) removal were lower than the adsorbent dosages used in the present study for the same level of removal.
Table 4 Maximum As(V), Cr(III) and Cr(VI) removal (%) achieved for the lowest initial metal concentrations examined during equilibrium experiments. Heavy metal
As(V) Cr(III) Cr(VI)
Initial metal concentration (mg/L)
90 185 190
Maximum removal (%)
Optimum adsorbent dose (g/L)
BC-RH
BC-SS
BC-SW
Soil
25 42 18
53 >99 89
55 >99 44
65 95 35
BC-RH 8 2 16
BC-SS 16 12 12
BC-SW 16 2 16
Soil 16 16 16
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3.3. Heavy metal desorption Batch desorption experiments were conducted in order to investigate whether heavy metal adsorption on biochars and soil is reversible or not. The maximum desorption rates were observed for the highest initial metal concentrations examined and thus for the cases where metal removal by the adsorbents was also maximum. The results indicated that desorption varied considerably not only among the metal examined but also among the adsorbents. As far as As(V) desorption is concerned, the maximum desorption rate was observed for the case of soil, as it desorbed approximately 24% of the total As(V) amount adsorbed. On the other hand, BC-SW, which was the most efficient adsorbent in removing As(V) from the aqueous solution (its removal rate reached 98%), it desorbed only 4% of the amount adsorbed. The corresponding percentage for BC-RH and BC-SS was close to zero and 10%, respectively, while they adsorbed 56 and 80% of As(V). These findings are of great importance since they suggest that although both biochars and soil have, in some cases, close As(V) removal capacities, biochars would be more effective in strongly retaining this metal, whereas soil would release a rather significant amount. From an environmental point of view, biochar application to soils could enhance soil ability to retain heavy metals, reducing their potential release to groundwater. Similar results were also found for Cr(VI) desorption. Also in this case, the soil released close to 20% of the Cr(VI) adsorbed. The corresponding percentages for biochars were all below 3.5%. The lowest desorption was observed for Cr(III), as desorption rates were below 4% for both biochars and soil. This is possibly related to the strong electrostatic interactions between negative surface charge of the adsorbents and the Cr(III) cations.
4. Conclusions The sorption process of Cr(III), Cr(VI) and As(V) onto biochars and soil was adequately described by the pseudo-second order kinetic model and Freundlich isotherm. Biochars and soil removed Cr(III) more efficiently than the two anions. BC-SS was the only material that managed to significantly remove Cr(VI) from the aqueous solution, with a removal percentage approximately equal to 90%, possibly due to its high (compared to the other biochars) Fe2O3 content. Removal rates for the other adsorbents were below 44%. Although soil was the most efficient material for As(V) removal, desorption experiments revealed that a significant amount of the adsorbed quantity could not be retained.
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