Conversion of solid waste to activated carbon to improve landfill sustainability

Abstract

Heterogeneous composite wastes from landfills were evaluated as precursors for the generation of activated carbon (AC). A single-step chemical activation process was applied involving irradiation with microwave energy and impregnation with KOH. The average percentage yield of AC from active landfill precursor was higher than that from closed landfill for all depths sampled. Increase in impregnation ratio and irradiation power decreased the average percentage yield for both landfill precursors (active: 38.1 to 33.1%; closed: 42.1 to 33.3%). The optimum pH range for adsorption of methylene blue was pH 6–7, while adsorption increased with increase in temperature over the range 30 to 50°C. Carbonyl and hydroxyl groups were the major functional groups on the surface of AC. The properties of the AC are potentially suitable for the removal of cationic dyes and pollutants. AC generated from the landfill composite was comparable to that from other biomass being managed through AC generation. This is the first report to demonstrate the possible reuse of landfill composite as AC. The reuse option of landfill composite could provide a means of sustainable management of landfilled municipal waste.

Keywords

Adsorption capacity activated carbon landfill composites microwave heating

Introduction

Landfill remains a major means of solid waste management despite concerted efforts at reducing over-reliance on this method of final waste disposal. There are hundreds of thousands of active, closed or abandoned landfills worldwide, with around 100,000 in the USA, >150,000 in Europe and an overwhelming majority in developing countries (Butt et al., 2014). Sustainable management of the huge tonnage of landfill waste to prevent immediate and future environmental pollution remains a significant and pressing challenge. Reuse and recovery of landfill composite via enhanced mining of landfill sites has been suggested as an alternative option to reduce negative environmental impact (Jain et al., 2014; Kaartinen et al., 2013). Typically, 50–60% of excavated landfill wastes are fine and intermediate components with limited reuse options (Jain et al., 2014).

Recently, biomass wastes, which are major constituents of some landfills, have been identified as potential precursors for the production of activated carbon (AC) (Li et al., 2008; Alslaibi et al., 2014; Septhupathi et al., 2015). A wide range of solid waste has been shown to be suitable for the production of AC, including plants (Islama et al., 2017; Tang et al., 2012), wood and sawdust (Ali and Mohd. Khan, 2012; Foo and Hameed, 2012a) and industrial sludge (Fu et al., 2013). Synthetic materials and papers have also been used as precursors for the production of mesoporous activated carbon (Nahil and Williams, 2012; Song et al., 2016). Most developing countries have a high percentage of organic matter in their waste stream, ranging from 40% to 85% of the total waste, which is disposed mainly in landfills (Hoornweg and Bhada-Tata, 2012). The International Panel on Climate Change (IPPC, 2006) estimates that organic waste constitutes 89–92% of waste generated in sub-Saharan Africa. Active and old (closed) landfills in Africa could therefore be a reservoir of potential precursor for AC. To the best of our knowledge, there are no published reports on the potential use of landfill composite waste as precursor for the production of AC.

The aim of the work presented here was to assess active and closed landfill composites as precursors for the production of AC in a single-step chemical activation process involving KOH and microwave heating. A further aim was to evaluate the quality of AC produced with regard to the particle size and depth of precursor used, its chemical functionalities and its adsorption mechanism for methylene blue (MB).

Materials and methods

Sampling sites

Active and closed landfill sites (Olusosun and Abule Egba landfills, respectively) located in the western and northern areas of Lagos, Nigeria, were used for this study. Both landfills are in proximity to major commercial centres of the state. The closed landfill had received waste for 25 years and was closed in 2009. The active landfill has been in operation since 1992. Waste is disposed at both landfills with no pre-disposal treatment. Manual sorting of recyclable wastes is carried out at both landfill sites, while organic or non-recyclable wastes are left to decompose.

Sampling

The landfill sites were systematically gridded into seven cells according to Resource Conservation and Recovery Act (RCRA) U.S. Environmental Protection Agency (USEPA) guidelines (USEPA, 2002). A total of three samples were obtained from the top layer of each of the cell at upper (0–15 cm), mid (16–35 cm) and lower (36–50 cm) depths. Approximately 500 g of sample was collected from each sampling point and stored in disinfected plastic containers.

Samples were dried in an oven at 105°C to a constant weight. The dried samples were separated by size and physical properties into composites of less degraded and more degraded components (more degraded <6.3mm, less degraded >6.3mm). Each composite sample was crushed and homogenized using a ball-mill to obtain representative samples excluding metals. The composites were identified by landfill type, size and depth: active landfill (A); closed landfill (C); more degraded (MD); less degraded (LD); upper depth (upper); mid depth (mid); lower depth (lower).

Sampling profile

Sampling for this research was designed to evaluate the suitability of the first receptor layer (between 5 and 30 cm) of the landfills, which reflects the early changes in the composition of the landfill waste. A shallow landfill sampling covering the whole expanse of the landfill was used to reveal the spatial-temporal nature of waste components within this landfill layer. A deeper in situ sampling was not considered, because the intended use of the landfill component is for an ex situ conversion process, which will be relatively unaffected by the immediate in situ conditions at the landfill. The IPCC recommends the evaluation of the first-layer landfill components for countries that lack data on the types and properties of solid waste before disposal, like the sampled sites. A linear trend in the degraded components of waste and the sampling depth was observed in both active and closed landfills: with increasing depth, the degraded component progressively decreased from 55.8% (w/w) to 43.5% (w/w) for active landfill, while the closed landfill degraded component increased from 58.6% (w/w) to 82.9% (w/w). The degraded component was higher in the closed landfill compared with the active landfill across the depth. The less degraded component had no definite trend with depth. A detailed waste component characterization across the sampling depth has previously been reported by Adelopo et al. (2017a).

Precursor characterization

The elemental and chemical characterization of the precursors had been carried out using scanning electron microscope/energy-dispersive x-ray spectroscopy (SEM/EDX), Fourier transform-infrared (FTIR) spectroscopy and proximate analysis, as reported in Adelopo et al. (2017b).

Microwave modification

A domestic microwave (LG Intellowave, model no. MB-382W/03, LG electronics Australia, New South Wales, Australia) was modified for the activation process. The oven had a frequency rating of 2.5 GHz, regulators to control the exposure time between 1 and 60 min, and power wattage settings of 90, 180, 360, 600 and 800W. A quartz tube with internal diameter of 4cm and length 28cm was run vertically through the microwave area, with gas inlet and outlet at each end to pass gas from the bottom of the microwave to the top outlet duct.

Activation preparation

Samples were further homogenized using a mortar and pestle to an average particle size of 0.25mm. An aliquot (5g) of sample was impregnated with 5g of KOH dissolved in 10ml of distilled water. The mixture was agitated using a stirrer at 80 rpm for 1 hour and then dried in an oven at 105°C for 24 hours to attain a constant weight. An aliquot (2g) of the mixture was loaded into a sample holder (combustion tube) in the reactor fixed within the microwave cavity. Pure nitrogen gas (99.5%) was passed through the microwave cavity at a flow rate of 15cm³/sec for 1 minute to purge the oxygen, and then the sample was irradiated for 10 minutes under the stream of nitrogen gas. The sample was allowed to cool under nitrogen gas flow. Power input and radiation time were the major process parameters, which ranged from 600 to 800W and time kept at 10 minutes during irradiation.

The activated samples were washed with 5% HCl and then continuously washed with distilled water until the pH was within a neutral range (6–7). The activated sample was then dried in the oven at 105°C to a constant weight. The percentage yield was determined as

% yield = \frac{(\begin{array}{l} Weight of precursor \\ before activation - \\ weight of precursor after activation \end{array}) \times 100}{Weight of precursor before activation}

(1)

The reproducibility of yield was evaluated through duplicate activation of the precursor and the relative percentage difference (RPD) was determined.

Adsorption by AC

Assessment of the adsorption capacity of AC samples was carried out using MB. A 25mg/l solution of MB in distilled water was prepared from a 100mg/l stock solution. An aliquot (10ml) of the 25mg/l MB solution was added to 10mg of activated carbon. The mixture was agitated in a shaking water bath (Brunswick C76, New Brunswick Scientific, New Jersey, USA) at 200 rpm for 2 h at 23°C. Supernatant solution (2ml) was removed by pipette after the solution had been allowed to settle, and MB concentration in the supernatant was determined by measuring absorbance at 664nm. The amount of adsorbed MB at equilibrium, q_e (mg/g), was calculated by

q_{e} = \frac{(C_{o} - C_{e}) V}{W}

(2)

where C_o and C_e (mg/l) are the initial and final concentration of MB (mg/l) before and after equilibrium, respectively. V is the volume of the solution (l), and W is the mass of dry adsorbent used (g).

The experimental procedure described above was modified to investigate the effect of adsorbent dosage (10, 25 and 50mg), pH and temperature. The pH effect on adsorption capacity was evaluated by adjusting the pH of the solution to 2–3 or 11–12 through the addition of 5% HCl or 0.1M KOH, respectively. The effect of temperature on adsorption capacity was determined by varying the temperature in the Labline orbit environ shaker (model 3527, Lab-line, New York, USA), to 23°C, 30°C and 50°C. The adsorption capacity of the activated carbon at any other time (q_t) during the analysis was determined according to equation (2).

Characterization of AC

Morphologies of AC samples were evaluated by scanning electron microscopy (SEM) using a Carl Zeiss EVO HD15 instrument set at an accelerating voltage of 20 keV. FTIR spectroscopy was run on a Bruker Alpha Attenuated Total Reflection-FTIR spectrophotometer with a frequency range of 4000–400 cm⁻¹. Background correction of the spectrum was carried out prior to every measurement. The surface area and pore structure parameters of AC were obtained by nitrogen adsorption at 77.5 K using a Micromeritics Gemini 2365 surface area analyser, Micromeritics Instrument Corporation, Georgia, USA.

Results and discussion

Activation conditions, percentage yield

Table 1 presents the activation conditions for each precursor, together with the associated yield and reproducibility for the AC produced. Reproducibility was assessed to determine whether the heterogeneous nature of the precursor affected the AC yield and its adsorption properties. At the same activation condition of 600 W and impregnation ratio of 1:1, the average yields of samples from active landfill were higher than from closed landfill for all sampling depths (upper, 23.8 and 19.3%; mid, 52.4 and 34.7%; lower, 35.7 and 27.0%, respectively). More degraded samples from the active landfill showed better reproducibility of percentage yield (up to 18% RPD) compared with more degraded samples from the closed landfill (up to 30% RPD).

Table 1.

AC yield, reproducibility and adsorption capacity with depth.

Samples			Reproducibility
Sample names		% yield^a	RPD % yield of samples 1 and 2	% uptake of MB	Adsorption capacity^b (mg/g)	pH
AMD upper	1	23.8		87	43.6	6.2
AMD upper	2	23.8	0	84.5	42.4	6.3
AMD mid	1	57.1		66.5	33.3	6.4
AMD mid	2	47.6	18	82.5	41.3	6.5
AMD lower	1	47.6		86	43	6.7
AMD lower	2	23.8	17	87	43.4	7.1
CMD upper	1	30.8		77	38.3	6.7
CMD upper	2	7.7	30	59	29.5	6.8
CMD mid	1	30.8		76.5	38.4	6.7
CMD mid	2	38.5	5	83.5	41.9	6.9
CMD lower	1	38.5		87.5	43.6	6.2
CMD lower	2	15.4	21	88	43.9	6.5
ALD upper		57.1		86	43	6.4
ALD mid		47.6		65	32	6.4
ALD lower		47.6		32	16	6.4
CLD upper		34.6		87	43	6.4
CLD mid		15.4		51	26	6.9
CLD lower		23.8		78	39	6.8

Activation conditions: power, 600W; impregnation ratio, 1:1; duration, 10min.

Adsorption capacity at 23°C, absorbent weight = 50mg.

RPD: Relative percentage difference.

For both active and closed landfills, the percentage yield did not show a definitive trend with depth. The pH of samples from both landfills was within the same narrow range, 6.2–7.1. These data indicated that under the same set of activation parameters, the percentage yield of AC from the heterogeneous precursor used (landfill composite) is fairly reproducible. A similar compositional trend of the precursor may have influenced the reproducibility.

Effect of irradiation power

The closed landfill yield of AC progressively decreased from 42.9 to 33.3% with increased irradiation power, while the active landfill precursors also had the lowest yield at the highest irradiation power (Table 2). The relatively low yields at 800 W may be attributed to increased gasification of precursor. During the activation process at 800 W, intense build-up of gaseous components within the reaction cavity was observed within 1–5 s of exposure to microwave radiation. The process of precursor degradation, volatilization and decomposition is known to increase with increasing microwave power (Foo and Hameed, 2012b).

Table 2.

AC yield with variations in wattage and impregnation ratio.

Samples	Ma/Ms	Microwave power (W)	% yield	pH
ALD upper	1:1	360	38.1	6.6
ALD upper	1:1	600	57.1	6.4
ALD upper	1:1	800	33.3	6.4
CLD upper	1:1	360	42.9	6.5
CLD upper	1:1	600	34.6	6.4
CLD upper	1:1	800	33.3	6.3
ALD upper	0.5:1	^a	47.6^a	6.2
ALD upper	1:1		57.1	6.4
ALD upper	2:1		33.3	6.4
ALD upper	0.5:1		14.3	6.4
CLD upper	0.5:1		28.6	6.3
CLD upper	1:1		42.9	6.4
CLD upper	2:1		47.6	6.2
CLD upper	0.5:1		28.6	6.5
CMD upper	0.5:1		38.1	6.3
CMD upper	1:1		19	6.7
AMD upper	0.5:1		42.9	6.2
AMD upper	1:1		23.8	6.2

Activation conditions: power, 600W; impregnation ratio, 1:1; duration, 10min.

A: active landfill; C: closed landfill; LD: less degraded; lower: lower depth; MD: more degraded; mid: mid depth; upper: upper depth.

Activating agent impregnation ratio is an important parameter in a chemical activation process, which influences both quantity and quality of the AC produced (Ferrera-Lorenzo et al., 2014; Njoku et al., 2014). A ratio of 0.5:1 was observed to be more suitable than 1:1 for more degraded sample, with yield enhancement from 19 to 38% and from 24 to 43% for the closed and active landfill, respectively (Table 2). This suggests that degraded precursor with an increased surface area absorbed more activating agent. Conversely, for less degraded samples from both landfills, an increase in impregnation ratio from 0.5:1 to 1:1 led to an increase in percentage AC yield (active 47.6 to 57.1%, closed 28.6 to 42.9%) (Table 2). Similarly, Foo and Hameed (2012b) observed that increasing the ratio of activation chemical reagent from 0.25 to 1.25 for mangosteen peel precursor increased the yield of AC from 76.03 to 88.01%.

Characterization of AC

Scanning electron microscopy (SEM)

Typical SEM micrographs of precursors and obtained AC revealed distinct differences in surface area caused by microwave activation, with well-defined porosity distributed across the surface area of AC (Figure 1). The relationship between the nature of the pore formation and adsorption capacity could not be inferred from the microphotographs. An adsorption study on the AC was therefore carried out using MB.

Adsorption studies

Effect of adsorbent quantity

Figure 2 shows the absorption capacity at 23°C of AC produced from precursor samples of each landfill. The adsorption capacities of the AC generally decreased with increase in weight of AC; 36–190 mg/g for 10 mg, 59–82mg/g for 25mg and 38–45 mg/g for 50mg of the adsorbent. The decrease in adsorption capacity of AC could be due to the splitting effect in the flux (concentration gradient) between the adsorbate and the adsorbent (Beekaroo and Mudhoo, 2011). As the quantity of adsorbent increased from 10g to 50mg with MB concentration kept constant at 25mg/l, there would be an increase in the number of surface sites of adsorbent available for the adsorbate adhesion, leading to a lower number of adsorbate molecules per site.

Adsorption capacity of AC with depth of precursor sampling

The adsorption capacity of AC produced from less degraded active landfill decreased with depth (upper, 43mg/g; mid, 32mg/g; lower, 16mg/g), while that of AC from the more degraded precursor of the closed landfill increases with depth (upper, 35mg/g; mid, 39mg/g; lower, 42mg/g) at 23°C using 50mg absorbent (Figure 2). This trend may reflect the elemental content of carbon in the precursor. The adsorption capability of 9 out of 12 of the AC samples followed the same trend as reported for elemental content of carbon in the elemental characterization studies of the precursors by Adelopo et al. (2017b), that is, active degraded, upper > mid > lower; closed degraded, lower > upper > mid. For both landfills, the average adsorption capacity of AC from more degraded precursors was higher than that of AC from less degraded precursors (active landfill 42 and 30mg/g, respectively; closed landfill 41 and 36mg/g, respectively). This may be due to higher fixed carbon content in the more degraded samples than in the less degraded samples; the fixed carbon of a biomass represents the carbon content available for fuel and energy conversion (García et al., 2013).

Table 3 compares the optimum absorption capacity of the landfill precursor AC with AC generated from other biogenic waste. The optimum adsorption capacity of AC generated from both landfills using MB was higher than values reported for AC from oil palm and tea waste.

Table 3.

Comparison of the maximum adsorption capacities of MB onto different biogenic waste adsorbents.

Adsorbents		Adsorption capacity (mg/g)	Reference
Coconut husk based activated carbon		66	Prauchner and Rodríguez-Reinoso (2012)
Tea waste		85.16	Uddin et al. (2009)
Activated carbon from oil palm wood		90.9	Tamai et al. (1996)
Black lignin liquor		92.51	Fu et al. (2013)
Mangosteen peel		345	Foo and Hameed (2012b)
This study:
Closed landfill composite activated carbon	More degraded	152
	Less degraded	101
Active landfill composite activated carbon	More degraded	190
	Less degraded	157

Temperature effect on adsorption

There was an increase in the percentage uptake of MB with increase in temperature from 30 to 50°C for all AC samples tested (Figure 3). However, only 60% of the AC samples indicated an increase in the percentage uptake of MB with increase in temperature from 23 to 30°C. The AC from more degraded precursors of active landfill (upper and lower layers) showed higher uptake of MB at 23°C than at 30°C. Statistical analysis of variance (ANOVA) was used to investigate whether there was a significant difference in percentage MB uptake at each of the three temperatures: 23, 30 and 50°C. Taking landfill type (closed and active) and sample nature (more degraded and less degraded) as fixed factors, ANOVA analysis showed that there was no significant statistical difference (p >.2) in percentage MB uptake at each of the temperatures.

pH effect on adsorption

The percentage uptake of MB onto the AC was affected by solution pH. For all the ACs investigated (six samples), uptake of MB was found to be optimal at a solution pH within the range 6–7 (Figure 4). This is similar to the observation made by Gerçel et al. (2007) and Karago et al. (2008), who observed an optimum adsorption of MB at the pH of 6 for AC produced from Euphorbia rigida and sunflower oil cake, respectively. Solution pH and ionic strength are known to be major factors influencing the adsorption process in solution (Foo and Hameed, 2012; Njoku et al., 2014). Solution pH effect on adsorption capacity is controlled by the electrostatic interactions between charged surface of adsorbents and the adsorbates present in the solution (Dogan et al., 2006; Foo and Hameed, 2012b; Karago et al., 2008; Moreno-Castilla, 2004).

In acidic medium, the AC would be neutral with –COOH groups, while the MB would carry a positive charge and thus compete with H⁺ for AC sites. At neutral pH, the AC would become more negatively charged, and consequently, MB cation uptake on the adsorption site would increase. In the basic medium, the drop in adsorption capacity could have been due to electrostatic repulsion between OH⁻ in the medium and the –COO⁻ surface functional group of the adsorbent. At pH 11–12, the adsorption capacity of AC of more degraded active landfill precursor progressively decreased with depth from 114 mg/g (upper) to 106mg/g (mid) to 73mg/g (lower) (Figure 4). There was no clear trend in MB adsorption with pH in relation to depth of sampling for other AC samples.

A total of 42 FTIR spectra of AC samples were analysed to investigate the nature of functional groups common to the ACs. Major peaks located at 3320–3380, 1635–1694, 990–998, 770–779 and 699cm⁻¹ were prominent in most (8 of 12) of the AC samples. The peak at 3320–3380cm⁻¹ was identified as stretching vibration of the hydroxyl group from alcohol or phenol (Ferrera-Lorenzo et al., 2014; Karago et al., 2008). The carboxyl group represented by the 1635–1694cm⁻¹ peak was attributed to C=C stretching vibration of olefin groups (Karago et al., 2008; Tran et al., 2017), while the peaks at 990–998, 779 and 699cm⁻¹ were ascribed to C–H out of plane bending of aliphatic groups (Karago et al., 2008; Suhas et al., 2007).

The FTIR spectra of the ACs of both landfills showed significant changes in functional groups when compared with spectra of their respective precursors (Figure 5).

Prominent peaks at 3690 and 3619 cm⁻¹ for the active landfill precursor, attributed to O–H vibration of clayey materials (possibly Si-OH), was absent in all AC samples of the landfill (Adelopo et al., 2017b). This could be attributed to dehydroxylation of the OH group during activation. OH vibration of clayey material is known to become less stable with increase in temperature, and the hydroxyl group can be oxidized to a carboxylic or aldehyde group (Suhas et al., 2007).

The intense peaks at around 1030–1033 and 1000–1009 cm⁻¹ attributed to silica/clay minerals (Si–O–Si and Si–O stretching vibration), which were present in the precursors of both landfills, are completely absent in ACs of lower and mid layers of the active landfill, but not for the upper layer (Adelopo et al., 2017b). In most (60%) of the closed landfill ACs, however, these peaks were retained but were less intense. This suggests that the source of these peaks in AC differs for each landfill. The AC peaks for active landfill sample could be mainly from clay mineral (Si–OH), which is quite soluble in alkali solvent and less resistant to heat, while those of the AC from closed landfill could be largely due to Si–O–Si stretching of silica, which is more stable to heat. A very weak aliphatic methyl peak at 2980–2984 cm⁻¹ was found in 95% of AC samples from the closed landfill, but was absent in most ACs of the active landfill. Other peaks at 1558–1560 cm⁻¹ were ascribed to C–O groups conjugated with aromatic rings (Foo et al., 2013). The peaks at 1440–1412cm⁻¹ were attributed to C–O–H in-plane bending of carboxylic carbon (Suhas et al., 2007), while those at 1340–1395cm⁻¹ were ascribed to conjugated moieties of the oxygen functional group of C=O stretching and C–O stretching in carboxylic groups (Ji et al., 2007). The peaks located at 1165 and 874cm⁻¹, attributed to C–O and C–H vibrations (Liou, 2010), were found in AC from both landfill types but were more prominent in the closed landfill than the active landfill.

Brunauer–Emmett–Teller (BET) surface analysis

For both closed and active landfills, the surface area of AC from more degraded precursor samples increased with increase in depth of sampling (closed landfill, 72.53 to 132.51 m²/g; active landfill, 34.02 to 105.15 m²/g), whereas that from less degraded precursors had no definite depth relationship (see Table 4). The surface area of AC from more degraded closed landfill samples followed the same trend as the adsorption capacity of MB, that is, an increase with increase in depth of landfill sampling.

Table 4.

BET determined surface and pore surface areas of AC.

Sample	BET surface area (m²/g)	Pore surface area (m²/g)
CMD upper	72.53	5.8
CMD mid	86.07	3.85
CMD lower	132.51	5.19
AMD upper	34.02	7.49
AMD mid	38.08	1.38
AMD lower	105.15	ND
ALD upper	51.64	0.87
ALD lower	92.89	11.33
CLD upper	31.75	2.58
CLD mid	89.33	4.59

ND: not determined.

A further study of the pore distribution in three selected samples revealed the distribution of pore sizes, with most pore diameters of the AC within the range 2–5nm. This indicates that the ACs were principally mesoporous. The ACs had average pore volumes of 0.160, 0.126 and 0.102 cm³.

N₂ adsorption isotherms for AC samples are shown in Figure 6. The plots show that N₂ adsorption follows type II (Figures 6(a) and (b)) and type III (Figure 6(c)) International Union of Pure and Applied Chemistry (IUPAC) adsorption.

Type II and III adsorption processes conform to the Freundlich adsorption model of unrestricted multilayer adsorption with the adsorbate–adsorbate interaction playing the major role (Sing et al., 1985). The hysteresis has a type H4 loop, which is attributed to narrow slit-like pores (Sing et al., 1985). The large Langmuir areas of AC (636, 466.1 and 361.8 m²/g) further strengthen the adsorption potential of the AC. Hu et al. (2001) observed adsorption isotherms of types I and II for the AC from coconut shells as the ratio of activating agent impregnation increased.

Effect of degradation

A similar trend was observed in the properties of ACs of more degraded precursors of both landfills compared with their less degraded precursors. AC from the more degraded samples had higher adsorption capacity and BET surface area than that from the less degraded samples (see Tables 1 and 4). Also, at equilibrium, the isothermal adsorption of MB onto AC for all degraded precursors conformed to Freundlich’s model regardless of the landfill type. This could indicate the effect of degradation between the precursors used (less degraded vs. more degraded) and the nature of the constituent waste.

Process scale-up potential

The potential scale up of the activation process from landfill composite could be conceptually considered a possibility with the robust reviews of economic, social and technological enhancement in landfill mining provided by previous research. Frändegård et al. (2013), Danthurebandara et al. (2015) and Zhou et al. (2015) identified and evaluated cost-effective technologies preferred for harnessing the landfilled component, which is the first challenge in landfill AC generation. According to Di Maria et al. (2013), a mechanical sorter with two- or three-dimensional outlets and near infrared selector scan can separate an average of 71,000 tonnes of landfilled municipal waste per year with 98% ferrous recovery potential. The screening process of the landfill composite waste for AC could be modified in line with this preferred technology. The cost point here is inherent in all processes of landfill mining for ex situ purposes.

Crushing and homogenizing the precursor is a major step in precursor treatment for all activation precursor types (Islama et al., 2017; Njoku et al., 2014). In this case, a granular activation agent could be introduced during crushing at a known ratio. High- and low-speed crushers have been successfully deployed in previous work for different landfill wastes (Shen et al., 2013; Sua-iam and Makul, 2013). A similar crusher with the desired mesh size could be suitable. On this basis, the process may not require any specialized equipment apart from the routine equipment used in the activation process. However, an upgrade of some parts may be required.

The microwave energy source may be the major challenge for the scale-up process. The large-scale application of microwave energy is quite limited due to the complexity of electromagnetic waves’ reflection and absorption in the microwave unit (Motasemi and Afzal, 2013). Despite its numerous advantages, a full-scale application of the microwave-assisted activation processes is still being studied and is a focus for our on-going research.

Conclusion

These research findings provide essential information on the potential use of municipal landfill composite as a precursor for AC generation and its MB adsorption capacity. The percentage yields of AC of between 7.7% and 57.1% were mainly influenced by the type of landfill (active or closed), composite nature (more or less degraded) and activation parameters (wattage power and impregnation ratio). Carbonyl and hydroxyl groups were the major functional groups on the surface of AC. With an adsorption capacity of 34–190 mg/g for MB and Langmuir surface area of 361.8–636m²/g, the AC could be used to adsorb a range of cationic dyes and pollutants. The properties of AC generated from the landfill composite are comparable with those of AC from other types of biomass, such as oil palm stone and black liquor lignin, which have been identified as having the potential to be sustainably managed by conversion to AC (Fu et al., 2013; Tamai et al., 1996).

The landfill composites from both active and closed landfills represent a novel source of precursor for AC production and a new reuse option for landfill composite. This creates an opportunity for integrated landfill resource management in which the AC generated could be deployed as landfill liner and/or leachate pond adsorbent.

The adsorption capacity of the landfill composite AC suggests that its reuse as a daily landfill covering would curtail organic and inorganic mass transfer within the landfill layers via leaching. This could reduce the substantial cost of daily landfill covering, as well as the amount of virgin soil applied. The cost of daily landfill covering is estimated as 50% of the operational cost of municipal landfills (Johannessen and Boyer, 1997).

The depth of both landfill composites significantly influenced both the adsorption capacity and the surface area of the generated AC, while age difference did not have a definite influence. A new research finding indicates a relationship between depth of landfill composite and the quality of AC produced. A limitation of the presently reported results, as with most landfill parameters, is that they are dependent on the type of waste disposed, climatic conditions and the landfill management practice. It is therefore difficult to generalize the findings for all landfill types.

Further research on deep sampling, robust landfill mining strategy and activation conditions is required for this concept to have utility at the field-scale level.

Footnotes

Declaration of conflicting interests

The authors declared no potential conflicts of interest with respect to the research, authorship, and/or publication of this article.

Funding

The authors received no financial support for the research, authorship, and/or publication of this article.

References

Adelopo

Haris

Alo

et al . (2017a) Seasonal variations in moisture content and the distribution of total organic carbon in landfill composites: Case of active and closed landfills in Lagos, Nigeria. International Journal of Environment and Waste Management 20: 171–185.

Adelopo

Haris

Alo

et al . (2017b) Comparative characterisation of closed and active landfill composites using EDX, FTIR and proximate techniques. Waste and Biomass Valorization 8: 1313–1323.

Ali

Mohd. Khan

(2012) Low cost adsorbent of organic pollutants from wastewater. Journal of Environment Management 113: 170–183.

Alslaibi

Abustana

Ahmad

et al . (2013) A review production of activated carbon from agricultural by products via conventional and microwave heating. Journal of Chemical Technology & Biotechnology 88: 1183–1190.

Beekaroo

Mudhoo

(2011) Adsorption of Reactive Red 158 Dye by Chemically Treated Cocos nucifera L. Shell Powder. [ebook] Dordrecht, Heidelberg, New York: Springer.

Butt

Gouda

Balochd

et al . (2014) Literature review of baseline study for risk analysis – the landfill leachate case. Environment International 63: 149–162.

Danthurebandara

Van Passel

Vanderreydt

et al . (2015) Assessment of environmental and economic feasibility of Enhanced Landfill Mining. Waste Management 45: 434–447.

Di Maria

Micale

Sordi

et al . (2013) Urban mining: Quality and quantity of recyclable and recoverable material mechanically and physically extractable from residual waste. Waste Management 33: 2594–2599.

Dogan

Alkana

Demirbas

et al . (2006) Adsorption kinetics of maxilon blue GRL onto sepiolite from aqueous solutions. Chemical Engineering Journal 124: 89–101.

10.

Frändegård

Krook

Svensson

et al . (2013) A novel approach for environmental evaluation of landfill mining. Journal of Cleaner Production 55: 24–34.

11.

Ferrera-Lorenzo

Fuente

Suárez-Ruiz

et al . (2014). KOH activated carbon from conventional and microwave heating system of a macroalgae waste from the Agar–Agar industry. Fuel Processing Technology 121: 25–31.

12.

Foo

Hameed

(2012a) Mesoporous activated carbon from wood sawdust by K₂CO₃ activation using microwave heating. Bioresource Technology 111: 425–432.

13.

Foo

Hameed

(2012b) Factors affecting the carbon yield and adsorption capability of the mangosteen peel activated carbon prepared by microwave assisted K₂CO₃ activation. Chemical Engineering Journal 180: 66–74.

14.

Foo

Lee

Hameed

(2013) Preparation of tamarind fruit seed activated carbon by microwave heating for the adsorptive treatment of landfill leachate: A laboratory column evaluation, Bioresource Technology 133: 599–605.

15.

Yue

Gao

et al . (2013) Preparation, characterization and application of lignin-based activated carbon from black liquor lignin by steam activation. Chemical Engineering Journal 228: 1074–1082.

16.

García

Pizarro

Lavín

et al . (2013) Biomass proximate analysis using thermogravimetry. Bioresource Technology 139: 1–4.

17.

Gerçel

Ozcan

et al. ( 2007) Preparation of activated carbon from a renewable bio-plant of Euphorbia rigida by H₂SO₄ activation and its adsorption behavior in aqueous solutions. Applied Surface Science 253: 4843–4852.

18.

Hoornweg

Bhada-Tata

(2012) What a Waste: A Global Review of Solid Waste Management. Washington, DC: Urban Development & Local Government Unit, World Bank.

19.

Srinivasan

(2001) Novel activation process for preparing highly microporous and mesoporous activated carbons. Carbon 39: 877–886.

20.

IPCC (International Panel on Climate Change) (2006) Guidelines for national greenhouse gas inventories Volume 5: Waste. Available at: www.ipcc-nggip.iges.or.jp/public/2006gl/vol5.html (accessed 20 August 2014).

21.

Islama Md

Ahmed

Khandaya

et al . (2017) Mesoporous activated carbon prepared from NaOH activation of rattan (Lacosperma secundiflorum) hydrochar for methylene blue removal. Ecotoxicology and Environmental Safety 138: 279–285.

22.

Jain

Powell

Smith

et al . (2014). Life-cycle inventory and impact evaluation of mining municipal solid waste landfills. Environmental Science & Technology 48: 2920−2927.

23.

Zhu

Wang

et al . (2007). Preparation of activated carbons by microwave heating KOH activation. Applied Surface Science 254: 506–512.

24.

Johannessen

Boyer

(1997) Observations of Solid Waste Landfills in Developing Countries: Africa, Asia, and Latin America. Washington, DC: Urban Development Division, Waste Management Anchor Team, World Bank.

25.

Kaartinen

Sormunen

Rintala

(2013) Case study on sampling, processing and characterization of landfilled municipal solid waste in the view of landfill mining. Journal of Cleaner Production 55: 56–66.

26.

Karago

Tay

Suat

et al . (2008) Activated carbons from waste biomass by sulfuric acid activation and their use on methylene blue adsorption. Bioresource Technology 99: 6214–6222.

27.

Zhanga

Penga

et al . (2008) Preparation of high surface area activated carbons from tobacco stems with K₂CO₃ activation using microwave radiation. Industrial Crops and Products 27: 341–347.

28.

Liou

. (2010) Development of mesoporous structure and high adsorption capacity of biomass based activated carbon by phosphoric acid and zinc chloride activation. Chemical Engineering Journal 158: 129–142.

29.

Moreno-Castilla

(2004) Adsorption of organic molecules from aqueous solutions on carbon materials. Carbon 42: 83–94.

30.

Motasemi

Afzal

(2013) A review on the microwave-assisted pyrolysis technique. Renewable and Sustainable Energy Reviews 28: 317–330.

31.

Nahil

Williams

(2012) Surface chemistry and porosity of nitrogen-containing activated carbons produced from acrylic textile waste. Chemical Engineering Journal 184: 228–237.

32.

Njoku

Foo

Asif

et al . (2014) Preparation of activated carbons from rambutan (Nephelium lappaceum) peel by microwave-induced KOH activation for acid yellow 17 dye adsorption. Chemical Engineering Journal 250: 198–204.

33.

Prauchner

Rodríguez-Reinoso

(2012) Chemical versus physical activation of coconut shell: A comparative study. Microporous and Mesoporous Materials 152: 163–171.

34.

Septhupathi

Bashir

MJK

Akbar

et al . (2015) Biomass-based shell activated carbon and palm shell carbon molecular sieve as gas separation adsorbents. Waste Management and Research 33: 303–312.

35.

Shen

Yuah

Pang

et al . (2013) Performance of anaerobic co-digestion of fruit and vegetable waste (FVW) and food waste (FW): Single-phase vs two-phase. Bioresource Technology 144: 80–85.

36.

Sing

KSW

Everett

Haul

RAW

et al . (1985). Reporting physisorption data for gas/solid systems with special reference to the determination of surface area and porosity. Pure and Applied Chemistry 57: 603–619.

37.

Song

Wei

et al . (2016) The application of prepared porous carbon material: Effect of different components on the heavy metal adsorption. Waste Management and Research 34: 534–541.

38.

Sua-iam

Makul

(2013) Use of increasing amount of bagasse ash waste to produce self-compacting concrete by adding limestone powder waste. Journal of Cleaner Production 57: 308–319.

39.

Suhas Carrott

PJM

Ribeiro Carrott

MML

(2007) Lignin from natural adsorbent to activated carbon: A review. Bioresource Technology 98: 2301–2312.

40.

Tamai

Kakii

Hirota

et al . (1996) Synthesis of extremely large mesoporous activated carbon and its unique adsorption for giant molecules. Chemistry of Materials 8: 454–462.

41.

Tang

Liu

Chen

(2012) Preparation and characterization of activated carbon from waste ramulus mori. Chemical Engineering Journal 203: 19–24.

42.

Tran

Chao

You

(2017) Activated carbons from golden shower upon different chemical activation methods: Synthesis and characterization. Adsorption Science and Technology 36: 95–113.

43.

Uddin

Islam

Mahmud

et al . (2009). Adsorptive removal of methylene blue by tea waste, Journal of Hazardous Materials 164: 53–60.

44.

USEPA (2002) RCRA Waste Sampling Draft Technical Guidance. Planning, Implementation and Assessment. EPA530-D-02002. Washington, DC: Office of Solid Waste, U.S. Environmental Protection Agency.

45.

Zhou

Gong

et al . (2015) A cost-benefit analysis of landfill mining and material recycling in China. Waste Management 35: 191–198.