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Characteristics of different molecular weight fractions of organic matter in landfill leachate and their role in soil sorption of heavy metals
Environmental Pollution 113 (2001) 331±339
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Characteristics of dierent molecular weight fractions of organic matter in land®ll leachate and their role in soil sorption of heavy metals N. Calace a,*, A. Liberatori b, B.M. Petronio a, M. Pietroletti a a Department of Chemistry, University ``La Sapienza'', p.le Aldo Moro 5, 00185 Rome, Italy Water Research Institute, National Research Council (CNR), via Reno 1, 00198 Rome, Italy
b
Received 27 March 2000; accepted 4 August 2000
``Capsule'': A land®ll leachate characterisation, based on the distribution of dierent molecular weights of constituents is proposed. Abstract We have characterised two kinds of municipal land®ll leachates derived from `old' and `young' municipal waste land®lls on the basis of the molecular weight distribution of the constituents, taking into account that the great variety of leachate constituents prevents any evaluation of the fate and of the role played by each component in the environmental impact. In the sample S1 (old leachate), the constituents were distributed over a wider range of molecular weights; high molecular weight fractions were present. In sample S2 (young leachate), the fractions are actually narrower at the lower molecular weights. The high molecular weight fractions of old leachates are found to be complex structures formed by condensed nuclei of carbons substituted by functional groups containing nitrogen, sulphur and oxygen atoms; the low molecular weight fractions of leachates are, instead, characterised by linear chains substituted by oxygenated functional groups such as carboxyl and/or alcoholic groups. After characterising each fraction we studied the role played by these fractions in the soil's capability for retaining heavy metals [copper (Cu) and cadmium(Cd)]. The Cd uptake increases only on the soil treated with sample S1 characterised by a higher pH value and by the presence of high molecular weight fractions. The Cu uptake also increases on the soil treated with sample S2, characterised by the sole presence of low molecular weight fractions. On the other hand, the metal adsorption tests performed on soil treated with the single fractions show that the amount of Cu and Cd retained by soil treated with the high molecular weight fractions of sample does not increase after 72 h of treatment and that the amount of Cu retained by the low molecular weight fractions of sample S1 and by the fractions of sample S2 increases, but does not justify the amount retained by soil treated with the total leachates. # 2001 Elsevier Science Ltd. All rights reserved. Keywords: Land®ll leachate; Molecular weights distribution; Spectroscopic characterisation; Metal soil sorption
1. Introduction Economic considerations continue to maintain land®lls as the most attractive disposal route for municipal solid waste. Alternative methods to land®lling (incineration, composting) are actually considered as volume reduction processes because they produce waste fractions (ashes, slag) which ultimately must be land®lled (Bingener and Crutzen, 1987; Emberton and Parker,
* Corresponding author. Tel./fax: +39-064-991-3723. E-mail address: [email protected] (N. Calace).
1987; Cossu, 1989; Gendebien et al., 1992; Nozhevnikova et al., 1992; Lee et al., 1993). Despite the evolution of land®ll technology from open, uncontrolled dumps to highly engineered facilities designed to eliminate or minimise the potential adverse impact of the waste on the surrounding environment, generation of contaminated leachate remains an inevitable consequence of the practice of waste disposal in land®lls. The subsequent migration of leachate away from land®ll boundaries and its release into the adjacent environment is a serious environmental pollution concern and a threat to public health and safety in both old
0269-7491/01/$ - see front matter # 2001 Elsevier Science Ltd. All rights reserved. PII: S0269-7491(00)00186-X
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and new facilities (MacFarlane et al., 1983; Mackay et al., 1985; Albaiges et al., 1986). Leachate is formed when the refuse moisture content exceeds its ®eld capacity, which is de®ned as the maximum moisture that can be retained in a porous medium without producing downward percolation (Blakey, 1982, 1989; Matejka et al., 1994). Soluble organic and inorganic compounds are encountered in the refuse at emplacement or are formed as a result of chemical and biological processes within the land®ll. Leachate formation creates a non-uniform and intermittent percolation of moisture through the refuse mass which results in the removal of these soluble compounds from the refuse and their dissolution in the leachate (Lema et al., 1988; Lisk, 1991; Wu and Wang, 1992; Oman and Hynning, 1993). The composition of land®ll leachate can exhibit considerable spatial and temporal variations. The great variety of leachate constituents prevents evaluation of the fate and the role played by each component in the environmental impact, so we propose a leachate characterisation based on the distribution of dierent molecular weights, taking into account the fact that the behaviour and reactivity of organic substances may partly depend on their molecular weight or size in solution (Chian, 1977; Chian and De Walle, 1977; Harmesen, 1983; Gourdon et al., 1989; Frimmel and Weis, 1991; Christensen et al., 1998). On the other hand, some researches highlighted the fundamental role played by dissolved organic matter (DOM) in the mobilisation/ immobilisation and transport of pollutants (McCarthy and Zachara, 1989; Harter and Naidu, 1995; Kaiser, 1998; Naidu and Harter, 1998) and some researches pointed out that the DOM interacts intensively with mineral surfaces altering signi®cantly their characteristics (Jardine et al., 1989; Amal et al., 1992; Godtfredsen and Stone, 1994; Evanko and Dzombak, 1998). The aim of this research is, therefore, to study the role played by the dierent molecular weight fractions of leachate constituents in soil sorption of copper (Cu) and cadmium (Cd). For this purpose, leachates were divided into fractions having dierent molecular weight ranges and soil-metal and soil-leachate-metal systems studied in batch conditions. 2. Experimental 2.1. Materials One leachate (S1; old leachate) was collected from a municipal waste land®ll more than 10 years old and another (S2; young leachate) from a municipal waste land®ll established only 4 years ago. The municipal waste land®lls were situated in Pomezia, a small city near Rome (Italy), and the same type of solid waste was
deposited on them. At both sites, the leachates were sampled and frozen at 20 C. S1 sample properties were: pH value of 8.5; chemical oxygen demand (COD) value of 2400 mg l 1; Cu concentration of 10 mg l 1; Lead (Pb) of 32 mg l 1, Zinc (Zn) of 60 mg l 1; nickel (Ni) of 284 mg l 1; and Cd of 0.5 mg l 1. S2 properties were: pH value of 6.4; COD value of 9100 mg l 1; Cu concentration of 3 mg l 1, Pb of 16 mg l 1; Zn of 120 mg l 1; Ni of 55 mg l 1; and Cd of 0.2 mg l 1. The super®cial (0±20 cm) soil features were a pH value of 6.5 (the soil was treated with distilled water, soil:distilled water ratio 1:2.5, for 3 h and then the pH value of the liquid phase was measured), a cation exchange capacity of 9.6 cmol kg 1 of soil, a speci®c surface area of 12 m2 g 1, a content of clay, silt and sand of 22.6, 25.0 and 52.4%, respectively. The organic matter content was 1.6% and the total heavy metal content was Cu 50 mg kg 1, Pb and Zn 90 mg kg 1, Ni 20 mg kg 1 and Cd not detectable. Model Centriplus Amicon concentrators supported low-binding membranes with dierent molecular weight cut-o (100 000, 50 000, 30 000, 10 000 Dalton) were used. Amicon low-binding membranes (molecular weight cut-o of 500 Dalton and diameter of 43 mm) were used in the ultra®ltration system. Cu(II) and Cd(II) solutions (from 0.03 to 0.3 mmol l 1) obtained by dissolving Cu(NO3)2 and Cd(NO3)2 salt in distilled water were employed, then checking the metal concentration with atomic absorption spectroscopy, and the pH values were from 5.7 to 5.4 (0.05). Reagents of analytical grade were used throughout. 2.2. Apparatus Molecular weight fractionation using Amicon concentrators was carried out with a VISMARA centrifuge model 3225R (max. speed 8000 g). The separation of the last two fractions was obtained by using an Amicon stirred ultra®ltration cell, model 8050, capacity 50 ml, supporting the Amicon membranes with a molecular weight cut-o of 500 Dalton. Elemental analyses were carried out in the Microanalysis Laboratory of the Italian Research Council, Monte Libretti, Rome (Italy) using a Carlo Erba 240-B model. 13 C-nuclear magnetic resonance (NMR) spectra were determined in D2O by a Varian spectrometer, model XL300. Samples were prepared by dissolving the dried residue (30 mg) in 1 ml of D2O in an NMR tube (5 mm). The operating conditions were: 75 MHz, pulse 45 , acquisition time 0.1 s, delay time 0.5 s. About 800 000 scans were accumulated. Spectra were performed in broadband decoupling. Under these conditions, a clear spectrum should be obtained with distinct single peaks for each
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carbon (C) atom; because of the complexity of our samples, the spectra show many poorly resolved signals. Fourier transform infrared (FTIR) spectra were recorded by means of a FTIR Philips spectrometer model PU9800 working in diuse re¯ectance conditions. For quantitative determinations, re¯ectance spectra were converted to Kubelka Munk units, which are directly proportional to the concentration of the scattering medium. Samples were prepared by drying 1 ml of liquid sample to which 100 mg of anhydrous KBr had been added; then sample discs were prepared with the dried residue homogenised in a mortar. Metal concentrations were determined by a model Varian AA 10 ¯ame atomic absorption spectrophotometer.
Soil (200 mg) was shaken up with 8 ml Cu(II) or Cd(II) solution of dierent concentrations (from 0.03 to 0.3 mmol l 1) in a controlled temperature box 201 C for 24 h. Preliminary experiments showed no signi®cant variation in metal sorption after 24 h and after batching treatment the pH value of solutions was 7.30.1 for samples S1 and 6.40.1 for samples S2. After centrifugation at 3000 g for 30 min the liquid phases were decanted and the soil washed with deionised water; the washes were then added to the liquid phase and metal concentration determined by atomic adsorption spectroscopy. The amount of sorbed metal was calculated by the dierence between initial and ®nal metal concentration.
2.3. Separation of leachate constituents
3. Results and discussion
The separation of leachate constituents based on the molecular weight size was carried out using Amicon concentrators designed for use in centrifuges with ®xedangle rotors (34 ) or a swinging bucket. The spin time was optimised for each cut o (100, 50, 70, 190 min, respectively, centrifugal force 3000 g) in order to obtain a ®nal retentate volume of 0.5 ml. Initial leachate volume used was 10 ml. The last fraction obtained with concentrators was further divided in a stirred Amicon ultra®ltration cell connected to a reservoir (5 l) containing dia®ltrate solution (deionized water) pressurised using nitrogen (3 atm). Since concentrator membranes contain trace amounts of glycerine, before using the membranes, washing was performed by centrifugation with 0.1 M NaOH (three times for 1 h) and the residual NaOH subsequently removed by washing with deionised water until a neutral pH value was obtained. After separation, each leachate fraction was again made up to initial volume (10 ml). A portion of each solution was dried under an infrared lamp (60 C), weighed and the obtained dried residue used for analytical characterisation (FTIR, 13C-NMR, elemental analysis).
3.1. Characterisation of molecular weight fractions Molecular weight distributions of leachate constituents are shown in Fig. 1. In the sample S1 (old leachate), the constituents were distributed over a wider range of molecular weights; high molecular weight fractions were present. The data con®rm the results obtained concerning the presence of humic-like substances (Artiola-Fortuny and Fuller, 1982; Weis et al.,
2.4. Adsorption isotherms Metal sorption experiments were carried out in batch conditions on the untreated soil and on the soil treated both with leachates and with the dierent leachate fractions. The treated soil was obtained by batching the untreated soil with leachates or with the dierent leachate fractions for 48 h (time experimentally determined by means of kinetic studies of sorption process, soil± liquid phase ratio 1:2). After centrifugation, the soil was washed with deionised water, lyophilised and used for the sorption experiments.
Fig. 1. The molecular weight fractions distribution of sample S1 (old leachate) and sample S2 (young leachate).
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1989; Calace and Petronio, 1997). In sample S2 (young leachate), the fractions are actually narrower at the lower molecular weights, in agreement with literature data (Chian, 1977; Chian and De Walle, 1977; Frimmel and Weis, 1991) for which the molecular weight distribution over a wider range increases with the increasing age of land®ll leachates. The elemental compositions of fractions are given in Table 1. In sample S1, the percentage of C was found to be highest in the 500±10 000 Dalton fraction and lowest in the 30 000±50 000 Dalton fraction. The nitrogen (N)/ C ratio was the same for the dierent molecular weight constituents while that of sulphur (ratio S/C) was the same except in the 500±10 000 Dalton fraction where it was lower. As regards the ratio hydrogen (H)/C data, the fractions with lower molecular weights had the highest values. Hence, the elemental composition of sample S1 fractions highlights the fact that there is not much dierence between them. In sample S2, the elemental composition is the same for each molecular weight fraction, which are all characterised by high H/C values; N and S content is negligible. Taking into account the data obtained, we performed some spectroscopic analyses to estimate the nature and kind of the functional groups characterising the dierent fractions. Indeed, the nature of the functional groups forming the compounds ultimately in¯uences the characteristics of the compound and can thus contribute greatly to the determination of mechanisms of metal accumulation and/or metal transport in soils. For the interpretation of land®ll leachate fraction spectra, we based on literature data referred to humic substances (Artiola-Fortuny and Fuller, 1982; Harmesen, 1983; Calace and Petronio, 1997) which are natural macromolecules having complex structures; any papers on land®ll leachate spectroscopic analysis are not found in the literature.
As regards sample S1, the spectroscopic analyses performed on each fraction highlighted some structural dierences. The FTIR spectrum of the >100 000 Dalton fraction (Fig. 2a) shows signals at 2916 and 2857 cm 1 which are attributed to asymmetric and symmetric stretching of methyl and methylene groups, as well as a band at 1406 cm 1 due to bending of CH2 adjacent to CO oxygen ( CH2 CO). Moreover, signals at 1661 and 1559 cm 1 due to the ®rst and second band of amides (stretching of CO and C N bonds, respectively) and a sorption at 1050 cm 1 associated with the C OH bond in secondary alcohols were present. 13CNMR spectrum (Fig. 3a) displayed a broad resonance band in the region 0±30 ppm due to aliphatic carbons of branched chains, a signal at 40 ppm due to the C N bond; poorly resolved signals are also present in the 60± 80 ppm region associated with C bound to O atoms, and in the 80±140 ppm region due to ole®nic carbons. The presence of CO groups is highlighted by resonances around 170 ppm. The spectroscopic features of the 50 0000-100 000 Dalton and 10 000-30 000 Dalton fractions are similar to the >100 000 Dalton one. Only the ratios among the intensities of sorption change. The spectroscopic results of 30 000±50 000 Dalton and 500±10 000 Dalton fractions are similar. The FTIR spectrum of the latter (Fig. 2b) shows a shoulder at 1670 cm 1 that, with the signals at 2990, 1600 and 1420 cm 1, can be attributed to CC. Moreover, absorption in the region of 1100 cm 1 is typical of stretching C O bonds in alcohols and ethers. 13C-NMR spectrum (Fig. 3b) shows signals in the 0±50 ppm region associated with the presence of aliphatic carbons in more or less
Table 1 Elemental analysis data of dierent molecular weight fractions of leachatesa C% N% H% O%
S% N/C O/C S/C
H/C
Sample S1 >100 000 D 50 000±100 000 D 30 000±50 000 D 10 000±30 000 D 500±10 000 D <500 D
45.0 47.5 36.2 44.4 54.2 44.1
1.07 1.25 0.94 1.20 0.58 1.67
1.52 1.57 1.76 1.40 1.95 2.58
Sample S2 10 000±30 000 D 500±10 000 D <500 D
53.8 0.45 7.70 37.60 0.45 0.007 0.52 0.007 1.72 53.6 0.45 8.80 36.70 0.45 0.007 0.51 0.007 1.97 57.2 0.48 8.95 33.07 0.32 0.007 0.89 0.002 1.88
a
7.14 6.70 5.30 5.60 6.00 3.72
5.70 6.20 5.30 5.20 8.80 9.50
41.09 38.35 52.26 43.60 30.42 41.01
0.14 0.12 0.12 0.11 0.09 0.07
0.68 0.60 1.08 0.74 0.42 0.70
0.009 0.010 0.010 0.010 0.004 0.014
The data are the average of ®ve determinations; standard deviation is less than 3%.
Fig. 2. Fourier transform infrared spectra of sample S1 (old leachate) fractions: (a) >100 000 D fraction; (b) 500±10000 D; (c) <500 D.
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Fig. 3.
13
335
C-NMR spectra of sample S1 (old leachate) fractions: (a) >10 000 D fraction; (b) 500±10000 D.
branched chains and of methoxyl groups, signals around 70 ppm due to alcohols and ethers, resonances at 129 ppm due to the CC bond and ®nally, signals around 180 ppm attributed to CO bond of carbonyl groups. Finally, the FTIR spectrum of the <500 Dalton fraction (Fig. 2c) highlight its prevalently aliphatic nature and the presence of a high inorganic ion content in the form of carbonates and nitrates (FTIR sorptions at 1460 and 1372 cm 1). In sample S2, the spectroscopic analyses of each fraction con®rm the homogeneity in the functional group distribution. The FTIR spectra (Fig. 4) show predominant signals at 1580 and 1440 cm 1 due to stretching and bending of CO in carboxylated groups (COO ) and two absorption bands at 2960 and 2900
cm 1 due to CH2 in long chains of fatty acids. Also 13CNMR spectra (not shown) show a large number of signals due to aliphatic carbons (10±40 ppm) and a resonance at 180 ppm due to CO bonds. 3.2. Metal sorption The adsorption isotherms, a graphical representation of the relationship between the equilibrium concentration of adsorbate in solution and the equilibrium concentration of adsorbate adsorbed by the soil particles, are shown in Figs. 5 and 6. These isotherms are obtained from batch tests on untreated soil and on the soil-leachates system. Cu is already sorbed to a greater extent than Cd onto the original soil in agreement with literature data (Yong et al., 1992). The Cu, indeed, has a
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higher anity for soil surface sorbing sites than the Cd (Yong et al., 1992). After the sample S1 treatments, both Cu and Cd were retained by the soil increases and the increase was not the same for two metals. This result is probably due to the dierent features of metals, such as anity to organic matter, formation of inner- and/or outer-sphere complexes, etc., as well as the structural features of the soil surface which changes with the leachate treatment. Numerous investigators (Tiller et
Fig. 4. Fourier transform infrared spectra of sample S2 (young leachate) fractions: (a) >10 000 D fraction; (b) 500±10000 D; (c) <500 D.
al., 1984; Brummer et al., 1988; Gerth et al., 1993; Naidu et al., 1994, 1997) showed that soil solution pH has a marked in¯uence on adsorption of metals that increases when soil solution pH increases and, between Cd and Cu, the former is more in¯uenced by pH value changes (Naidu et al., 1994). On the other hand, the Cu can interact, forming stable complexes with high and low molecular weight organic substances while the Cd can only interact with the high molecular weight organic compounds (Godtfredsen and Stone, 1994; Naidu et al., 1997). Sample S1 is characterised by a high pH value and by the presence of both high molecular weight organic substances and low molecular weight organic substances. In general, it can be hypothesised that the sample S1 in¯uences the soil sorption capability both acting on pH value by increasing it and creating new sorbing sites due to the presence of organic matter. This hypothesis is con®rmed by the tests performed with the sample S2, characterised by pH value similar to pH value of original soil and by the presence of the only low molecular weight organic compounds. After treatment with the sample S2, the soil retention capability only increases for Cu. Our results ®t the Langmuir equation and the experimental data have an H-type pattern according to Giles' classi®cation (Giles et al., 1974a, b). These curves show a long plateau and are characteristic of interactions that are essentially electrostatic in nature between species adsorbing and soil active surface. The length of the plateau indicates the diculty of formation of a second layer because of charge repulsion between the adsorbates and those approaching from the solution. Moreover, some authors have demonstrated that sorption of
Fig. 5. Graphical representation of copper sorption. S1, old leachate; S2, young leachate.
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Fig. 6. Graphical representation of cadmium sorption. S1, old leachate; S2, young leachate.
Cu and Cd onto clay minerals follows Freundlich isotherms, whereas sorption onto humic substances follows the Langmuir isotherms (Yong et al., 1992) con®rming the hypothesis about the important role played by high molecular weight organic matter in heavy metal sorption process. The soil treated with each fraction of sample S1 (Table 2) shows a dierent pattern. The metal adsorption tests performed on soil treated with the single fractions show that the amount of Cu and Cd retained by soil treated with the high molecular weight fractions of sample does not increase after 72 h of treatment. Indeed, the soil treated with the high molecular weight fractions does not change the retention capacity of either Cu or Cd. After treatment with low molecular weight fractions, on the other hand, only the amount of retained Cu is weakly increased. Moreover, the amount of Cu sorbed onto the soil treated with low molecular weight fractions does not justify the value of qmax found for the soil treated with the total sample S1. In this case, the pH value increase of soil does not seem to in¯uence the metal retention capacity (the pH values of fractions were 8.50.1) but, on the basis of the data obtained, it seems that the increase of the metal retention capacity of the soil treated with the total leachate is not due to a particular fraction of leachate but to the presence of all fractions at the same time. We have already said that after treatment with sample S2, the soil increases metal retention capacity only in the case of Cu. The Cd sorption on soil, indeed, does not seem to be in¯uenced by the presence of sample S2.
Table 2 The maximum amount of metals retained from soilsa Cu qmax
Cd qmax
Original soil
4.37
1.57
S1 Total S1 >100 000 D S1 50 000±100 000 D S1 30 000±50 000 D S1 10 000±30 000 D S1 500±10 000 D S1 <500 D
6.81 4.39 4.31 4.06 3.81 4.73 4.88
3.65 1.81 1.70 1.71 1.59 1.69 1.66
S2 Total S2 10 000±30000 D S2 500±10 000 D S2 <500 D
6.02 5.20 4.95 5.29
1.67 1.34 1.33 1.42
a The data are the average of ®ve determinations; standard deviation is less than 5%.
Even treating the soil with each fraction increases Cu retention while Cd retention actually decreases (Table 2). 4. Conclusion Since the composition of land®ll leachate can exhibit considerable spatial and temporal variations, an interesting approach to characterise the municipal land®ll leachates consists of the separation of leachate constituents in dierent molecular weight fractions. In particular, the old leachate constituents were distributed
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over a wider range of molecular weights; high molecular weight fractions were present. The young leachate fractions, instead, are actually narrower at the lower molecular weights, in agreement with literature data. Moreover, the nature of the functional groups forming the compounds can in¯uence the characteristics of the compound and can thus contribute greatly to the determination of mechanisms of metal accumulation and/or metal transport in soils. The high molecular weight fractions of old leachates are found to be complex structures formed by condensed nuclei of carbons substituted by functional groups containing N, S and O atoms; the low molecular weight fractions of leachates are, instead, characterised by linear chains substituted by oxygenated functional groups such as carboxyl and/ or alcoholic groups. The impact of the land®ll leachates on the environment can be estimated on the basis of its chemical characteristics (pH value, type and content of organic matter, etc.) and an evaluation of the role played by leachate constituents on the fate of metals in the soils can be done. We found that the Cd uptake increases only on the soil treated with sample S1, characterised by a higher pH value and by the presence of high molecular weight fractions. The Cu uptake also increases on the soil treated with sample S2, characterised by the sole presence of low molecular weight fractions. On the other hand, the metal adsorption tests performed on soil treated with the single fractions show that the amount of Cu and Cd retained by soil treated with the high molecular weight fractions of sample does not increase after 72 h of treatment and that the amount of Cu retained by the low molecular weight fractions of sample S1 and by the fractions of sample S2 increases, but does not justify the amount retained by soil treated with the total leachates. Taking into account that the leachate±soil interactions depend on the nature of soil type and on the chemical features of leachate constituents, we can hence hypothesise that leachates are able to increase the metal retention capacity of the soil but that this ability does not depend on a particular fraction of leachate, but on the presence of all fractions at the same time. References Albaiges, J., Casado, F., Ventura, F., 1986. Organic indicators of groundwater pollution by a sanitary land®ll. Water Research 20, 1153±1159. Amal, R., Raper, J.A., Waite, T.D., 1992. Eect of fulvic acid adsorption on the aggregation kinetics and structure of hematite particles. Journal of Colloid and Interface Science 151, 244±257. Artiola-Fortuny, J., Fuller, W.H., 1982. Humic substances in land®ll leachates: I. Humic acid extraction and identi®cation. Journal of Environmental Quality 11 (4), 663±669. Bingener, H.G., Crutzen, P.J., 1987. The production of methane from solid wastes. Journal of Geophysical Research 92, 2182±2187.
Blakey, N.C., 1982. In®ltration and Adsorption of Water by Domestic Wastes in Land®ll Sites. Hawell Land®ll Leachate Symp. Harwell, Oxon, UK. Blakey, N.C., 1989. Model prediction of land®ll leachate production. In: Christensen, T.H., Cossu, R., Stegmann, R. (Eds.), Sanitary Land®lling: Process, Technology and Environmental Impact. Academic Press, San Diego, CA, pp. 17±34. Brummer, G.W., Gerth, J., Tiller, K.G., 1988. Reaction kinetics of the adsorption and desorption of Ni, Zn and Cd by geothite. I. Adsorption and diusion of metals. Journal of Soil Science 39, 35± 52. Calace, N., Petronio, B.M., 1997. Characterization of high molecular weight organic compounds in land®ll leachate: humic substances. Journal of Environmental Science Health A32 (8), 2229±2244. Chian, E.S.K., 1977. Stability of organic matter in land®ll leachates. Water Research 11, 225±232. Chian, E.S.K., De Walle, F.B., 1977. Characterization of soluble organic matter in leachate. Environmental Science and Technology 11, 158±163. Christensen, J.B., Jensen, D.L., Gron, C., Filip, Z., Christensen, T.H., 1998. Characterization of the dissolved organic carbon in land®ll leachate-polluted groundwater. Water Research 32, 125±135. Cossu, R., 1989. Role of land®lling in solid waste management. In: Christensen, T.H., Cossu, R., Stegmann, R. (Eds.), Sanitary Land®lling: Process, Technology and Environmental Impact. Academic Press, San Diego, CA, pp. 3±9. Emberton, J.R., Parker, A., 1987. The problems associated with building on land®ll sites. Waste Managment and Research 5, 473±482. Evanko, C.R., Dzombak, D.A., 1998. In¯uence of structural features on sorption of NOM-analogue organic acids to geothite. Environmental Science and Technology 32, 2846±2855. Frimmel, F.H., Weis, M., 1991. Ageing eects of high molecular weight organic acids which can be isolated from land®ll leachates. Water Science and Technology 23, 419±426. Gendebien, A., Pauwels, M., Constant, M., Ledrut-Damanet, M.J., Willumsen, H.C., Butson, J., Fabry, R., Ferrero, G.L., Nyns, E.L., 1992. Land®ll Gas: From Environment to Energy. Oce for Ocial Publications of the European Community, Luxembourg. Gerth, J., Brummer, G.W., Tiller, K.G., 1993. Retention of Ni, Zn and Cd by Si-associated geothite. Zeitschrift fur P¯anzenernahrung und Bodekunde 156, 123±129. Giles, C.H., Smith, D., Huitson, A., 1974a. A general treatment and classi®cation of the solute adsorption isotherm. I. Theoretical. Journal of Colloid Interface Science 47, 755±765. Giles, C.H., D'Silva, A.P., Easton, I.A., 1974b. A general treatment and classi®cation of the solute adsorption Isotherm. II. Experimental interpretation. Journal of Colloid Interface Science 47, 766± 778. Godtfredsen, K.L., Stone, A.T., 1994. Solubilization of manganese dioxide-bound copper by naturally occurring organic compounds. Environmental Science and Technology 28, 1450±1458. Gourdon, R., Comei, C., Vermande, P., Veron, J., 1989. Fractionation of the organic matter of a land®ll leachate before and after aerobic or anaerobic biological treatment. Water Research 23, 167±173. Harmesen, J., 1983. Identi®cation of organic compounds in leachate from a waste tip. Water Research 6, 699±714. Harter, R.D., Naidu, R., 1995. Role of metal-organic complexation on metal sorption by soils. Advances in Agronomy 55, 219±264. Jardine, P.M., Weber, N.L., McCarthy, J.F., 1989. Mechanisms of dissolved organic carbon adsorption on soil. Soil Science Society of America Journal 53, 1378±1385. Kaiser, K., 1998. Fractionation of dissolved organic matter aect by polivalent metal cations. Organic Geochemistry 28 (12), 849±854. Lee, J.J., Jung, I.H., Lee, W.B., Kim, J.O., 1993. Computer and experimental simulations of the production of methane gas from municipal solid waste. Water Science and Technology 27, 225±234.
N. Calace et al. / Environmental Pollution 113 (2001) 331±339 Lema, J.M., Mendez, R., Blazquez, R., 1988. Characteristics of land®ll leachates and alternatives for their treatment: a review. Water, Air and Soil Pollution 40, 223±250. Lisk, D.J., 1991. Environmental eects of land®lls. The Science of the Total Environment 100, 415±467. MacFarlane, D.S., Cherry, J.A., Gillham, R.W., Sudicky, E.A., 1983. Migration of contaminants in groundwater by a land®ll: a case study. Journal of Hydrology 63, 1±29. Mackay, D.M., Roberts, P.V., Cherry, J.A., 1985. Transport of organic contaminants in groundwater. Environmental Science and Technology 19, 384±392. Matejka, G., Rinke, M., Mejbri, R., Bril, H., 1994. Pollution engendree par un lixiviat de dicharge d'ordures menageres: bilan hydrique et characterization. Environmental Technology 15, 313±322. McCarthy, J.F., Zachara, J.M., 1989. Subsurface transport of contaminants. Environmental Science and Technology 23, 496±502. Naidu, R., Harter, R.D., 1998. Eectiveness of dierent organic ligands on sorption and extractibility of cadmium by soils. Soil Science Society America Journal 62, 644±650. Naidu, R., Bolan, N.S., Kookana, R.S., Tiller, K.G., 1994. Ionic strength and pH eects on the sorption of cadmium and the surface charge of soils. European Journal of Soil Science 45, 419±429.
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Naidu, R., Kookana, R.S., Sumner, M.E., Harter, R.D., Tiller, K.G., 1997. Cadmium sorption and transport in variable charge soils: a review. Journal of Environmental Quality 26, 602±617. Nozhevnikova, A.N., Lebedev, V.S., Lifshitz, A.B., 1992. Microbiological processes occurring in land®lls. Proceedings of the International Symposium on Anaerobic Digestion of Solid Waste, Venice, Italy pp. 303±312. Oman, C., Hynning, P., 1993. Identi®cation of organic compounds in municipal land®ll leachates. Environmental Pollution 80, 265± 271. Tiller, K.G., Gerth, J., Brummer, G., 1984. The relative anities by Cd, Ni and Zn for dierent clay fractions and geothite. Geoderma 34, 17±35. Weis, W., Abbt-Braun, G., Frimmel, F.H., 1989. Humic-like substances from land®ll leachates Ð characterization and comparison with terrestrial and aquatic humic substances. The Science of the Total Environment 81/82, 343±352. Wu, S.C., Wang, M.S., 1992. Modelling the organic strength of leachate from sanitary land®lls. Bulletin of the College of Engineering, N.T.U 55, 149±164. Yong, R.N., Mohamed, A.M.O., Warkentin, B.P., 1992. Principles of Contaminant Transport in Soils. Elsevier, The Netherlands.
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Characteristics of dierent molecular weight fractions of organic matter in land®ll leachate and their role in soil sorption of heavy metals N. Calace a,*, A. Liberatori b, B.M. Petronio a, M. Pietroletti a a Department of Chemistry, University ``La Sapienza'', p.le Aldo Moro 5, 00185 Rome, Italy Water Research Institute, National Research Council (CNR), via Reno 1, 00198 Rome, Italy
b
Received 27 March 2000; accepted 4 August 2000
``Capsule'': A land®ll leachate characterisation, based on the distribution of dierent molecular weights of constituents is proposed. Abstract We have characterised two kinds of municipal land®ll leachates derived from `old' and `young' municipal waste land®lls on the basis of the molecular weight distribution of the constituents, taking into account that the great variety of leachate constituents prevents any evaluation of the fate and of the role played by each component in the environmental impact. In the sample S1 (old leachate), the constituents were distributed over a wider range of molecular weights; high molecular weight fractions were present. In sample S2 (young leachate), the fractions are actually narrower at the lower molecular weights. The high molecular weight fractions of old leachates are found to be complex structures formed by condensed nuclei of carbons substituted by functional groups containing nitrogen, sulphur and oxygen atoms; the low molecular weight fractions of leachates are, instead, characterised by linear chains substituted by oxygenated functional groups such as carboxyl and/or alcoholic groups. After characterising each fraction we studied the role played by these fractions in the soil's capability for retaining heavy metals [copper (Cu) and cadmium(Cd)]. The Cd uptake increases only on the soil treated with sample S1 characterised by a higher pH value and by the presence of high molecular weight fractions. The Cu uptake also increases on the soil treated with sample S2, characterised by the sole presence of low molecular weight fractions. On the other hand, the metal adsorption tests performed on soil treated with the single fractions show that the amount of Cu and Cd retained by soil treated with the high molecular weight fractions of sample does not increase after 72 h of treatment and that the amount of Cu retained by the low molecular weight fractions of sample S1 and by the fractions of sample S2 increases, but does not justify the amount retained by soil treated with the total leachates. # 2001 Elsevier Science Ltd. All rights reserved. Keywords: Land®ll leachate; Molecular weights distribution; Spectroscopic characterisation; Metal soil sorption
1. Introduction Economic considerations continue to maintain land®lls as the most attractive disposal route for municipal solid waste. Alternative methods to land®lling (incineration, composting) are actually considered as volume reduction processes because they produce waste fractions (ashes, slag) which ultimately must be land®lled (Bingener and Crutzen, 1987; Emberton and Parker,
* Corresponding author. Tel./fax: +39-064-991-3723. E-mail address: [email protected] (N. Calace).
1987; Cossu, 1989; Gendebien et al., 1992; Nozhevnikova et al., 1992; Lee et al., 1993). Despite the evolution of land®ll technology from open, uncontrolled dumps to highly engineered facilities designed to eliminate or minimise the potential adverse impact of the waste on the surrounding environment, generation of contaminated leachate remains an inevitable consequence of the practice of waste disposal in land®lls. The subsequent migration of leachate away from land®ll boundaries and its release into the adjacent environment is a serious environmental pollution concern and a threat to public health and safety in both old
0269-7491/01/$ - see front matter # 2001 Elsevier Science Ltd. All rights reserved. PII: S0269-7491(00)00186-X
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and new facilities (MacFarlane et al., 1983; Mackay et al., 1985; Albaiges et al., 1986). Leachate is formed when the refuse moisture content exceeds its ®eld capacity, which is de®ned as the maximum moisture that can be retained in a porous medium without producing downward percolation (Blakey, 1982, 1989; Matejka et al., 1994). Soluble organic and inorganic compounds are encountered in the refuse at emplacement or are formed as a result of chemical and biological processes within the land®ll. Leachate formation creates a non-uniform and intermittent percolation of moisture through the refuse mass which results in the removal of these soluble compounds from the refuse and their dissolution in the leachate (Lema et al., 1988; Lisk, 1991; Wu and Wang, 1992; Oman and Hynning, 1993). The composition of land®ll leachate can exhibit considerable spatial and temporal variations. The great variety of leachate constituents prevents evaluation of the fate and the role played by each component in the environmental impact, so we propose a leachate characterisation based on the distribution of dierent molecular weights, taking into account the fact that the behaviour and reactivity of organic substances may partly depend on their molecular weight or size in solution (Chian, 1977; Chian and De Walle, 1977; Harmesen, 1983; Gourdon et al., 1989; Frimmel and Weis, 1991; Christensen et al., 1998). On the other hand, some researches highlighted the fundamental role played by dissolved organic matter (DOM) in the mobilisation/ immobilisation and transport of pollutants (McCarthy and Zachara, 1989; Harter and Naidu, 1995; Kaiser, 1998; Naidu and Harter, 1998) and some researches pointed out that the DOM interacts intensively with mineral surfaces altering signi®cantly their characteristics (Jardine et al., 1989; Amal et al., 1992; Godtfredsen and Stone, 1994; Evanko and Dzombak, 1998). The aim of this research is, therefore, to study the role played by the dierent molecular weight fractions of leachate constituents in soil sorption of copper (Cu) and cadmium (Cd). For this purpose, leachates were divided into fractions having dierent molecular weight ranges and soil-metal and soil-leachate-metal systems studied in batch conditions. 2. Experimental 2.1. Materials One leachate (S1; old leachate) was collected from a municipal waste land®ll more than 10 years old and another (S2; young leachate) from a municipal waste land®ll established only 4 years ago. The municipal waste land®lls were situated in Pomezia, a small city near Rome (Italy), and the same type of solid waste was
deposited on them. At both sites, the leachates were sampled and frozen at 20 C. S1 sample properties were: pH value of 8.5; chemical oxygen demand (COD) value of 2400 mg l 1; Cu concentration of 10 mg l 1; Lead (Pb) of 32 mg l 1, Zinc (Zn) of 60 mg l 1; nickel (Ni) of 284 mg l 1; and Cd of 0.5 mg l 1. S2 properties were: pH value of 6.4; COD value of 9100 mg l 1; Cu concentration of 3 mg l 1, Pb of 16 mg l 1; Zn of 120 mg l 1; Ni of 55 mg l 1; and Cd of 0.2 mg l 1. The super®cial (0±20 cm) soil features were a pH value of 6.5 (the soil was treated with distilled water, soil:distilled water ratio 1:2.5, for 3 h and then the pH value of the liquid phase was measured), a cation exchange capacity of 9.6 cmol kg 1 of soil, a speci®c surface area of 12 m2 g 1, a content of clay, silt and sand of 22.6, 25.0 and 52.4%, respectively. The organic matter content was 1.6% and the total heavy metal content was Cu 50 mg kg 1, Pb and Zn 90 mg kg 1, Ni 20 mg kg 1 and Cd not detectable. Model Centriplus Amicon concentrators supported low-binding membranes with dierent molecular weight cut-o (100 000, 50 000, 30 000, 10 000 Dalton) were used. Amicon low-binding membranes (molecular weight cut-o of 500 Dalton and diameter of 43 mm) were used in the ultra®ltration system. Cu(II) and Cd(II) solutions (from 0.03 to 0.3 mmol l 1) obtained by dissolving Cu(NO3)2 and Cd(NO3)2 salt in distilled water were employed, then checking the metal concentration with atomic absorption spectroscopy, and the pH values were from 5.7 to 5.4 (0.05). Reagents of analytical grade were used throughout. 2.2. Apparatus Molecular weight fractionation using Amicon concentrators was carried out with a VISMARA centrifuge model 3225R (max. speed 8000 g). The separation of the last two fractions was obtained by using an Amicon stirred ultra®ltration cell, model 8050, capacity 50 ml, supporting the Amicon membranes with a molecular weight cut-o of 500 Dalton. Elemental analyses were carried out in the Microanalysis Laboratory of the Italian Research Council, Monte Libretti, Rome (Italy) using a Carlo Erba 240-B model. 13 C-nuclear magnetic resonance (NMR) spectra were determined in D2O by a Varian spectrometer, model XL300. Samples were prepared by dissolving the dried residue (30 mg) in 1 ml of D2O in an NMR tube (5 mm). The operating conditions were: 75 MHz, pulse 45 , acquisition time 0.1 s, delay time 0.5 s. About 800 000 scans were accumulated. Spectra were performed in broadband decoupling. Under these conditions, a clear spectrum should be obtained with distinct single peaks for each
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carbon (C) atom; because of the complexity of our samples, the spectra show many poorly resolved signals. Fourier transform infrared (FTIR) spectra were recorded by means of a FTIR Philips spectrometer model PU9800 working in diuse re¯ectance conditions. For quantitative determinations, re¯ectance spectra were converted to Kubelka Munk units, which are directly proportional to the concentration of the scattering medium. Samples were prepared by drying 1 ml of liquid sample to which 100 mg of anhydrous KBr had been added; then sample discs were prepared with the dried residue homogenised in a mortar. Metal concentrations were determined by a model Varian AA 10 ¯ame atomic absorption spectrophotometer.
Soil (200 mg) was shaken up with 8 ml Cu(II) or Cd(II) solution of dierent concentrations (from 0.03 to 0.3 mmol l 1) in a controlled temperature box 201 C for 24 h. Preliminary experiments showed no signi®cant variation in metal sorption after 24 h and after batching treatment the pH value of solutions was 7.30.1 for samples S1 and 6.40.1 for samples S2. After centrifugation at 3000 g for 30 min the liquid phases were decanted and the soil washed with deionised water; the washes were then added to the liquid phase and metal concentration determined by atomic adsorption spectroscopy. The amount of sorbed metal was calculated by the dierence between initial and ®nal metal concentration.
2.3. Separation of leachate constituents
3. Results and discussion
The separation of leachate constituents based on the molecular weight size was carried out using Amicon concentrators designed for use in centrifuges with ®xedangle rotors (34 ) or a swinging bucket. The spin time was optimised for each cut o (100, 50, 70, 190 min, respectively, centrifugal force 3000 g) in order to obtain a ®nal retentate volume of 0.5 ml. Initial leachate volume used was 10 ml. The last fraction obtained with concentrators was further divided in a stirred Amicon ultra®ltration cell connected to a reservoir (5 l) containing dia®ltrate solution (deionized water) pressurised using nitrogen (3 atm). Since concentrator membranes contain trace amounts of glycerine, before using the membranes, washing was performed by centrifugation with 0.1 M NaOH (three times for 1 h) and the residual NaOH subsequently removed by washing with deionised water until a neutral pH value was obtained. After separation, each leachate fraction was again made up to initial volume (10 ml). A portion of each solution was dried under an infrared lamp (60 C), weighed and the obtained dried residue used for analytical characterisation (FTIR, 13C-NMR, elemental analysis).
3.1. Characterisation of molecular weight fractions Molecular weight distributions of leachate constituents are shown in Fig. 1. In the sample S1 (old leachate), the constituents were distributed over a wider range of molecular weights; high molecular weight fractions were present. The data con®rm the results obtained concerning the presence of humic-like substances (Artiola-Fortuny and Fuller, 1982; Weis et al.,
2.4. Adsorption isotherms Metal sorption experiments were carried out in batch conditions on the untreated soil and on the soil treated both with leachates and with the dierent leachate fractions. The treated soil was obtained by batching the untreated soil with leachates or with the dierent leachate fractions for 48 h (time experimentally determined by means of kinetic studies of sorption process, soil± liquid phase ratio 1:2). After centrifugation, the soil was washed with deionised water, lyophilised and used for the sorption experiments.
Fig. 1. The molecular weight fractions distribution of sample S1 (old leachate) and sample S2 (young leachate).
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1989; Calace and Petronio, 1997). In sample S2 (young leachate), the fractions are actually narrower at the lower molecular weights, in agreement with literature data (Chian, 1977; Chian and De Walle, 1977; Frimmel and Weis, 1991) for which the molecular weight distribution over a wider range increases with the increasing age of land®ll leachates. The elemental compositions of fractions are given in Table 1. In sample S1, the percentage of C was found to be highest in the 500±10 000 Dalton fraction and lowest in the 30 000±50 000 Dalton fraction. The nitrogen (N)/ C ratio was the same for the dierent molecular weight constituents while that of sulphur (ratio S/C) was the same except in the 500±10 000 Dalton fraction where it was lower. As regards the ratio hydrogen (H)/C data, the fractions with lower molecular weights had the highest values. Hence, the elemental composition of sample S1 fractions highlights the fact that there is not much dierence between them. In sample S2, the elemental composition is the same for each molecular weight fraction, which are all characterised by high H/C values; N and S content is negligible. Taking into account the data obtained, we performed some spectroscopic analyses to estimate the nature and kind of the functional groups characterising the dierent fractions. Indeed, the nature of the functional groups forming the compounds ultimately in¯uences the characteristics of the compound and can thus contribute greatly to the determination of mechanisms of metal accumulation and/or metal transport in soils. For the interpretation of land®ll leachate fraction spectra, we based on literature data referred to humic substances (Artiola-Fortuny and Fuller, 1982; Harmesen, 1983; Calace and Petronio, 1997) which are natural macromolecules having complex structures; any papers on land®ll leachate spectroscopic analysis are not found in the literature.
As regards sample S1, the spectroscopic analyses performed on each fraction highlighted some structural dierences. The FTIR spectrum of the >100 000 Dalton fraction (Fig. 2a) shows signals at 2916 and 2857 cm 1 which are attributed to asymmetric and symmetric stretching of methyl and methylene groups, as well as a band at 1406 cm 1 due to bending of CH2 adjacent to CO oxygen ( CH2 CO). Moreover, signals at 1661 and 1559 cm 1 due to the ®rst and second band of amides (stretching of CO and C N bonds, respectively) and a sorption at 1050 cm 1 associated with the C OH bond in secondary alcohols were present. 13CNMR spectrum (Fig. 3a) displayed a broad resonance band in the region 0±30 ppm due to aliphatic carbons of branched chains, a signal at 40 ppm due to the C N bond; poorly resolved signals are also present in the 60± 80 ppm region associated with C bound to O atoms, and in the 80±140 ppm region due to ole®nic carbons. The presence of CO groups is highlighted by resonances around 170 ppm. The spectroscopic features of the 50 0000-100 000 Dalton and 10 000-30 000 Dalton fractions are similar to the >100 000 Dalton one. Only the ratios among the intensities of sorption change. The spectroscopic results of 30 000±50 000 Dalton and 500±10 000 Dalton fractions are similar. The FTIR spectrum of the latter (Fig. 2b) shows a shoulder at 1670 cm 1 that, with the signals at 2990, 1600 and 1420 cm 1, can be attributed to CC. Moreover, absorption in the region of 1100 cm 1 is typical of stretching C O bonds in alcohols and ethers. 13C-NMR spectrum (Fig. 3b) shows signals in the 0±50 ppm region associated with the presence of aliphatic carbons in more or less
Table 1 Elemental analysis data of dierent molecular weight fractions of leachatesa C% N% H% O%
S% N/C O/C S/C
H/C
Sample S1 >100 000 D 50 000±100 000 D 30 000±50 000 D 10 000±30 000 D 500±10 000 D <500 D
45.0 47.5 36.2 44.4 54.2 44.1
1.07 1.25 0.94 1.20 0.58 1.67
1.52 1.57 1.76 1.40 1.95 2.58
Sample S2 10 000±30 000 D 500±10 000 D <500 D
53.8 0.45 7.70 37.60 0.45 0.007 0.52 0.007 1.72 53.6 0.45 8.80 36.70 0.45 0.007 0.51 0.007 1.97 57.2 0.48 8.95 33.07 0.32 0.007 0.89 0.002 1.88
a
7.14 6.70 5.30 5.60 6.00 3.72
5.70 6.20 5.30 5.20 8.80 9.50
41.09 38.35 52.26 43.60 30.42 41.01
0.14 0.12 0.12 0.11 0.09 0.07
0.68 0.60 1.08 0.74 0.42 0.70
0.009 0.010 0.010 0.010 0.004 0.014
The data are the average of ®ve determinations; standard deviation is less than 3%.
Fig. 2. Fourier transform infrared spectra of sample S1 (old leachate) fractions: (a) >100 000 D fraction; (b) 500±10000 D; (c) <500 D.
N. Calace et al. / Environmental Pollution 113 (2001) 331±339
Fig. 3.
13
335
C-NMR spectra of sample S1 (old leachate) fractions: (a) >10 000 D fraction; (b) 500±10000 D.
branched chains and of methoxyl groups, signals around 70 ppm due to alcohols and ethers, resonances at 129 ppm due to the CC bond and ®nally, signals around 180 ppm attributed to CO bond of carbonyl groups. Finally, the FTIR spectrum of the <500 Dalton fraction (Fig. 2c) highlight its prevalently aliphatic nature and the presence of a high inorganic ion content in the form of carbonates and nitrates (FTIR sorptions at 1460 and 1372 cm 1). In sample S2, the spectroscopic analyses of each fraction con®rm the homogeneity in the functional group distribution. The FTIR spectra (Fig. 4) show predominant signals at 1580 and 1440 cm 1 due to stretching and bending of CO in carboxylated groups (COO ) and two absorption bands at 2960 and 2900
cm 1 due to CH2 in long chains of fatty acids. Also 13CNMR spectra (not shown) show a large number of signals due to aliphatic carbons (10±40 ppm) and a resonance at 180 ppm due to CO bonds. 3.2. Metal sorption The adsorption isotherms, a graphical representation of the relationship between the equilibrium concentration of adsorbate in solution and the equilibrium concentration of adsorbate adsorbed by the soil particles, are shown in Figs. 5 and 6. These isotherms are obtained from batch tests on untreated soil and on the soil-leachates system. Cu is already sorbed to a greater extent than Cd onto the original soil in agreement with literature data (Yong et al., 1992). The Cu, indeed, has a
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higher anity for soil surface sorbing sites than the Cd (Yong et al., 1992). After the sample S1 treatments, both Cu and Cd were retained by the soil increases and the increase was not the same for two metals. This result is probably due to the dierent features of metals, such as anity to organic matter, formation of inner- and/or outer-sphere complexes, etc., as well as the structural features of the soil surface which changes with the leachate treatment. Numerous investigators (Tiller et
Fig. 4. Fourier transform infrared spectra of sample S2 (young leachate) fractions: (a) >10 000 D fraction; (b) 500±10000 D; (c) <500 D.
al., 1984; Brummer et al., 1988; Gerth et al., 1993; Naidu et al., 1994, 1997) showed that soil solution pH has a marked in¯uence on adsorption of metals that increases when soil solution pH increases and, between Cd and Cu, the former is more in¯uenced by pH value changes (Naidu et al., 1994). On the other hand, the Cu can interact, forming stable complexes with high and low molecular weight organic substances while the Cd can only interact with the high molecular weight organic compounds (Godtfredsen and Stone, 1994; Naidu et al., 1997). Sample S1 is characterised by a high pH value and by the presence of both high molecular weight organic substances and low molecular weight organic substances. In general, it can be hypothesised that the sample S1 in¯uences the soil sorption capability both acting on pH value by increasing it and creating new sorbing sites due to the presence of organic matter. This hypothesis is con®rmed by the tests performed with the sample S2, characterised by pH value similar to pH value of original soil and by the presence of the only low molecular weight organic compounds. After treatment with the sample S2, the soil retention capability only increases for Cu. Our results ®t the Langmuir equation and the experimental data have an H-type pattern according to Giles' classi®cation (Giles et al., 1974a, b). These curves show a long plateau and are characteristic of interactions that are essentially electrostatic in nature between species adsorbing and soil active surface. The length of the plateau indicates the diculty of formation of a second layer because of charge repulsion between the adsorbates and those approaching from the solution. Moreover, some authors have demonstrated that sorption of
Fig. 5. Graphical representation of copper sorption. S1, old leachate; S2, young leachate.
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Fig. 6. Graphical representation of cadmium sorption. S1, old leachate; S2, young leachate.
Cu and Cd onto clay minerals follows Freundlich isotherms, whereas sorption onto humic substances follows the Langmuir isotherms (Yong et al., 1992) con®rming the hypothesis about the important role played by high molecular weight organic matter in heavy metal sorption process. The soil treated with each fraction of sample S1 (Table 2) shows a dierent pattern. The metal adsorption tests performed on soil treated with the single fractions show that the amount of Cu and Cd retained by soil treated with the high molecular weight fractions of sample does not increase after 72 h of treatment. Indeed, the soil treated with the high molecular weight fractions does not change the retention capacity of either Cu or Cd. After treatment with low molecular weight fractions, on the other hand, only the amount of retained Cu is weakly increased. Moreover, the amount of Cu sorbed onto the soil treated with low molecular weight fractions does not justify the value of qmax found for the soil treated with the total sample S1. In this case, the pH value increase of soil does not seem to in¯uence the metal retention capacity (the pH values of fractions were 8.50.1) but, on the basis of the data obtained, it seems that the increase of the metal retention capacity of the soil treated with the total leachate is not due to a particular fraction of leachate but to the presence of all fractions at the same time. We have already said that after treatment with sample S2, the soil increases metal retention capacity only in the case of Cu. The Cd sorption on soil, indeed, does not seem to be in¯uenced by the presence of sample S2.
Table 2 The maximum amount of metals retained from soilsa Cu qmax
Cd qmax
Original soil
4.37
1.57
S1 Total S1 >100 000 D S1 50 000±100 000 D S1 30 000±50 000 D S1 10 000±30 000 D S1 500±10 000 D S1 <500 D
6.81 4.39 4.31 4.06 3.81 4.73 4.88
3.65 1.81 1.70 1.71 1.59 1.69 1.66
S2 Total S2 10 000±30000 D S2 500±10 000 D S2 <500 D
6.02 5.20 4.95 5.29
1.67 1.34 1.33 1.42
a The data are the average of ®ve determinations; standard deviation is less than 5%.
Even treating the soil with each fraction increases Cu retention while Cd retention actually decreases (Table 2). 4. Conclusion Since the composition of land®ll leachate can exhibit considerable spatial and temporal variations, an interesting approach to characterise the municipal land®ll leachates consists of the separation of leachate constituents in dierent molecular weight fractions. In particular, the old leachate constituents were distributed
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over a wider range of molecular weights; high molecular weight fractions were present. The young leachate fractions, instead, are actually narrower at the lower molecular weights, in agreement with literature data. Moreover, the nature of the functional groups forming the compounds can in¯uence the characteristics of the compound and can thus contribute greatly to the determination of mechanisms of metal accumulation and/or metal transport in soils. The high molecular weight fractions of old leachates are found to be complex structures formed by condensed nuclei of carbons substituted by functional groups containing N, S and O atoms; the low molecular weight fractions of leachates are, instead, characterised by linear chains substituted by oxygenated functional groups such as carboxyl and/ or alcoholic groups. The impact of the land®ll leachates on the environment can be estimated on the basis of its chemical characteristics (pH value, type and content of organic matter, etc.) and an evaluation of the role played by leachate constituents on the fate of metals in the soils can be done. We found that the Cd uptake increases only on the soil treated with sample S1, characterised by a higher pH value and by the presence of high molecular weight fractions. The Cu uptake also increases on the soil treated with sample S2, characterised by the sole presence of low molecular weight fractions. On the other hand, the metal adsorption tests performed on soil treated with the single fractions show that the amount of Cu and Cd retained by soil treated with the high molecular weight fractions of sample does not increase after 72 h of treatment and that the amount of Cu retained by the low molecular weight fractions of sample S1 and by the fractions of sample S2 increases, but does not justify the amount retained by soil treated with the total leachates. Taking into account that the leachate±soil interactions depend on the nature of soil type and on the chemical features of leachate constituents, we can hence hypothesise that leachates are able to increase the metal retention capacity of the soil but that this ability does not depend on a particular fraction of leachate, but on the presence of all fractions at the same time. References Albaiges, J., Casado, F., Ventura, F., 1986. Organic indicators of groundwater pollution by a sanitary land®ll. Water Research 20, 1153±1159. Amal, R., Raper, J.A., Waite, T.D., 1992. Eect of fulvic acid adsorption on the aggregation kinetics and structure of hematite particles. Journal of Colloid and Interface Science 151, 244±257. Artiola-Fortuny, J., Fuller, W.H., 1982. Humic substances in land®ll leachates: I. Humic acid extraction and identi®cation. Journal of Environmental Quality 11 (4), 663±669. Bingener, H.G., Crutzen, P.J., 1987. The production of methane from solid wastes. Journal of Geophysical Research 92, 2182±2187.
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