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Comparative study of the removal of phenolic compounds by biological and non-biological adsorbents
Journal of Hazardous Materials 164 (2009) 1439–1446
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Comparative study of the removal of phenolic compounds by biological and non-biological adsorbents Abel E. Navarro a,∗ , Norma A. Cuizano b , Jose C. Lazo c , María R. Sun-Kou c , Bertha P. Llanos b a
Department of Chemistry, Graduate School of Arts and Sciences, New York University, New York, NY 10003, United States Departamento Académico de Química, Facultad de Ciencias y Filosofía, Universidad Peruana Cayetano Heredia, Lima L31, Peru c Sección Química, Departamento de Ciencias, Pontificia Universidad Católica del Perú, Lima L32, Peru b
a r t i c l e
i n f o
Article history: Received 10 February 2008 Received in revised form 10 September 2008 Accepted 17 September 2008 Available online 27 September 2008 Keywords: Exchanged bentonite Cross-linked algae Partition coefficient pH Adsorption
a b s t r a c t The ability of biological and non-biological adsorbents to remove 2-nitrophenol (2-NP) and 2chlorophenol (2-CP) from aqueous solutions in batch experiments at room temperature was compared. The marine seaweeds Macrocystis integrifolia Bory (S1) and Lessonia nigrescens Bory (S2) were crosslinked with CaCl2 to enhance their mechanical properties. Natural bentonite was chemically exchanged with hexadecyltrimethylammonium bromide (B1) and bencyltriethylammonium chloride (B2) to increase their affinity towards organic compounds as well. The adsorption capacity of all of the adsorbents strongly depends on solution pH, whereas equilibrium assays showed a mixed mechanism according to the Langmuir and Freundlich isotherms. The maximum adsorption capacity of 2-NP follows the trend: S1 > S2 > B2 > B1 within the range of 97.37 and 18.64 mg g−1 whereas for 2-CP, it ranged between 24.18 and 9.95 mg g−1 with the trend: S1 > S2 > B2 ≈ B1. The importance of the octanol–water partition coefficient as the main factor on the adsorption of these compounds on two different kinds of adsorbents is discussed. © 2008 Elsevier B.V. All rights reserved.
1. Introduction At present, one of the most serious concerns faced in the natural environment is chemical contamination with organic and inorganic substances catalyzed by the presence of heavy metals and phenolic compounds [1,2]. The presence of such compounds in an aquatic medium originates changes in chemical (pH, chemical oxygen demand, alkalinity, acidity, dissolved oxygen, etc.), physical (color, temperature, odor, viscosity, turbidity, etc.) and biological properties, damaging the water quality for human usage and upsetting the environmental equilibrium [2,3]. The uptake of toxic and heavy metals ions from waste waters and industrial wastes by certain types of microbial biomass especially algae [4,5], fungi [6,7] and other biological adsorbents [8,9] has been the subject of many recent studies. Natural clays have also shown an increasing number of applications on the removal of heavy metals [10–12]. Among the different pollutants of aquatic systems, phenols are considered as priority contaminants since they are harmful to plants, animals and humans, even at low concentrations [3,13]. Due to their extensive use and slow degradation, nitrated and chlorinated phenols readily enter the environment by agricultural runoff
∗ Corresponding author. Tel.: +1 646 2862082; fax: +1 212 9954475. E-mail address: [email protected] (A.E. Navarro). 0304-3894/$ – see front matter © 2008 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2008.09.077
of pesticides, effluents from oil refineries and plastic industries and by leachates emerging from waste deposits produced by microbial hydrolysis and photodegradation of several organophosphorous pesticides, such as parathion [14,15]. Activated carbon is one of the most efficient adsorbents for these organic molecules, as it possesses a high surface area per unit mass and exhibits a high affinity for phenolic compounds [16]. Unfortunately, due to the high cost of activated carbon, it is not an economical adsorbent compared with low cost, naturally occurring alternatives. These non-expensive adsorbents remove trace amounts of organic contaminants from wastewaters where other sophisticated techniques such as ionic exchange, osmosis, and solvent extraction are not applicable. Consequently, the environmental biotechnology is in a constant search of alternative, less expensive and viable techniques for the removal of these phenolic compounds at very low concentrations. Adsorption using naturally occurring adsorbents has become a cheap and efficient tool, easily applicable in the detoxification of residual waters from industries and mines showing positive results. High efficiency on the elimination of artificial colorants and phenols from aqueous solutions by marine seaweeds and natural clays has been observed in previous studies [17–22]. Peru is widely known for its diversity of natural resources. Marine seaweeds are in such abundance on the shores of beaches as to cause unsightly accumulation. Also, natural clays of the type
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montmorillonite, saponite and kaolinite are extracted from natural mines in the northern part of the country at very low price. For reasons of accessibility and abundance, Peruvian algae and natural clay of the type bentonite were chosen and compared for the removal of 2-nitrophenol and 2-chlorophenol from aqueous solutions. Several research groups have investigated the adsorptive capacity of other adsorbents such as lignins, coconut shells and papermill sludges [15,23,24] in the removal of nitrated and chlorinated phenols. Previous work has already been carried out utilizing marine algae and clays for the elimination of phenolic compounds [17,19] but none have compared the adsorption capacity of these two adsorbents or explained a relationship between their chemical and mechanical properties and adsorption. 2. Materials and methods 2.1. Biological adsorbent – algae The marine seaweeds were obtained from the beaches of Marcona and Tacna in Peru and identified as Macrocystis integrifolia Bory (S1) and Lessonia nigrescens Bory (S2), respectively. After collection, both brown marine algae were washed twice with tap water in order to remove adhering soil, other microscopic algae, insect larvae, etc. In the laboratory, they were washed with double-distilled water and dried in sunlight and then in an oven at 60 ◦ C for 24 h until all the moisture was evaporated. The dried seaweeds were ground and separated by size by means of sieves reaching a particle size between 150 and 212 m. Rubin et al. [25] indicated that untreated Sargassum muticum was easily decomposed in presence of phenolic compounds. Moreover, the decomposition products interfere in the absorbance measurements and consequently, in the determination of phenolic compounds in aqueous solution. Therefore, to enhance their mechanical properties, the algae were cross-linked with calcium chloride following the procedure cited by Blanco et al. [4], vacuum filtered and washed with double-distilled water to remove the excess of CaCl2 . Finally, the seaweeds were dried at 60 ◦ C and stored under refrigeration until their use. 2.2. Non-biological adsorbent – clays The natural clays used in this study were obtained from Piura, Peru and classified as bentonite. In order to separate the pure montmorillonite fraction from any adhering materials the clay samples were purified by controlled sedimentation. The samples were left to settle, decanted and dried at 60 ◦ C in an oven for 24 h. Finally, the bentonite was passed through a 2 m sieve and taken for further treatment. The bentonite was chemically modified with two different quaternary ammonium salts: bencyltriethylammonium bromide (BTEA) and hexadecyltrimethylammonium chloride (HDMTA), which have shown high affinity towards organic compounds [20,22], to enhance their adsorption capacity. 20 g of bentonite suspended in 2 L of distilled water were combined with two equivalents of its cation exchange capacity (CEC) of the ammonium salts (99% Sigma–Aldrich). The final suspension was stirred at 250 rpm for 12 h at room temperature. The modified bentonites were vacuum filtered and vigorously rinsed with double-distilled water to eliminate the excess of ammonium salts, dried at 60 ◦ C in an oven for 48 h and stored until their use. The bentonites were identified according to the ammonium salt used in their modification as follows: bentonite treated with HDMTA (B1) and bentonite treated with BTEA (B2).
The CEC of the bentonite (0.06 mmol g−1 ) was determined using the methylene blue standard method (ASTM C-837-81). 2.3. Characterization of adsorbents The specific surface area of both types of adsorbents was measured. For the algae, the methylene blue adsorption method was used. The calculation was made assuming a value of 108 (Å)2 for the ionic cross-sectional area of methylene blue [26]. The surface area of the modified bentonites was determined on the basis of the Brunauer, Emmet and Teller (BET) method using an ASAP 2000 (Micromeritics). Infrared spectra of all the adsorbents were recorded on a PerkinElmer 1600 infrared spectrometer in order to identify the binding groups present in their surface. FT-IR samples were prepared by combining 0.5 g of powered potassium bromide with 1 mg of dried adsorbent 24 h prior to recording. Thermal stability and presence of volatile impurities in the adsorbents were analyzed by thermogravimetric studies using a Linseis STA 1600 analyzer at a heating rate of 5 ◦ C/min up to 1000 ◦ C. In order to compare the functional groups observed by FT-IR, elemental composition of the adsorbents was evaluated by standard methods and ICP multi-elemental analysis. An Environmental Scanning Electron Microscope (PHILIPS ESEM XL-30 TMP) was selected to elucidate the surface texture and morphology of the adsorbents. Prior to the observation, the surface of the samples was coated with a thin and electrically conductive gold film. 2.4. Batch adsorption test Duplicated batch adsorption experiments were carried out at room temperature by allowing an accurately weighed amount of cross-linked algae and exchanged bentonites to reach equilibrium with 2-nitrophenol (2-NP) and 2-chlorophenol (2-CP) solutions of known concentrations. Initial concentrations of 2-NP and 2-CP were held between 50 and 1000 mg/L. The pH was adjusted by adding concentrated solutions of HCl and NaOH. The bentonite samples were shaken at 500 rpm for 6 h, as established in preliminary results [22] and the algal samples were shaken at 200 rpm during 24 h to assure the completeness of the adsorption. At the end of the equilibrium the samples were filtered and subsequently analyzed for residual concentration of 2-NP and 2-CP by UV–vis spectrometry (SHIMADZU UV-MINI 1240) at wavelengths of 274 and 276 nm for 2-NP and 2-CP, respectively. 2.5. Data analysis The amount of adsorbed phenolic compound is expressed as adsorption capacity (q, mg/g) calculated as follows: q=
(Ci − Cf ) × V m
(1)
where m is the mass of adsorbent expressed in g, V is the volume of the solution in L and Ci and Cf are the initial and at the equilibrium concentrations, respectively expressed in mg/L. 3. Results and discussion 3.1. Adsorbent characterization 3.1.1. Specific surface area Specific surface areas of 820 and 1512 m2 g−1 for the brown algae S1 and S2, respectively were observed by means of the methylene blue method in the water-wet state of the algae. On other
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Fig. 1. FTIR spectra of cross-linked algae: Macrocystis integrifolia Bory (S1) and Lessonia nigrescens Bory (S2).
hand, bentonite samples B1 and B2 showed a surface area of 24 and 34 m2 g−1 with the BET method (adsorption and desorption of nitrogen). These algae report higher surface areas and demonstrate higher porosity and a better distribution of active sites on their surface when compared to the algae Gelidium (375 m2 g−1 ) [21] and S. muticum (485 m2 g−1 ) [26] as well as other naturally occurring adsorbents such as lignin (19.72 m2 g−1 ) [15] and dried sewage sludge [27]. The bentonites, upon treatment with the ammonium salts underwent a decrease in the surface area compared to other natural (66 m2 g−1 ) [18] and modified clays such as porous clay heterostructure (305.5 m2 g−1 ) [28]. This loss of surface area is attributed to the full occupation of the interlamellar spaces of the bentonite by organic cations containing a large hydrophobic alkyl chain (HDMTA) and aromatic ring (BTEA), slightly reducing their porosity. The substantial difference in the specific surface areas between these two types of adsorbents is explained by the use of different methods in their determination. It is known that in aqueous solution, the algae swell and increase their dimensions because of the
water filling their pores [25], whereas these changes in the volume of organophilic clays are rarely seen. It has been proposed that the surface area of algae is more accurately determined by the methylene blue method, where the adsorbent is better resembled at the moment of the adsorption. On the other hand, organophilic clays are almost unaffected by the presence of water so the BET method is considered more accurate than other techniques. 3.1.2. Fourier transform infrared spectroscopy and composition The FT-IR revealed the presence of the typical functional groups in algae (Fig. 1) as well as the interlamellar insertion of organic cations in the clay’s structure (Fig. 2). As previously shown [29], the algae S1 and S2 display peaks at 3410, 1634 and 1425 and at 1160 and 817 cm−1 attributed to the presence of polyalginate and fucoidan, respectively. As for the modified bentonites (Fig. 2), common peaks can be found at 1636 and 3445 cm−1 related to the vibration of O–H bonds [30,31] that belongs to remaining water molecules physically adsorbed on the clays. Stretching vibrations Si–O (1040 cm−1 ) and Si–O–Si (570 cm−1 ) are also observed [30].
Fig. 2. FTIR spectra of bentonites exchanged with hexadecyltrimethylammonium chloride (B1) and with bencyltriethylammonium bromide (B2).
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Table 1 Elemental composition of the marine seaweeds. Composition (%)
S1
S2
N Ca P Fe Fiber
1.81 5.51 1.89 1.11 11.1
4.12 5.46 1.38 0.15 10.7
Table 2 Elemental composition of the modified bentonites. Composition (%)
B1
B2
Na Mg Al2 O3 SiO2 K Ca Ti Mn Fe
0.04 0.38 11.6 62.3 0.05 0.48 0.01 0.01 0.56
0.05 0.42 11.9 64.5 0.05 0.38 0.01 0.01 0.69
The only peculiar difference between B1 and B2 FT-IR spectra (Fig. 2) is two sharp peaks in B1 around 2850–2920 cm−1 associated with asymmetric and symmetric stretching of C–C bonds due to the length of the aliphatic chain of HDMTA [32]. To corroborate these results Tables 1 and 2 show the elemental composition of the adsorbents, exhibiting in the algae a high content of calcium due to the cross-linkage and fibers assumed to correspond to the polyalginates and fucoidans [5,29,31]. According to the ICP multi-element analysis of the exchanged bentonites, silica and alumina are the main constituents as reflected in the FT-IR analysis. 3.1.3. Thermogravimetric analysis Figs. 3 and 4 show the thermogravimetric analysis of crosslinked algae and modified bentonite, respectively. The thermal studies of algae showed that both algae underwent three steps decomposition process when heated from 20 to 550 ◦ C. Initial step (∼4% weight loss), which was small in the range in 25–110 ◦ C, could be attributed to the loss of adsorbed water [33]. The second step from 240 to 320 ◦ C (∼6% weight loss) corresponds to volatile substances releases proceeding from inorganic compound decomposition as well as organic matter degradation. At this range of
Fig. 3. Thermogravimetric curves of cross-linked S1 and S2 seaweeds.
Fig. 4. Thermogravimetric curves of exchanged bentonites B1 and B2.
temperature some carbohydrates degrade (around 300 ◦ C). Finally, the weight loss at temperatures higher than 480 ◦ C is caused by fiber and lignin pyrolysis [34]. As for the modified bentonites, B1 shows a two-steps decomposition process when heated from 20 to 500 ◦ C. The first weight loss of 1.5% occurs between 25 and 100 ◦ C, also associated with the loss of adsorbed water. The second step (∼5% weight loss) occurs in the range between 150 and 300 ◦ C, corresponding to two events: (i) the water loss associated with the residual exchange cations that remain in the bentonite upon the treatment with the ammonium salt or (ii) the early decomposition of the ammonium salt that has been inserted in the bentonite [35,36]. Finally, a third step between 300 and 550 ◦ C, corresponds to structural dehydroxylation of the bentonite (∼1% weight loss). This dehydroxylation was also observed by Xi et al. in organophilic bentonites, and continues until 636 ◦ C [37]. On the other hand, the modified bentonite B2, shows a thermal curve similar to B1, but the weight loss percentages are lower for B2. Since, the water loss in the first step is the same for B1 and B2; we believe that a lower exchange with the ammonium salt BTEA has happened. The difference in this initial loss of water between both types of adsorbents is attributed to the hygroscopicity of the polysaccharic composition of the algae compared to the hydrophobicity of modified bentonites. These results show that the cross-linkage of algae with calcium and the incorporation of quaternary ammonium ions in bentonites show mechanical stability and heat resistance up to 300 ◦ C compared to other adsorbents where oxidative decomposition has been observed at lower temperatures [33,34]. 3.1.4. Scanning electron microscopy The electron micrograph scans clearly reveal the surface texture and morphologies of the modified adsorbents (Figs. 5 and 6). It is evident from the micrographs that both algae show a very well-defined organization of the polysaccharic structures in their surfaces where there is an even surface texture. Contradictorily, both exchanged bentonites showed a repetitive formation of “layer over layer” structure. Moreover, the micrograph of B2 displays layers on the clays that have become more planar and irregular than the ones observed in B1. This can be attributed to a less efficient packing of aromatic organic ions BTEA compared to the aliphatic organic ions HDMTA between the interlamellar spaces of the clay.
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Fig. 5. SEM images showing the surface morphologies of S1 (left) and S2 (right) at ×1280 magnification.
Fig. 6. SEM images showing the surface morphologies of B1 (left) and B2 (right) at ×6925 magnification.
3.2. Sorption studies 3.2.1. Effect of initial pH value pH is one of the most important factors in the adsorption of environmental contaminants in aqueous solutions since it influences the solution chemistry of phenols (i.e. hydrolysis, redox reactions, polymerization and complexation). pH also has a strong influence on the speciation and the sorption availability of phenolic compounds as well as the ionic state of functional groups on the surface of adsorbents [7,8,29]. Fig. 7 shows the effect of solution pH on the adsorption of 2NP at room temperature. For algae, the highest adsorption capacity
Fig. 7. Effect of pH on 2-NP adsorption by S1 (), S2 ((), B1 (() and B2 (). Initial conditions: 100 mg of adsorbent at initial concentration of 2-NP of 100 mg L−1 .
is reported at low pH (q = 6.2 and 7.1 mg g−1 for S1 and S2 at pH 4 and 3, respectively). Contradictorily for exchanged bentonites show an adsorption capacity that gradually increases with the pH (q = 8.8 mg g−1 for B1 at pH 10), whereas for B2, there is not observable pH effect within the range values of 2–10, reaching a steady adsorption capacity around 2.0 mg g−1 . These phenomena can be explained as the product of electrostatic repulsion/attraction. Recent studies [29] demonstrated that the algae S1 and S2 have apparent ionization constants around 3.0 where at higher pH values, the algae’s surfaces are negatively charged. At the same time, 2-NP (pKa = 7.23) becomes negatively charged with increasing pH. Consequently, the neutral state of 2-NP promotes its adsorption on cross-linked seaweeds. On the other hand, the exchange of ions with HDMTA cations in bentonites creates a permanent and non-pH-dependent positive charge evenly distributed on the surface, shown by SEM analysis, that preferentially binds to negatively charged 2-nitrophenolate which exists in higher concentrations at high pH values. Likewise, as stated in Section 3.1.4, the micrographs revealed a scattered packing of BTEA in the clay B2, which added to the high hydrophobicity of the aromatic ring, reduces the pH effect in the process. The relationship between 2-CP uptake and solution pH for crosslinked algae and exchanged bentonite is illustrated in Fig. 8. The graph clearly exhibits that lower pH values led to higher 2-CP uptake for both kind of adsorbents (q = 10.5, 5.0, 18 and 12 mg g−1 at pH values 2, 3, 4 and 4 for S1, S2, B1 and B2, respectively). According to their octanol–water partition coefficients, 2-NP is more polar than 2-CP (log Kow = 1.77 and 2.15, respectively) [38,39], therefore 2-CP (pKa = 8.48) does not have a distribution of electron density as expanded as 2-NP and will display a better adsorption on neutral surfaces. Algae corroborate this behavior by displaying their higher adsorption capacity at pH lower than their respective apparent ionization constants [29]. In the case of exchanged
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Fig. 8. Effect of pH on 2-CP adsorption by S1 (), S2 ((), B1 (() and B2 (). Initial conditions: 100 mg of adsorbent in 100 mL of solution at initial concentration of 2-CP of 100 mg L−1 .
Fig. 9. Linearized Langmuir isotherms obtained from 2-NP adsorption by S1 (), S2 ((), B1 (() and B2 (). Initial conditions: 500 mg of adsorbent in 100 mL of solution at the optimum pH.
bentonites, the decrease on pH makes 2-CP more neutral and friendlier to the aliphatic (HDMTA) and aromatic (BTEA) quaternary ammonium ions. A slight decrease on the adsorption is observed at pH values lower than 4 which can be attributed to the polarization of 2-CP by the formation of complexes or protonation of the hydroxyl group in excess of hydronium ions. Exchanged bentonites also show an increasing adsorption at high pH values (pH ≈ 10) most likely due to the electrostatic affinity between the deprotonated 2-CP and the positive charge of the ammonium ions. Based on these results, we propose a completely electrostatic adsorption of 2-NP and 2-CP by cross-linked algae and a polarity-dependent adsorption for bentonites (apolar adsorbates are adsorbed by apolar mechanisms and polar adsorbates by polar mechanisms) obeying the Traube rule and Rebinder effect. 3.2.2. Equilibrium modeling The adsorption of a substance from one phase to the surface of another in a specific system leads to a thermodynamically defined distribution of that substance between the phases as the system reaches the equilibrium. This distribution can be expressed in terms of adsorption isotherms. The Langmuir equation assumes that (i) the solid surface presents a finite number of identical sites which are energetically uniform; (ii) there is no interaction between adsorbed species, meaning that the amount adsorbed has no influence on the rate of adsorption and (iii) a monolayer is formed when the solid surface reaches the saturation. While the Langmuir isotherm assumes that enthalpy of adsorption does not depend on the amount adsorbed, the Freundlich isotherm is derived assuming a logarithmic decrease in the enthalpy of adsorption with the increase in the fraction of occupied sites and based on sorption on heterogeneous surfaces. Both isotherm equations can be described by their linearized forms and the Langmuir and Freundlich constants can be determined, according to Eqs. (2) and (3), respectively. 1 1 1 = + × qmax × Cf q qmax b
(2)
ln q = ln kf +
(3)
1 n
× ln Cf
where qmax (mg g−1 ) and b (L mg−1 ) are the Langmuir constants related to the maximum adsorption capacity and to the adsorption
Fig. 10. Linearized Freundlich isotherms obtained from 2-NP adsorption by S1 (), S2 ((), B1 (() and B2 (). Initial conditions: 500 mg of adsorbent in 100 mL of solution at the optimum pH.
energy, respectively and kf and 1/n are the Freundlich constants related to the adsorption capacity and the adsorption intensity, respectively. The linearized forms of Langmuir and Freundlich isotherms obtained at room temperature for the adsorption of 2-NP and 2-CP are given in Figs. 9–12, whereas Table 3 presents the corresponTable 3 Isotherm constants for adsorption of phenolic compounds (2-NP and 2-CP) by crosslinked algae and modified bentonites. Adsorbent
Langmuir
Freundlich
qmax (mg g−1 )
b (L mg−1 )
R2
kf
n
R2
2-NP S1 S2 B1 B2
97.37 71.28 18.64 23.02
0.0005 0.0005 0.0316 0.001
0.937 0.965 0.988 0.99
0.035 0.02 1.824 0.068
0.978 0.973 2.358 1.31
0.976 0.977 0.984 0.987
2-CP S1 S2 B1 B2
24.18 17.33 9.95 10.04
0.0135 0.0363 0.0087 0.0086
0.995 0.986 0.988 0.989
1.391 3.297 0.916 0.566
2.269 3.714 2.841 2.302
0.984 0.946 0.992 0.961
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Fig. 11. Linearized Langmuir isotherms obtained from 2-CP adsorption by S1 (), S2 ((), B1 (() and B2 (). Initial conditions: 500 mg of adsorbent in 100 mL of solution at the optimum pH.
dent constants along with the coefficients of linear correlation (R2 ) associated with each linearized model. As can be seen in Table 3, the maximum value for the adsorption capacity for 2-NP was obtained by the marine seaweed S1 (qmax = 97.37 mg g−1 ). However, the exchanged bentonite B1 presents the highest b constant value (0.0316 L mg−1 ), which represents the adsorbent/pollutant affinity according to Langmuir theory and the highest adsorption capacity (kf = 1.824) and adsorption intensity (n = 2.358) according to the Freundlich isotherm. Regarding the adsorption of 2-CP, the maximum adsorption capacity is reported by the alga S1 (qmax = 97.37 mg g−1 ) according to Langmuir and the alga S2 has the maximum adsorption capacity (kf = 3.297) and adsorption intensity (n = 3.714) according to Freundlich. According to the observed specific surface areas of the algae, we expected a higher adsorption for S2. Early studies showed that S1 presents a higher fucoidan concentration in its structure, whereas S2 possibly has a higher concentration of alginates and/or higher protein content [29]. We believe that the difference in the composition of both algae alters their adsorption capacities, regardless their specific surface area.
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As for the coefficients of correlation, we can postulate that both models are adequate for modeling the isotherm of the removal of 2-NP and 2-CP by the four studied adsorbents. Therefore, a mixed adsorptive mechanism under these experimental conditions can be suggested. Finally, calculated Langmuir and Freundlich constants confirm the differences in the adsorption mechanisms by cross-linked algae and exchanged bentonites elucidated by the effect of pH. In the adsorption of 2-NP, S1 and S2 showed about 3–4-fold the maximum adsorption capacity (qmax ) of B1 and B2. This difference can be explained by the efficient polar interaction between the neutral 2-NP and polyalginic acids present on the surface of the algae, whereas, the adsorption by exchanged bentonites at high pH value can only be explained through the interaction of the 2-nitrophenolate anion with the quaternary ammonium cation. In order to approach the ammonium cation, however the 2-nitrophenolate ion has to go through a hostile 16 carbonaliphatic chain for B1 and through an aromatic ring for B2. Such obstacles are the reason for the inferior adsorption capacity of 2-NP by exchanged bentonites compared to cross-linked algae. In the case of 2-CP, S1 and S2 displayed 2-fold the maximum adsorption capacity of exchanged bentonites at low pH. Assuming a completely neutral 2-CP in both cases, the explanation would be the formation of dipole–dipole interactions or hydrogen bonding between 2-CP and the polysaccharides of the algae and weak dipole-induced dipole interactions between 2-CP and the aliphatic/aromatic motifs of B1 and B2. This second case is not as dramatic as that of 2-NP which is reflected with the smaller differences in qmax between algae and bentonites. Future work will involve the effect of temperature on the adsorption of 2-NP and 2-CP, thermodynamic analysis of the process determining the strength of interaction between 2-CP and 2-NP with the adsorbents and more important, the existence of hydrogen bonding formation of both phenolic compounds with algae. 4. Conclusions Based on their adsorption capacity of 2-NP and 2-CP, biological and non biological adsorbents have been compared. Marine seaweeds Macrocystis integrifolia Bory (S1) and Lessonia nigrescens Bory (S2), cross-linked with CaCl2 , displayed a higher affinity towards 2-NP and 2-CP compared to the bentonite exchanged with quaternary ammonium salts: hexadecyltrimethylammonium bromide (B1) and with bencyltriethylammonium chloride (B2). Batch experiments at room temperature report a strong solution pH effect and a maximum adsorption capacity of 97.37 and 24.18 mg g−1 of 2-NP and 2-CP, respectively by the alga S1. According to the equilibrium studies, the adsorption of 2-NP is ranked: S1 > S2 > B2 > B1 whereas for 2-CP is: S1 > S2 > B2 ≈ B1, following in all cases a mixed mechanism according to Langmuir and Freundlich models. The calculated constants from the isotherms show promising adsorbent/pollutant affinity for their use on the removal of phenolic compounds in real solutions. Finally, octanol–water partition coefficient plays an important role on the adsorption of organic compounds by altering the binding interactions between two different adsorbates with the same adsorbent and between the same adsorbate with different adsorbents. Acknowledgements
Fig. 12. Linearized Freundlich isotherms obtained from 2-CP adsorption by S1 (), S2 ((), B1 (() and B2 (). Initial conditions: 500 mg of adsorbent in 100 mL of solution at the optimum pH.
The authors thank the National Council of Science and Technology of Peru for the grant 152-2006-CONCYTEC-OAJ used in carrying out this work. We also wish to extend our grateful appreciation to
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Rosario Portales, Todd Kelly and Richard Buran for their contributions and critical comments regarding this research. References [1] Z. Aksur, D. Akpinar, Competitive biosorption of phenol and chromium (VI) from binary mixtures onto dried aerobic activated sludge, Biochem. Eng. 7 (2001) 87–99. [2] N. Cuizano, A. Navarro, Biosorción de metales pesados por algas marinas: Posible solución a la contaminación a bajas concentraciones, An. Quím. 104 (2008) 120–125. [3] L. Keith, W. Telliard, Priority pollutants I – a perspective view, Environ. Sci. Technol. 13 (1979) 416–423. [4] D. Blanco, B. Llanos, N. Cuizano, H. Maldonado, A. Navarro, Optimización de la adsorción de cadmio divalente en Lessonia trabeculata mediante reticulación con CaCl2 , Rev. Soc. Quím. Perú 70 (2004) 147–157. [5] T. Davis, B. Volesky, A. Mucci, A review of the biochemistry of heavy metal biosorption by brown algae, Water Res. 37 (2003) 4311–4330. [6] A. Kapoor, T. Virarghavan, Fungal biosorption – an alternative treatment option for heavy metal bearing wastewaters: a review, Bioresour. Technol. 53 (1995) 195–203. [7] K. Ramos, B. Llanos, H. Maldonado, A. Navarro, Evidencias del mecanismo de adsorción de cadmio divalente en Lentinus edodes, An. Quím. 103 (2007) 36–40. [8] A. Ofomaja, Y.-S. Ho, Effect of pH on cadmium biosorption by coconut copra meal, J. Hazard. Mater. 139 (2007) 356–362. [9] C. Lacher, R. Smith, Sorption of Hg (II) by Potamogeton natans dead biomass, Miner. Eng. 15 (2002) 187–191. [10] A. Osorio, D. Bazán, M. Carhuancho, N. Salas, G. Zárate, R. Lengua, R. Aguirre, E. Becerra, Aplicación de la montmorillonita en la descontaminación de efluentes mineros, Rev. Soc. Quím. Perú 70 (2004) 18–26. [11] O. Abollino, M. Aceto, M. Malandrino, C. Sarzanini, E. Mentasti, Adsorption of heavy metals on Na-montmorillonite. Effect of pH and organic substances, Water Res. 37 (2003) 1619–1627. [12] P. Stathi, K. Litina, D. Gournis, T. Giannopoulos, Y. Deligiannakis, Physicochemical study of novel organoclays as heavy metal ion adsorbents for environmental remediation, J. Colloid Interface Sci. 316 (2007) 298–309. [13] EPA Method 604, Phenols in Federal Register, Environmental Protection Agency, Part VIII, 40 CFR Part 136, 58, Revi, October 26, 1984. [14] J. Fawell, S. Hunt, Environmental Toxicology: Organic Pollutants, 1st ed., Ellis Horwood, England, 1998. [15] S. Allen, B. Koumonova, Z. Kircheva, S. Nenkova, Adsorption of 2-nitrophenol by technical hydrolysis of lignin: kinetics, mass transfer and equilibrium studies, Ind. Eng. Chem. Res. 44 (2005) 2281–2287. [16] J. Kilduff, C. King, Effect of carbon adsorbent surface properties on the uptake and solvent regeneration of phenol, Ind. Eng. Chem. Res. 36 (1997) 1603–1613. [17] F. Banat, B. Al-Bashir, S. Al-Asheh, O. Hayajneh, Adsorption of phenol by bentonite, Environ. Pollut. 107 (2000) 391–398. [18] E. Tuesta, M. Vivas, M. Sun-Kou, A. Gutarra, Modificación Química de arcillas y su aplicación en la retención de colorantes, Rev. Soc. Quím. Perú 71 (2005) 26–36. [19] K. Vasanth, V. Ramamurthi, S. Sivanesan, Biosorption of malachite green, a cationic dye onto Pithophora sp., a fresh water algae, Dyes Pigments 69 (2006) 102–107.
[20] S. Yapar, V. Özbudak, A. Dias, A. Lopes, Effect of adsorbent concentration to the adsorption of phenol on hexadecyl trimethyl ammonium-bentonite, J. Hazard. Mater. B121 (2005) 135–139. [21] V. Vilar, C. Botelho, R. Boaventura, Methylene blue adsorption by algal biomass based materials: biosorbents characterization and process behaviour, J. Hazard. Mater. 147 (2007) 120–132. [22] J. Lazo, A. Navarro, M. Sun-Kou, Bertha Llanos, Empleo de Arcillas Modificadas para la adsorción de fenol presente en soluciones acuosas, Rev. Soc. Quím. Perú 73 (2007) 166–170. [23] M. Radica, K. Palanivelu, Adsorptive removal of chlorophenols from aqueous solutions by low cost adsorbent – Kinetics and isotherm analysis, J. Hazard. Mater. B138 (2006) 116–124. [24] N. Calace, E. Nardi, B. Petronio, M. Pietroletti, Adsorption pf phenols by papermill sludges, Environ. Pollut. 118 (2002) 315–319. [25] E. Rubin, P. Rodriguez, R. Herrero, M. Sastre de Vicente, Biosorption of phenolic compounds by the brown alga Sargassum muticum, J. Chem. Technol. Biotechnol. 81 (2006) 1093–1099. [26] E. Rubin, P. Rodriguez, R. Herrero, J. Cremades, J. Barbara, M. Sastre de Vicente, Removal of methylene blue from aqueous solutions using as biosorbent Sargassum muticum: an invasive microalga in Europe, J. Chem. Technol. Biotechnol. 80 (2005) 291–298. [27] U. Thawornchaisit, K. Pakulanon, Application of dried sewage sludge as phenol biosorbent, Bioresour. Technol. 98 (2007) 140–144. [28] S. Arellano-Cardenas, T. Gallardo-Velazquez, G. Osorio-Revilla, M. Lopez-Cortez, B. Gomez-Perea, Adsorption of phenol and dichlorophenols from aqueous solutions by porous clay hetero-structure (PCH), J. Mex. Chem. Soc. 49 (2005) 287–291. [29] A. Navarro, R. Portales, M. Sun-Kou, B. Llanos, Effect of pH on phenol biosorption by marine seaweed, J. Hazard. Mater. 156 (2008) 405–411. [30] M. Boufatit, H. Ait-Amar, W. McWhinnie, Development of an Algerian material montmorillonite clay. Adsorption of phenol, 2-chlorophenol and 2,4,6trichlorophenol from aqueous solutions onto montmorillonite exchanged with transition metal complexes, Desalination 206 (2007) 394–406. [31] P. Sheng, Y. Ting, J. Chen, L. Hong, Sorption of lead, copper, cadmium, zinc and nickel by marine algal biomass: characterization of biosorptive capacity and investigation of mechanism, J. Colloid Interface Sci. 275 (2004) 131–141. [32] C. Pouchert, The Aldrich Library of Infrared Spectra, Aldrich Chemical Company, 1975. [33] V. Padmavathy, P. Vasudevan, S. Dhingra, Thermal and spectroscopic studies on sorption of nickel (II) ion on protonated baker’s yeast, Chemosphere 52 (2003) 1807–1817. [34] N. Farinella, G. Matos, M. Arruda, Grape bagasse as a potential biosorbent of metals in effluent treatments, Bioresour. Technol. 98 (2007) 1940–1946. [35] H. He, J. Duchet, J. Galy, J. Gerard, Influence of cationic surfactant removal on the thermal stability of organoclays, J. Colloid Interface Sci. 295 (2006) 202–208. [36] C. Hedley, G. Yuan, B. Theng, Thermal analysis of montmorillonites modified with quaternary phosphonium and ammonium surfactants, Appl. Clays Sci. 35 (2007) 180–188. [37] Y. Xi, Z. Ding, H. He, R. Frost, Structure of organoclays an X-ray diffraction and thermogravimetric analysis study, J. Colloid Interface Sci. 277 (2004) 116–120. [38] V. Makovskaya, J. Dean, W. Tomlinson, M. Comber, Octanol-water partition coefficients of substituted phenols and their correlation with molecular descriptors, Anal. Chim. Acta 315 (1995) 193–200. [39] C. Hansch, A. Leo, Substituent Constants for Correlation Analysis in Chemistry and Biology, Wiley, New York, 1979, p. 339.
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Comparative study of the removal of phenolic compounds by biological and non-biological adsorbents Abel E. Navarro a,∗ , Norma A. Cuizano b , Jose C. Lazo c , María R. Sun-Kou c , Bertha P. Llanos b a
Department of Chemistry, Graduate School of Arts and Sciences, New York University, New York, NY 10003, United States Departamento Académico de Química, Facultad de Ciencias y Filosofía, Universidad Peruana Cayetano Heredia, Lima L31, Peru c Sección Química, Departamento de Ciencias, Pontificia Universidad Católica del Perú, Lima L32, Peru b
a r t i c l e
i n f o
Article history: Received 10 February 2008 Received in revised form 10 September 2008 Accepted 17 September 2008 Available online 27 September 2008 Keywords: Exchanged bentonite Cross-linked algae Partition coefficient pH Adsorption
a b s t r a c t The ability of biological and non-biological adsorbents to remove 2-nitrophenol (2-NP) and 2chlorophenol (2-CP) from aqueous solutions in batch experiments at room temperature was compared. The marine seaweeds Macrocystis integrifolia Bory (S1) and Lessonia nigrescens Bory (S2) were crosslinked with CaCl2 to enhance their mechanical properties. Natural bentonite was chemically exchanged with hexadecyltrimethylammonium bromide (B1) and bencyltriethylammonium chloride (B2) to increase their affinity towards organic compounds as well. The adsorption capacity of all of the adsorbents strongly depends on solution pH, whereas equilibrium assays showed a mixed mechanism according to the Langmuir and Freundlich isotherms. The maximum adsorption capacity of 2-NP follows the trend: S1 > S2 > B2 > B1 within the range of 97.37 and 18.64 mg g−1 whereas for 2-CP, it ranged between 24.18 and 9.95 mg g−1 with the trend: S1 > S2 > B2 ≈ B1. The importance of the octanol–water partition coefficient as the main factor on the adsorption of these compounds on two different kinds of adsorbents is discussed. © 2008 Elsevier B.V. All rights reserved.
1. Introduction At present, one of the most serious concerns faced in the natural environment is chemical contamination with organic and inorganic substances catalyzed by the presence of heavy metals and phenolic compounds [1,2]. The presence of such compounds in an aquatic medium originates changes in chemical (pH, chemical oxygen demand, alkalinity, acidity, dissolved oxygen, etc.), physical (color, temperature, odor, viscosity, turbidity, etc.) and biological properties, damaging the water quality for human usage and upsetting the environmental equilibrium [2,3]. The uptake of toxic and heavy metals ions from waste waters and industrial wastes by certain types of microbial biomass especially algae [4,5], fungi [6,7] and other biological adsorbents [8,9] has been the subject of many recent studies. Natural clays have also shown an increasing number of applications on the removal of heavy metals [10–12]. Among the different pollutants of aquatic systems, phenols are considered as priority contaminants since they are harmful to plants, animals and humans, even at low concentrations [3,13]. Due to their extensive use and slow degradation, nitrated and chlorinated phenols readily enter the environment by agricultural runoff
∗ Corresponding author. Tel.: +1 646 2862082; fax: +1 212 9954475. E-mail address: [email protected] (A.E. Navarro). 0304-3894/$ – see front matter © 2008 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2008.09.077
of pesticides, effluents from oil refineries and plastic industries and by leachates emerging from waste deposits produced by microbial hydrolysis and photodegradation of several organophosphorous pesticides, such as parathion [14,15]. Activated carbon is one of the most efficient adsorbents for these organic molecules, as it possesses a high surface area per unit mass and exhibits a high affinity for phenolic compounds [16]. Unfortunately, due to the high cost of activated carbon, it is not an economical adsorbent compared with low cost, naturally occurring alternatives. These non-expensive adsorbents remove trace amounts of organic contaminants from wastewaters where other sophisticated techniques such as ionic exchange, osmosis, and solvent extraction are not applicable. Consequently, the environmental biotechnology is in a constant search of alternative, less expensive and viable techniques for the removal of these phenolic compounds at very low concentrations. Adsorption using naturally occurring adsorbents has become a cheap and efficient tool, easily applicable in the detoxification of residual waters from industries and mines showing positive results. High efficiency on the elimination of artificial colorants and phenols from aqueous solutions by marine seaweeds and natural clays has been observed in previous studies [17–22]. Peru is widely known for its diversity of natural resources. Marine seaweeds are in such abundance on the shores of beaches as to cause unsightly accumulation. Also, natural clays of the type
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montmorillonite, saponite and kaolinite are extracted from natural mines in the northern part of the country at very low price. For reasons of accessibility and abundance, Peruvian algae and natural clay of the type bentonite were chosen and compared for the removal of 2-nitrophenol and 2-chlorophenol from aqueous solutions. Several research groups have investigated the adsorptive capacity of other adsorbents such as lignins, coconut shells and papermill sludges [15,23,24] in the removal of nitrated and chlorinated phenols. Previous work has already been carried out utilizing marine algae and clays for the elimination of phenolic compounds [17,19] but none have compared the adsorption capacity of these two adsorbents or explained a relationship between their chemical and mechanical properties and adsorption. 2. Materials and methods 2.1. Biological adsorbent – algae The marine seaweeds were obtained from the beaches of Marcona and Tacna in Peru and identified as Macrocystis integrifolia Bory (S1) and Lessonia nigrescens Bory (S2), respectively. After collection, both brown marine algae were washed twice with tap water in order to remove adhering soil, other microscopic algae, insect larvae, etc. In the laboratory, they were washed with double-distilled water and dried in sunlight and then in an oven at 60 ◦ C for 24 h until all the moisture was evaporated. The dried seaweeds were ground and separated by size by means of sieves reaching a particle size between 150 and 212 m. Rubin et al. [25] indicated that untreated Sargassum muticum was easily decomposed in presence of phenolic compounds. Moreover, the decomposition products interfere in the absorbance measurements and consequently, in the determination of phenolic compounds in aqueous solution. Therefore, to enhance their mechanical properties, the algae were cross-linked with calcium chloride following the procedure cited by Blanco et al. [4], vacuum filtered and washed with double-distilled water to remove the excess of CaCl2 . Finally, the seaweeds were dried at 60 ◦ C and stored under refrigeration until their use. 2.2. Non-biological adsorbent – clays The natural clays used in this study were obtained from Piura, Peru and classified as bentonite. In order to separate the pure montmorillonite fraction from any adhering materials the clay samples were purified by controlled sedimentation. The samples were left to settle, decanted and dried at 60 ◦ C in an oven for 24 h. Finally, the bentonite was passed through a 2 m sieve and taken for further treatment. The bentonite was chemically modified with two different quaternary ammonium salts: bencyltriethylammonium bromide (BTEA) and hexadecyltrimethylammonium chloride (HDMTA), which have shown high affinity towards organic compounds [20,22], to enhance their adsorption capacity. 20 g of bentonite suspended in 2 L of distilled water were combined with two equivalents of its cation exchange capacity (CEC) of the ammonium salts (99% Sigma–Aldrich). The final suspension was stirred at 250 rpm for 12 h at room temperature. The modified bentonites were vacuum filtered and vigorously rinsed with double-distilled water to eliminate the excess of ammonium salts, dried at 60 ◦ C in an oven for 48 h and stored until their use. The bentonites were identified according to the ammonium salt used in their modification as follows: bentonite treated with HDMTA (B1) and bentonite treated with BTEA (B2).
The CEC of the bentonite (0.06 mmol g−1 ) was determined using the methylene blue standard method (ASTM C-837-81). 2.3. Characterization of adsorbents The specific surface area of both types of adsorbents was measured. For the algae, the methylene blue adsorption method was used. The calculation was made assuming a value of 108 (Å)2 for the ionic cross-sectional area of methylene blue [26]. The surface area of the modified bentonites was determined on the basis of the Brunauer, Emmet and Teller (BET) method using an ASAP 2000 (Micromeritics). Infrared spectra of all the adsorbents were recorded on a PerkinElmer 1600 infrared spectrometer in order to identify the binding groups present in their surface. FT-IR samples were prepared by combining 0.5 g of powered potassium bromide with 1 mg of dried adsorbent 24 h prior to recording. Thermal stability and presence of volatile impurities in the adsorbents were analyzed by thermogravimetric studies using a Linseis STA 1600 analyzer at a heating rate of 5 ◦ C/min up to 1000 ◦ C. In order to compare the functional groups observed by FT-IR, elemental composition of the adsorbents was evaluated by standard methods and ICP multi-elemental analysis. An Environmental Scanning Electron Microscope (PHILIPS ESEM XL-30 TMP) was selected to elucidate the surface texture and morphology of the adsorbents. Prior to the observation, the surface of the samples was coated with a thin and electrically conductive gold film. 2.4. Batch adsorption test Duplicated batch adsorption experiments were carried out at room temperature by allowing an accurately weighed amount of cross-linked algae and exchanged bentonites to reach equilibrium with 2-nitrophenol (2-NP) and 2-chlorophenol (2-CP) solutions of known concentrations. Initial concentrations of 2-NP and 2-CP were held between 50 and 1000 mg/L. The pH was adjusted by adding concentrated solutions of HCl and NaOH. The bentonite samples were shaken at 500 rpm for 6 h, as established in preliminary results [22] and the algal samples were shaken at 200 rpm during 24 h to assure the completeness of the adsorption. At the end of the equilibrium the samples were filtered and subsequently analyzed for residual concentration of 2-NP and 2-CP by UV–vis spectrometry (SHIMADZU UV-MINI 1240) at wavelengths of 274 and 276 nm for 2-NP and 2-CP, respectively. 2.5. Data analysis The amount of adsorbed phenolic compound is expressed as adsorption capacity (q, mg/g) calculated as follows: q=
(Ci − Cf ) × V m
(1)
where m is the mass of adsorbent expressed in g, V is the volume of the solution in L and Ci and Cf are the initial and at the equilibrium concentrations, respectively expressed in mg/L. 3. Results and discussion 3.1. Adsorbent characterization 3.1.1. Specific surface area Specific surface areas of 820 and 1512 m2 g−1 for the brown algae S1 and S2, respectively were observed by means of the methylene blue method in the water-wet state of the algae. On other
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Fig. 1. FTIR spectra of cross-linked algae: Macrocystis integrifolia Bory (S1) and Lessonia nigrescens Bory (S2).
hand, bentonite samples B1 and B2 showed a surface area of 24 and 34 m2 g−1 with the BET method (adsorption and desorption of nitrogen). These algae report higher surface areas and demonstrate higher porosity and a better distribution of active sites on their surface when compared to the algae Gelidium (375 m2 g−1 ) [21] and S. muticum (485 m2 g−1 ) [26] as well as other naturally occurring adsorbents such as lignin (19.72 m2 g−1 ) [15] and dried sewage sludge [27]. The bentonites, upon treatment with the ammonium salts underwent a decrease in the surface area compared to other natural (66 m2 g−1 ) [18] and modified clays such as porous clay heterostructure (305.5 m2 g−1 ) [28]. This loss of surface area is attributed to the full occupation of the interlamellar spaces of the bentonite by organic cations containing a large hydrophobic alkyl chain (HDMTA) and aromatic ring (BTEA), slightly reducing their porosity. The substantial difference in the specific surface areas between these two types of adsorbents is explained by the use of different methods in their determination. It is known that in aqueous solution, the algae swell and increase their dimensions because of the
water filling their pores [25], whereas these changes in the volume of organophilic clays are rarely seen. It has been proposed that the surface area of algae is more accurately determined by the methylene blue method, where the adsorbent is better resembled at the moment of the adsorption. On the other hand, organophilic clays are almost unaffected by the presence of water so the BET method is considered more accurate than other techniques. 3.1.2. Fourier transform infrared spectroscopy and composition The FT-IR revealed the presence of the typical functional groups in algae (Fig. 1) as well as the interlamellar insertion of organic cations in the clay’s structure (Fig. 2). As previously shown [29], the algae S1 and S2 display peaks at 3410, 1634 and 1425 and at 1160 and 817 cm−1 attributed to the presence of polyalginate and fucoidan, respectively. As for the modified bentonites (Fig. 2), common peaks can be found at 1636 and 3445 cm−1 related to the vibration of O–H bonds [30,31] that belongs to remaining water molecules physically adsorbed on the clays. Stretching vibrations Si–O (1040 cm−1 ) and Si–O–Si (570 cm−1 ) are also observed [30].
Fig. 2. FTIR spectra of bentonites exchanged with hexadecyltrimethylammonium chloride (B1) and with bencyltriethylammonium bromide (B2).
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Table 1 Elemental composition of the marine seaweeds. Composition (%)
S1
S2
N Ca P Fe Fiber
1.81 5.51 1.89 1.11 11.1
4.12 5.46 1.38 0.15 10.7
Table 2 Elemental composition of the modified bentonites. Composition (%)
B1
B2
Na Mg Al2 O3 SiO2 K Ca Ti Mn Fe
0.04 0.38 11.6 62.3 0.05 0.48 0.01 0.01 0.56
0.05 0.42 11.9 64.5 0.05 0.38 0.01 0.01 0.69
The only peculiar difference between B1 and B2 FT-IR spectra (Fig. 2) is two sharp peaks in B1 around 2850–2920 cm−1 associated with asymmetric and symmetric stretching of C–C bonds due to the length of the aliphatic chain of HDMTA [32]. To corroborate these results Tables 1 and 2 show the elemental composition of the adsorbents, exhibiting in the algae a high content of calcium due to the cross-linkage and fibers assumed to correspond to the polyalginates and fucoidans [5,29,31]. According to the ICP multi-element analysis of the exchanged bentonites, silica and alumina are the main constituents as reflected in the FT-IR analysis. 3.1.3. Thermogravimetric analysis Figs. 3 and 4 show the thermogravimetric analysis of crosslinked algae and modified bentonite, respectively. The thermal studies of algae showed that both algae underwent three steps decomposition process when heated from 20 to 550 ◦ C. Initial step (∼4% weight loss), which was small in the range in 25–110 ◦ C, could be attributed to the loss of adsorbed water [33]. The second step from 240 to 320 ◦ C (∼6% weight loss) corresponds to volatile substances releases proceeding from inorganic compound decomposition as well as organic matter degradation. At this range of
Fig. 3. Thermogravimetric curves of cross-linked S1 and S2 seaweeds.
Fig. 4. Thermogravimetric curves of exchanged bentonites B1 and B2.
temperature some carbohydrates degrade (around 300 ◦ C). Finally, the weight loss at temperatures higher than 480 ◦ C is caused by fiber and lignin pyrolysis [34]. As for the modified bentonites, B1 shows a two-steps decomposition process when heated from 20 to 500 ◦ C. The first weight loss of 1.5% occurs between 25 and 100 ◦ C, also associated with the loss of adsorbed water. The second step (∼5% weight loss) occurs in the range between 150 and 300 ◦ C, corresponding to two events: (i) the water loss associated with the residual exchange cations that remain in the bentonite upon the treatment with the ammonium salt or (ii) the early decomposition of the ammonium salt that has been inserted in the bentonite [35,36]. Finally, a third step between 300 and 550 ◦ C, corresponds to structural dehydroxylation of the bentonite (∼1% weight loss). This dehydroxylation was also observed by Xi et al. in organophilic bentonites, and continues until 636 ◦ C [37]. On the other hand, the modified bentonite B2, shows a thermal curve similar to B1, but the weight loss percentages are lower for B2. Since, the water loss in the first step is the same for B1 and B2; we believe that a lower exchange with the ammonium salt BTEA has happened. The difference in this initial loss of water between both types of adsorbents is attributed to the hygroscopicity of the polysaccharic composition of the algae compared to the hydrophobicity of modified bentonites. These results show that the cross-linkage of algae with calcium and the incorporation of quaternary ammonium ions in bentonites show mechanical stability and heat resistance up to 300 ◦ C compared to other adsorbents where oxidative decomposition has been observed at lower temperatures [33,34]. 3.1.4. Scanning electron microscopy The electron micrograph scans clearly reveal the surface texture and morphologies of the modified adsorbents (Figs. 5 and 6). It is evident from the micrographs that both algae show a very well-defined organization of the polysaccharic structures in their surfaces where there is an even surface texture. Contradictorily, both exchanged bentonites showed a repetitive formation of “layer over layer” structure. Moreover, the micrograph of B2 displays layers on the clays that have become more planar and irregular than the ones observed in B1. This can be attributed to a less efficient packing of aromatic organic ions BTEA compared to the aliphatic organic ions HDMTA between the interlamellar spaces of the clay.
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Fig. 5. SEM images showing the surface morphologies of S1 (left) and S2 (right) at ×1280 magnification.
Fig. 6. SEM images showing the surface morphologies of B1 (left) and B2 (right) at ×6925 magnification.
3.2. Sorption studies 3.2.1. Effect of initial pH value pH is one of the most important factors in the adsorption of environmental contaminants in aqueous solutions since it influences the solution chemistry of phenols (i.e. hydrolysis, redox reactions, polymerization and complexation). pH also has a strong influence on the speciation and the sorption availability of phenolic compounds as well as the ionic state of functional groups on the surface of adsorbents [7,8,29]. Fig. 7 shows the effect of solution pH on the adsorption of 2NP at room temperature. For algae, the highest adsorption capacity
Fig. 7. Effect of pH on 2-NP adsorption by S1 (), S2 ((), B1 (() and B2 (). Initial conditions: 100 mg of adsorbent at initial concentration of 2-NP of 100 mg L−1 .
is reported at low pH (q = 6.2 and 7.1 mg g−1 for S1 and S2 at pH 4 and 3, respectively). Contradictorily for exchanged bentonites show an adsorption capacity that gradually increases with the pH (q = 8.8 mg g−1 for B1 at pH 10), whereas for B2, there is not observable pH effect within the range values of 2–10, reaching a steady adsorption capacity around 2.0 mg g−1 . These phenomena can be explained as the product of electrostatic repulsion/attraction. Recent studies [29] demonstrated that the algae S1 and S2 have apparent ionization constants around 3.0 where at higher pH values, the algae’s surfaces are negatively charged. At the same time, 2-NP (pKa = 7.23) becomes negatively charged with increasing pH. Consequently, the neutral state of 2-NP promotes its adsorption on cross-linked seaweeds. On the other hand, the exchange of ions with HDMTA cations in bentonites creates a permanent and non-pH-dependent positive charge evenly distributed on the surface, shown by SEM analysis, that preferentially binds to negatively charged 2-nitrophenolate which exists in higher concentrations at high pH values. Likewise, as stated in Section 3.1.4, the micrographs revealed a scattered packing of BTEA in the clay B2, which added to the high hydrophobicity of the aromatic ring, reduces the pH effect in the process. The relationship between 2-CP uptake and solution pH for crosslinked algae and exchanged bentonite is illustrated in Fig. 8. The graph clearly exhibits that lower pH values led to higher 2-CP uptake for both kind of adsorbents (q = 10.5, 5.0, 18 and 12 mg g−1 at pH values 2, 3, 4 and 4 for S1, S2, B1 and B2, respectively). According to their octanol–water partition coefficients, 2-NP is more polar than 2-CP (log Kow = 1.77 and 2.15, respectively) [38,39], therefore 2-CP (pKa = 8.48) does not have a distribution of electron density as expanded as 2-NP and will display a better adsorption on neutral surfaces. Algae corroborate this behavior by displaying their higher adsorption capacity at pH lower than their respective apparent ionization constants [29]. In the case of exchanged
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Fig. 8. Effect of pH on 2-CP adsorption by S1 (), S2 ((), B1 (() and B2 (). Initial conditions: 100 mg of adsorbent in 100 mL of solution at initial concentration of 2-CP of 100 mg L−1 .
Fig. 9. Linearized Langmuir isotherms obtained from 2-NP adsorption by S1 (), S2 ((), B1 (() and B2 (). Initial conditions: 500 mg of adsorbent in 100 mL of solution at the optimum pH.
bentonites, the decrease on pH makes 2-CP more neutral and friendlier to the aliphatic (HDMTA) and aromatic (BTEA) quaternary ammonium ions. A slight decrease on the adsorption is observed at pH values lower than 4 which can be attributed to the polarization of 2-CP by the formation of complexes or protonation of the hydroxyl group in excess of hydronium ions. Exchanged bentonites also show an increasing adsorption at high pH values (pH ≈ 10) most likely due to the electrostatic affinity between the deprotonated 2-CP and the positive charge of the ammonium ions. Based on these results, we propose a completely electrostatic adsorption of 2-NP and 2-CP by cross-linked algae and a polarity-dependent adsorption for bentonites (apolar adsorbates are adsorbed by apolar mechanisms and polar adsorbates by polar mechanisms) obeying the Traube rule and Rebinder effect. 3.2.2. Equilibrium modeling The adsorption of a substance from one phase to the surface of another in a specific system leads to a thermodynamically defined distribution of that substance between the phases as the system reaches the equilibrium. This distribution can be expressed in terms of adsorption isotherms. The Langmuir equation assumes that (i) the solid surface presents a finite number of identical sites which are energetically uniform; (ii) there is no interaction between adsorbed species, meaning that the amount adsorbed has no influence on the rate of adsorption and (iii) a monolayer is formed when the solid surface reaches the saturation. While the Langmuir isotherm assumes that enthalpy of adsorption does not depend on the amount adsorbed, the Freundlich isotherm is derived assuming a logarithmic decrease in the enthalpy of adsorption with the increase in the fraction of occupied sites and based on sorption on heterogeneous surfaces. Both isotherm equations can be described by their linearized forms and the Langmuir and Freundlich constants can be determined, according to Eqs. (2) and (3), respectively. 1 1 1 = + × qmax × Cf q qmax b
(2)
ln q = ln kf +
(3)
1 n
× ln Cf
where qmax (mg g−1 ) and b (L mg−1 ) are the Langmuir constants related to the maximum adsorption capacity and to the adsorption
Fig. 10. Linearized Freundlich isotherms obtained from 2-NP adsorption by S1 (), S2 ((), B1 (() and B2 (). Initial conditions: 500 mg of adsorbent in 100 mL of solution at the optimum pH.
energy, respectively and kf and 1/n are the Freundlich constants related to the adsorption capacity and the adsorption intensity, respectively. The linearized forms of Langmuir and Freundlich isotherms obtained at room temperature for the adsorption of 2-NP and 2-CP are given in Figs. 9–12, whereas Table 3 presents the corresponTable 3 Isotherm constants for adsorption of phenolic compounds (2-NP and 2-CP) by crosslinked algae and modified bentonites. Adsorbent
Langmuir
Freundlich
qmax (mg g−1 )
b (L mg−1 )
R2
kf
n
R2
2-NP S1 S2 B1 B2
97.37 71.28 18.64 23.02
0.0005 0.0005 0.0316 0.001
0.937 0.965 0.988 0.99
0.035 0.02 1.824 0.068
0.978 0.973 2.358 1.31
0.976 0.977 0.984 0.987
2-CP S1 S2 B1 B2
24.18 17.33 9.95 10.04
0.0135 0.0363 0.0087 0.0086
0.995 0.986 0.988 0.989
1.391 3.297 0.916 0.566
2.269 3.714 2.841 2.302
0.984 0.946 0.992 0.961
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Fig. 11. Linearized Langmuir isotherms obtained from 2-CP adsorption by S1 (), S2 ((), B1 (() and B2 (). Initial conditions: 500 mg of adsorbent in 100 mL of solution at the optimum pH.
dent constants along with the coefficients of linear correlation (R2 ) associated with each linearized model. As can be seen in Table 3, the maximum value for the adsorption capacity for 2-NP was obtained by the marine seaweed S1 (qmax = 97.37 mg g−1 ). However, the exchanged bentonite B1 presents the highest b constant value (0.0316 L mg−1 ), which represents the adsorbent/pollutant affinity according to Langmuir theory and the highest adsorption capacity (kf = 1.824) and adsorption intensity (n = 2.358) according to the Freundlich isotherm. Regarding the adsorption of 2-CP, the maximum adsorption capacity is reported by the alga S1 (qmax = 97.37 mg g−1 ) according to Langmuir and the alga S2 has the maximum adsorption capacity (kf = 3.297) and adsorption intensity (n = 3.714) according to Freundlich. According to the observed specific surface areas of the algae, we expected a higher adsorption for S2. Early studies showed that S1 presents a higher fucoidan concentration in its structure, whereas S2 possibly has a higher concentration of alginates and/or higher protein content [29]. We believe that the difference in the composition of both algae alters their adsorption capacities, regardless their specific surface area.
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As for the coefficients of correlation, we can postulate that both models are adequate for modeling the isotherm of the removal of 2-NP and 2-CP by the four studied adsorbents. Therefore, a mixed adsorptive mechanism under these experimental conditions can be suggested. Finally, calculated Langmuir and Freundlich constants confirm the differences in the adsorption mechanisms by cross-linked algae and exchanged bentonites elucidated by the effect of pH. In the adsorption of 2-NP, S1 and S2 showed about 3–4-fold the maximum adsorption capacity (qmax ) of B1 and B2. This difference can be explained by the efficient polar interaction between the neutral 2-NP and polyalginic acids present on the surface of the algae, whereas, the adsorption by exchanged bentonites at high pH value can only be explained through the interaction of the 2-nitrophenolate anion with the quaternary ammonium cation. In order to approach the ammonium cation, however the 2-nitrophenolate ion has to go through a hostile 16 carbonaliphatic chain for B1 and through an aromatic ring for B2. Such obstacles are the reason for the inferior adsorption capacity of 2-NP by exchanged bentonites compared to cross-linked algae. In the case of 2-CP, S1 and S2 displayed 2-fold the maximum adsorption capacity of exchanged bentonites at low pH. Assuming a completely neutral 2-CP in both cases, the explanation would be the formation of dipole–dipole interactions or hydrogen bonding between 2-CP and the polysaccharides of the algae and weak dipole-induced dipole interactions between 2-CP and the aliphatic/aromatic motifs of B1 and B2. This second case is not as dramatic as that of 2-NP which is reflected with the smaller differences in qmax between algae and bentonites. Future work will involve the effect of temperature on the adsorption of 2-NP and 2-CP, thermodynamic analysis of the process determining the strength of interaction between 2-CP and 2-NP with the adsorbents and more important, the existence of hydrogen bonding formation of both phenolic compounds with algae. 4. Conclusions Based on their adsorption capacity of 2-NP and 2-CP, biological and non biological adsorbents have been compared. Marine seaweeds Macrocystis integrifolia Bory (S1) and Lessonia nigrescens Bory (S2), cross-linked with CaCl2 , displayed a higher affinity towards 2-NP and 2-CP compared to the bentonite exchanged with quaternary ammonium salts: hexadecyltrimethylammonium bromide (B1) and with bencyltriethylammonium chloride (B2). Batch experiments at room temperature report a strong solution pH effect and a maximum adsorption capacity of 97.37 and 24.18 mg g−1 of 2-NP and 2-CP, respectively by the alga S1. According to the equilibrium studies, the adsorption of 2-NP is ranked: S1 > S2 > B2 > B1 whereas for 2-CP is: S1 > S2 > B2 ≈ B1, following in all cases a mixed mechanism according to Langmuir and Freundlich models. The calculated constants from the isotherms show promising adsorbent/pollutant affinity for their use on the removal of phenolic compounds in real solutions. Finally, octanol–water partition coefficient plays an important role on the adsorption of organic compounds by altering the binding interactions between two different adsorbates with the same adsorbent and between the same adsorbate with different adsorbents. Acknowledgements
Fig. 12. Linearized Freundlich isotherms obtained from 2-CP adsorption by S1 (), S2 ((), B1 (() and B2 (). Initial conditions: 500 mg of adsorbent in 100 mL of solution at the optimum pH.
The authors thank the National Council of Science and Technology of Peru for the grant 152-2006-CONCYTEC-OAJ used in carrying out this work. We also wish to extend our grateful appreciation to
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