Potential use of Caulerpa fastigiata biomass for

Potential use of Caulerpa fastigiata biomass for removal of lead: kinetics, isotherms, thermodynamic, and characterization studies B. Sarada, M. Krishna Prasad, K. Kishore Kumar & Ch. V. R. Murthy

Environmental Science and Pollution Research ISSN 0944-1344 Environ Sci Pollut Res DOI 10.1007/s11356-013-2008-z

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Author's personal copy Environ Sci Pollut Res DOI 10.1007/s11356-013-2008-z

RESEARCH ARTICLE

Potential use of Caulerpa fastigiata biomass for removal of lead: kinetics, isotherms, thermodynamic, and characterization studies B. Sarada & M. Krishna Prasad & K. Kishore Kumar & Ch. V. R. Murthy Received: 6 May 2012 / Accepted: 12 July 2013 # Springer-Verlag Berlin Heidelberg 2013

Abstract The present study attempts to analyze the biosorption trend of biosorbent Caulerpa fastigiata (macroalgae) biomass for removal of toxic heavy metal ion Pb (II) from solution as a function of initial metal ion concentration, pH, temperature, sorbent dosage, and biomass particle size. The sorption data fitted with various isotherm models and Freundlich model was the best one with correlation coefficient of 0.999. Kinetic study results revealed that the sorption data on Pb (II) with correlation coefficient of 0.999 can best be represented by pseudosecond-order. The biosorption capacity (qe) of Pb (II) is 16.11±0.32 mg g−1 on C. fastigiata biomass. Thermodynamic studies showed that the process is exothermic (ΔH° negative). Free energy change (ΔG°) with negative sign reflected the feasibility and spontaneous nature of the process. The SEM studies showed Pb (II) biosorption on selective grains of the biosorbent. The FTIR spectra indicated bands corresponding to –OH, COO−, –CH, C=C, C=S, and –C–C– groups were involved in the biosorption process. The XRD pattern of the C. fastigiata was found to be mostly amorphous in nature.

Responsible editor: Elena Maestri B. Sarada Department of Biotechnology, GITAM University, Visakhapatnam 530 003, India M. K. Prasad Department of Chemical Engineering, GMR Institute of Technology, Rajam, 532127 Srikakulam District, Andhra Pradesh, India K. K. Kumar Department of Plant Sciences, School of Life Sciences, University of Hyderabad, Hyderabad, Andhra Pradesh 500046, India K. K. Kumar (*) : C. V. R. Murthy Department of Chemical Engineering, University College of Engineering, Andhra University, Visakhapatnam 530 003, Andhra Pradesh, India e-mail: [email protected]

Keywords Biosorption . Free energy change . Lead (II) . Caulerpa fastigiata . FTIR . SEM

Introduction Process industries, such as battery manufacturing, metal plating, and finishing, are a prime source of Pb (II) pollution in the environment. The current lead level in the drinking water was fixed as 50 and 10 μg L−1 by the Environmental Protection Agency and World Health Organization, respectively. Lead accumulates mainly in bones, brain, kidney, and muscles and may cause many serious disorders like anemia, kidney diseases, nervous disorders, and sickness even death. It is therefore essential to remove Pb (II) from wastewater before disposal. Several different conventional methods were applied to remove excessive heavy metals from aqueous solutions including chemical precipitation, ion exchange, evaporation, electroplating, and membrane processes. However, these methods are either inefficient or expensive, when heavy metals exist in low concentrations (Wilde and Benermann 1993; Kuyucak and Volesky 1998). Consequently, it is urgent to find new technologies or biomaterials for removing heavy metal ions from wastewater. New technologies involving the removal of metals ions from wastewaters have directed attention to biosorption based on metal binding capacities of various biological materials. So, biosorption can be a promising alternative method to treat industrial effluents, mainly because of its low cost, high metal binding capacity, high efficiency in dilute effluents, and environmental friendly (Volesky and Holan 1995). Application of algae, plant, bacteria, fungi, and yeast biomasses have proved to be potential materials (Luo et al. 2006; Kishore et al. 2006; Ahmet and Mustafa 2008; Gupta and Rastogi 2008; Vítor et al. 2008; Kishore et al. 2009; Lalhruaitluanga et al. 2010) for removal of heavy metal ions. There are many reports and reviews on the biosorption of lead metal ion on marine algae

Author's personal copy Environ Sci Pollut Res

(Ahmet and Mustafa 2008; Gupta and Rastogi 2008; Vítor et al. 2008), green seaweed (Tien 2002; Tuzun et al. 2005), and freshwater green algal species (Sudhir and Tripathi 2008; Sibel et al. 2009) with varying removal efficiencies, maximum adsorption capacities (qmax), and binding constants. Caulerpa fastigiata is abundant and a renewable nontoxic algae which could be obtained very economically on a large scale locally from oceans. In the present study, the biosorption capacity of dried algae has been investigated as the function of environmental parameters, adsorption time, pH values, and solid/liquid ratio. In addition, equilibrium and kinetic studies were carried out. Langmuir and Freundlich isotherm models were applied to fit the experimental data. The biosorbent was characterized by employing instrumental techniques, viz, Fourier transform infrared spectroscopy (FTIR), X-ray diffraction (XRD), and scanning electron microscopy (SEM).

Experimental Preparation of biosorbent C. fastigiata

100 mL solution with various metal concentrations (37.1±1.45, 59.3±0.73, 85.2±0.86, and 109.7±0.84 mg L−1) and required amount of adsorbent (2.5, 5.0, 7.5, and 10.0 g L−1) using orbital shaker. Initially, the effect of contact time (0–120 min) on the sorption capacity of C. fastigiata was evaluated. The equilibrium time obtained was used in all experiments, and the experiments were repeated thrice for all conditions of study. Analytical procedure The concentrations of unadsorbed Pb (II) ions in the sample supernatant liquid were determined using an atomic absorption spectrophotometer (PerkinElmer AA200) with an air acetylene flame. C and CT were determined and tabulated for subsequent analysis of the data. The metal uptake (qe) was calculated using the general definition as follows: qe ¼

V ðC−C e Þ M

ð1Þ

Where qe is the metal uptake in milligrams per gram biomass, V is the volume of metal containing solution in contact with the biosorbent in liters, C and Ce are the initial and equilibrium (residual) concentration of metal in the solution in milligrams per liter, respectively, and M is the amount of added biosorbent in grams. Metal percent of removal by C. fastigiata was determined by Eq. 2 as follows:

C. fastigiata The green macroalgae under study was collected from the east coast of India and was identified with authentication by the faculty of the Dept. of Botany, A.U, Visakhapatnam. The biomass was washed thrice using demineralized water for the removal of surface debris, particulate matter, and salts. After washing, the algae powder was dried in a hot air oven (70 °C for 24 h). A domestic mixture was used to reduce the particle size of the alga. The powdered biomass was sieved and separated into several particle sizes ranging between 0.074 and 0.15 mm. The macroalgal powder was preserved in a humidity control oven to maintain a standard humidity throughout for equilibrium studies during the entire period of study.

Where R is the percentage of Pb (II) adsorbed by biomass, C is the initial concentration of metal ions in milligrams per liter, and Ct is the concentration of metal ions at time t in milligrams per liter.

Preparation of stock solution

Characterization of biomass

All the chemicals used were of analytical reagent grade. In order to avoid interference with other elements in the wastewater, the experiments were conducted with aqueous solution of Pb (II). Stock Pb (II) solution (1,000 mg/L) was prepared by dissolving appropriate amount of Pb(NO)3 in Milli-Q water. All the working solutions were prepared by diluting the stock solution. The pH of the test solutions was adjusted (Systronics 361) using reagent grade hydrochloric acid and sodium hydroxide solutions.

FTIR studies The biomass prior and after adsorption was air dried and demoisturized at 60 °C in humidity control oven. The powder was analyzed by FTIR (PerkinElmer 1600) by potassium bromide pellet method in the wave number range of 400 to 4,000 cm−1. The FTIR study is intended to provide a deeper understanding of the interaction between the biomass cell surface and the metal ions.

Equilibrium studies

X-ray diffraction analysis

The experiments were carried out in 250 mL Erlenmeyer conical flasks at a constant agitation speed (160 rpm) with

The XRD of each of the biomass powder sample was obtained using XRD-6000 Shimadzu, Japan Model. The diffracted X-

Rð%Þ ¼

C−Ct 100 Ct

ð2Þ


ray intensities were recorded as a function of 2 , at a scan speed of 1.2°/min, and pattern was recorded from 10 to 70°. SEM studies The dried biomass powders and the corresponding metal ion loaded powders were first coated with ultra thin film of gold by an ion sputter (JFC-1100) and then were exposed under electron microscope (JEOL, JXA-8100). For this purpose, the working height was kept at 15 mm, with working voltage ranging between 10 to 25 kV.

The effect of contact time was studied on percent adsorption of Pb (II) over a time period of 5–90 min, using l g of C. fastigiata biomass powder (diameter of particle; dp=0.074mm), initial Pb (II) concentration is 59.304 mg L−1 at pH 4.5, and temperature 25 °C. Percentage adsorption increased from 75.05 to 81.94 % during a contact time period from 5 to 60 min. The rapid initial sorption was likely due to extracellular polymeric sites (ionisable) binding, and the slower sorption resulted from intracellular binding (Areco and dos Santos Afonso 2010) on alga. Similar studies were performed with 2.5, 5, and 7.5 g L−1 of biomass concentrations in aqueous solutions, and the results indicated that the 1 h is the optimum time of contact for the range of concentrations used.

Results and discussion Effect of pH Effect of contact time Experiments were conducted to estimate the time required to reach the sorption equilibrium by taking an initial charge of 100 mL of aqueous solution containing Pb (II) ions and required quantities of biomass. The mixture was shaken in an orbital shaker, the samples were drawn at regular time intervals, and the metal concentration was estimated using AAS. The data of concentration of metal ion Ct in solution with time are shown for different quantities of biomass in Fig. 1. Experimental results show a faster uptake at initial stages of contact, and subsequent slowing down as the equilibrium is approached. In the initial stages of contact, large numbers of vacant sites are available, and hence the uptake is faster. The slowing down of metal uptake later is due to difficulty in occupying the remaining vacant sites. Repulsive forces between the adsorbed Pb (II) ions and aqueous Pb (II) may also most probably contribute to the slowing down of uptake of metal at equilibrium. Earlier, similar results have been reported for the Pb (II) ions binding on seed husk of Calophyllum inophyllum (Lawal et al. 2010).

pH of the biosorption medium is one of the essential parameters affecting the uptake of heavy metal ions from aqueous solutions by biosorbents (Areco and dos Santos Afonso 2010). This parameter is directly related to the ability of hydrogen ions and metal ions to bind themselves to active sites on the biosorbent surface (Ahmet et al. 2008). Generally, metal biosorption involves complex mechanisms of ion exchange, chelation, adsorption by physical forces, and ion entrapment in inter and intrafibrillar capillaries and spaces of the cell structural network of a biosorbents (Volesky and Holan 1995). Depending on the pH value of the aqueous solution, the functional groups on the cell wall participate in metal ion binding (Dhananjay and Gaur 2011). The effect of pH can be explained by ion-exchange mechanism of sorption in which a significant role is played by the functional groups of biomass that have cation-exchange properties (Areco and dos Santos Afonso 2010). The effect of pH on the biosorption of Pb (II) ions onto C. fastigiata biomass was studied in the pH range 2–8, and the results are presented in Fig. 2. The uptake of biomass was found to increase with the increase in pH. It was found to have a high content of carboxyl groups that render it susceptible to pH changes. Similar observations were reported earlier by several investigators (Wang and Chen 2009; Yi-Chao and Shui-Ping 2011). Figure 2 shows that an increase in initial pH from 2.0 to 5.0 resulted in biosorption of Pb (II) from 25.8 to 88.4 % at an initial metal concentration of 37.13 mg L−1. Optimum pH for Pb (II) biosorption was found to be 5.0. A drastic fall in percent biosorption as low as 19.68 % was observed at pH value of 8.0. It may be because of the tendency of precipitation of lead as Pb (OH)2 (Yi-Chao and Shui-Ping 2011). Effect of metal ion concentration

Fig. 1 Variation of Pb (II) concentration in solution with time at various quantities of biomass at 25 °C, particle diameter 0.074 mm, C 59.304 mg L−1, and pH 4.5

The effect of initial Pb (II) concentration on the metal uptake was shown in Fig. 3. The adsorption capacity (qe) of the


Fig. 2 Effect of pH on the lead removal percentage at various initial Pb (II) concentrations at temperature of 25 °C, biomass weight 10 g L−1, and particle diameter 0.074 mm

biomass increased from 3.28 to 8.77 mg g−1 with increasing Pb (II) concentration from 37.1 to 109.7 mg L−1 at sorbent dose of 10 g L−1 with the temperature of 25 °C and pH 5. An increase in the initial ion concentration provides a larger driving force to overcome all mass transfer resistances between the solid and the aqueous phase, which results in higher metal ion adsorption. Similar observations also were made by earlier investigators (Mohammad et al. 2011) in their studies on the adsorption of lead.

Fig. 4 Variation of % Pb (II) sorption with biomass weight at temp 25 °C, pH 5, and C 59.304 mg L−1 and at various particle sizes

to 10 g L−1. This decrease could be due to the concentration gradient between the sorbent and the sorbate; an increase in biomass caused a decrease in the amount of metal sorbed onto a unit weight of the algae. Moreover, the increase in percentage biosorption of metals by increasing the biomass dosage is due to an increase in the number of active sites and surface area available for biosorption. Similar trends have been reported in the literature (Taqvi et al. 2006; Gupta and Rastogi 2008; Yi-Chao and Shui-Ping 2011). From Fig. 4, we can conclude that percentage adsorption increased with decrease in biomass dosage.

Effect of adsorbent dosage Biosorbent dosage determines the potential of biosorbent through the number of binding sites available to remove metal ions at a specified initial metal ion concentration. The effect of amount of biomass was studied on the biosorption of Pb (II) using C. fastigiata. The percent removal of Pb (II) on C. fastigiata ranged from 67.91 to 81.56 (Fig. 4) at a pH value of 5, and the uptaking capacity declined from 6.13 to 4.84 mg g−1 when increasing the biomass dosage from 2.5

Fig. 3 Variation of % removal of Pb (II) with various initial concentrations at temperature of 25 °C, pH 5, particle diameter 0.074 mm, and biomass 10 g L−1

Effect of adsorbent size The effect of particle size in the range 0.074 to 0.150 mm on (Fig. 5) adsorption percentage of Pb (II) on C. fastigiata was investigated. A decrease of about 10–12 % in the percent adsorption was found with an increase in the particle size by twofold. The decrease in the average particle size of the adsorbent increased the surface area. The ion exchange capacities much depends on the specific surface area available for solute–surface interaction, which is accessible to solute.

Fig. 5 Variation of % Pb (II) sorption with various adsorbent sizes at temp 25 ° C, pH 5, and C 59.304 mg L−1 using C. fastigiata as biosorbent


adsorption (Dąbrowski 2001). This equation is applicable to the physical or chemical adsorption on solid surface with one type of adsorption active center. As long as the limitations are clearly recognized, the Langmuir equation can be used for describing equilibrium conditions for adsorption behavior in different adsorbate–adsorbent systems or for varied conditions within any given system. Linear form of the Langmuir equation is given as follows: Ce 1 Ce þ ¼ qm b qm qe

Fig. 6 Variation of % removal of Pb (II) with various temperatures at different initial concentration of Pb (II) at biomass 10 g L−1 and pH 5

Consequently, it is expected that the ion exchange capacity had increased with increasing surface area of adsorbent. The rate of exchange is generally controlled by the rate of ion diffusion within the particle, and this is related to the size of particles (Kumar et al. 2006; Sari and Tuzen 2008). Effect of temperature All the experiments with Pb (II) were conducted in the temperature range of 25–40 °C, the percent removal of lead by C. fastigiata biomass decreases from 88.41 to 81.51 % with increase in temperature in the range 25–40 °C at (Fig. 6) initial concentration of 37.125 mg L−1. In most of the chemical reactions, the temperature is expected to activate the process increasing the heat or mass transport processes. Sorption capacity of the biomass has decreased with increase in temperature; rise in temperature has a tendency to desorb the adsorbed metal ions from the surface to the solution. The same trend was observed for other initial metal concentrations and was supported by the earlier report (Aksu 2001). Equilibrium isotherms Langmuir isotherm The equation proposed by Langmuir is universally applicable to chemisorption with some limitations involving physical

ð3Þ

Where qm (in milligrams per gram) is the maximum amount of the metal ion per unit weight of adsorbent to form a complete monolayer on the surface. “qe” is equilibrium adsorption capacity (in milligrams per gram), “Ce” is the equilibrium concentration of the adsorbate in solution (in milligrams per liter), and b is a constant which accounts for the affinity of the binding sites (in liters per milligram). qm represents the limiting adsorption capacity when the surface is fully covered with metal ions and helps in the evaluation of adsorption performance, particularly in cases where the sorbent did not reach its full saturation during contact. It is the most widely used simple two parameter equation (Langmuir 1918). From the plots between (Ce/qe) and Ce, the slope (1/qm) and the intercept (1/qmb) can be calculated. The Langmuir constant used to determine the suitability of the adsorbent to adsorbate by using dimensionless factor RL (Hall separation factor) is calculated as follows: RL ¼

1 1 þ bC

ð4Þ

0< RL 1 indicates unfavorable for adsorption, RL =1 indicates linear adsorption, and RL=0 indicates irreversible adsorption. The linearized Langmuir adsorption isotherms of Pb (II) onto C. fastigiata were obtained at different temperatures. Adsorption constants and correlation coefficients are shown in Table 1. Ce/qe versus Ce plot yielded a straight line with R2 (0.999) indicating that the sorption data could be represented by the Langmuir model. The higher adsorption capacity, qm(»1), indicated the strong electrostatic force of attraction, and b is a constant which accounts for the affinity of the binding sites (in

Table 1 Langmuir and Freundlich isotherm model parameters for Pb (II) onto C. fastigiata Temp. (°K)

qm (mg g−1)

qe (exp) (mg g−1)

b (Lmg−1)

C0 (mgL−1)

RL

R2

KF {(mgg−1) (mgL−1)n}

nf

R2

298 303 308 313

15.948 17.513 16.583 17.985

16.109 8.58 6.227 4.837

0.0653 0.0753 0.0898 0.1033

37.125 59.304 85.176 109.68

0.292 0.205 0.152 0.122

0.999 0.998 0.991 0.996

1.3424 1.084 0.8926 0.8005

1.5812 1.4825 1.3883 1.4164

0.993 0.997 0.999 0.993

Author's personal copy Environ Sci Pollut Res Table 2 Kinetic parameters for Pb (II) biosorption onto Caulerpa fastigiata Pseudo-second-order kinetics

W (g/ L)

qe (cal) (mgg−1)

2.5 16.447 5 8.7413 7.5 6.2775 10 4.8828

Pseudo-first-order kinetics

qe (exp) (mgg−1)

k2 (g/ R2 mg min)

k1 qe (mgg−1) (h−1)

R2

16.109 8.58 6.227 4.837

0.0319 0.0834 0.1799 0.2238

1.775 1.333 0.8564 0.772

0.933 0.929 0.928 0.974

0.9994 0.9997 0.9999 0.9998

0.022 0.031 0.021 0.022

liters per milligram). Moreover, the b values are 0.0653 and 0.0753 L mg−1, indicating that the biosorption capacity of C. fastigiata biomass for Pb (II) is higher. From the value of b, a dimensionless factor RL at different initial metal ion concentrations was calculated, and the values are shown in Table 1. The adsorption of Pb (II) on algal surface is, thus, a highly favorable process. Furthermore, it is observed that the sorption of lead is more favorable at higher Pb (II) initial concentration (109 mg L−1) than for the lower ones (37.125 mg L−1).

Freundlich isotherm An adsorption isotherm was proposed by Boedecker (Dąbrowski 2001) which was later modified by Freundlich (Freundlich 1926). The Freundlich adsorption equation can be written as follows: 1 n

qe ¼ K f C e f

ð5Þ

Fig. 8 Pseudo-second-order kinetic for biosorption of Pb (II) at 25 °C, 59.304 mg L−1, and pH 4.5

Taking logarithm on both sides, lnqe ¼ lnK f þ

1 lnC e nf

ð6Þ

Where “qe” is equilibrium adsorption capacity (in milligrams per gram), “Ce” is the equilibrium concentration of the adsorbate in solution, “Kf”, and nf are constants related to the adsorption process such as adsorption capacity and intensity, respectively. This empirical model has shown best fit for nonideal sorption on heterogeneous surfaces as well as multilayer sorption. The Freundlich isotherm is also more widely used but provides no information on the monolayer adsorption capacity, in contrast to the Langmuir model. Freundlich isotherm has been derived by assuming an exponentially decaying sorption site energy distribution. The coefficient of determination for this case is 0.993, and the values of nf and Kf (Table 1) are found to be 1.3424 g L−1 and 1.5812{(in milligrams per gram) (in milligrams per liter)n} at 25 °C. Freundlich constant nf between 1 and 10 indicates a trend more favorable for biosorption by macroalgae C. fastigiata. This is also suggestive that the metal ion under study could well be separated from its aqueous solution with high adsorption capacity. The values of high correlation coefficients indicated that the Pb (II) sorption data were very well represented by Freundlich model. The Freundlich constant nf was greater than 1, at all temperatures as well as initial metal Table 3 Thermodynamic parameters for Pb (II) biosorption onto C. fastigiata

Fig. 7 Pseudo-first-order kinetic plot for biosorption of Pb (II) at 25 °C, 59.304 mg L−1, and pH 4.5

Temp. (K)

C0 (mg L−1)

ΔH° (kJ mol−1)

ΔS° (J mol−1 K)

ΔG° (kJ mol−1)

298 303 308 313

37.125 59.304 85.176 109.68

−27.591 −24.526 −21.879 −19.991

−75.6848 −66.8465 −59.5648 −54.9638

−4.34579 −4.02446 −3.70314 −3.38181


concentrations representing that adsorption intensity of the sorbate on the sorbent surface was high, reflecting the favorable sorption even at high metal concentration.

expression as follows (Ho and Mckay 1999; 2000): dq ¼ k 2 ð qe − qt Þ 2 dt

ð9Þ

Adsorption kinetic models Several kinetic models have been proposed earlier to identify the mechanism of solute adsorption from aqueous solution onto the adsorbent. They are pseudo-first-order/Lagergren kinetic model, First order reversible kinetic model, Ritchie’ second order kinetic model, and pseudo-second-order kinetic model. In the present study, pseudo-first-order and pseudosecond-order kinetic models have been attempted to fit the present biosorption data (Table 2). Pseudo-first-order/Lagergren kinetic model The pseudo-first-order or Lagergren kinetic rate equation for the adsorption of liquid–solid system was derived based on solid adsorption capacity. It is one of the most widely used adsorption rate equations for adsorption of a solute from a liquid solution (Taqvi et al. 2006; Suddhodan and Mishra 2006). The pseudo-first-order kinetic equation can be expressed as follows: dq ¼ k 1 ð qe − qt Þ dt

ð7Þ

Where “qe” is the amount of solute adsorbed at equilibrium per unit mass of adsorbent (in milligrams per gram), “qt” is the amount of solute adsorbed at any given time “t”, and “k1” is the rate constant. By using the boundary conditions and simplifying, the Eq. 7 yields. lnðqe − qt Þ ¼ lnqe − k 1 t

ð8Þ

“k1” can be computed from the slope of the linear plot between ln (qe – qt) versus “t” for different adsorption parameters such as pH, temperature, adsorbate concentration, adsorbent dose, and particle size. The pseudo-first-order rate constant k1could be obtained from the slope of the plot between log (qe – qt) and time ‘t’. Figure 7 shows that the Lagergren pseudo-first-order kinetic plot does not fit well for the adsorption of Pb (II) onto C. fastigiata, as it does not follow a straight line.

On integration for boundary conditions when t=0 to t>0 and qt =0 to qt >0 and further simplifications, Eq. 9 becomes, t 1 1 ¼ þ t 2 qt k 2 qe qe

ð10Þ

The plot (Fig. 8) of t/qt versus t of Eq. (10) gave a linear relationship from which the qe and k2 values were determined. The rate constants and the correlation coefficients for pseudosecond-order kinetic model were calculated and summarized in Table 2. These values showed that the pseudo-second-order kinetic plot fits well with the adsorption data. The value of correlation coefficient R2 for the pseudo-second-order adsorption model is relatively high (>0.9994), and the adsorption capacities calculated by the model are also close to those determined by experiments. However, the values of R2 for the pseudo-first-order are not satisfactory. Therefore, it has been concluded that the pseudo-second-order adsorption model is more suitable to describe the adsorption kinetics of lead over this algal biomass.

Thermodynamic parameters Gibbs free energy ΔG is the thermodynamic criterion at constant P and T for deciding whether the chemical process does occur/proceed or not. The spontaneity of the reaction can also be judged by the sign and magnitude of ΔG°. A negative sign for ΔG° is an indication of the spontaneity of any chemical process. To design any chemical process system, one should have the knowledge of changes that are expected to occur during chemical reaction. The rate and extent of changes are more informative in the design of process equipment. In view of the above, analysis has been carried out on the values of thermodynamic parameters on the biosorption of Pb (II) by C. fastigiata. The thermodynamic parameters such as changes in standard free energy change ΔG°, Enthalpy ΔH°, and Entropy ΔS° for any given adsorption process could be determined from the Equations: ΔG0 ¼ −RTInK C

ð11Þ

Pseudo-second-order kinetic model In view of the above, the fitness of the sorption data was tested using pseudo-second-order reaction model. The pseudosecond-order reaction model could be expressed by the rate

Where ΔG° is the free energy change, expressed as joule per mole. Kc is the apparent equilibrium constant for the process. Kc can be derived from:


Fig. 9 FT-IR spectrum (a) unloaded algal biomass and (b) Pb (II)-loaded algal biomass

KC ¼ log

CS Ce

CS Ce

ð12Þ

¼−

ΔH 0 ΔS 0 þ 2:303RT R

ð13Þ

can be defined as “adsorption affinity”. Cs is the concentration of metal ion (in milligrams) in solid adsorbent. Ce is CS Ce

equilibrium metal concentration in milligrams per liter. The enthalpy changes (ΔH°) and entropy changes (ΔS°) for the adsorption process for all the initial metal concentrations in the aqueous solutions were obtained from the plots of log CCSe drawn against 1/T. The calculated thermodynamic data are compiled in Table 3. Large negative value for ΔG° indicates the spontaneity of biosorption process at a given temperature. The free energy values increased positively with increase in temperature for

Author's personal copy Environ Sci Pollut Res Fig. 10 SEM images of (a) before (b) lead-treated C. fastigiata powder

the adsorption of Pb (II), which shows that the spontaneity of the biosorption process reduces with increase in temperature.

The negative ΔH° values indicated the exothermic nature of the adsorption. The negative values of ΔS° suggested a

Fig. 11 XRD pattern of (a) untreated (b) and treated with Pb (II) C. fastigiata


decrease (Ahmet et al. 2008) in the randomness at solid/ solution interface during the adsorption of Pb (II) ions onto C. fastigiata.

2009; Leila et al. 2009; Suleman et al. 2009; Vítor et al. 2009; Rajesh et al. 2010; Munagapati et al. 2010). Scanning electron microscopy

Characterization of biomass Fourier transform infrared spectroscopy The shift of the bands and the changes in signal intensity allow the identification of the functional groups involved in metal sorption. The FTIR differences of spectra in pure algal biomass and in metal ion loaded algal biomass was compared to determine whether the observed differences are due to interaction of the metals ions with functional groups (Fig. 9a). The broad absorption peak at 3304.78 cm−1 indicates the presence of bounded -OH and -NH groups, and it is shifted (Fig. 9b) to 3343.07 cm−1, which may be due to the strong interaction with Pb+2 metal ion. These results indicated the involvement of these functional groups in biosorption process. The band at 2175.00 cm−1 is due to the –C≡N in the polyacrylonitrile and is shifted to 2,137.95 cm−1 for Pb+2-loaded biomass. The band at 1,654.99 cm−1 is due to the –C=O of carboxylic acid and is shifted to 1,639.01 cm−1 for Pb+2-loaded biomass. The band at 1,420.14, 1,248.04, and 1,033.82 cm−1 were due to the C–N, –SO3, and C–O in benzene ring groups, and the peaks were slightly shifted to 1,419.60, 1,254.04, and 1,034.87 cm−1, respectively, for Pb+2-loaded biomass. These results indicated the involvement of –SO3 functional group in biosorption process. The characteristic absorption peaks detected in the pure biomass at 2,936.13, 1,549.01, and 667.95 cm−1 were representing C–H stretching vibrations, amide (N–H), and N-containing bioligands were not detected in metal loaded biomass. Because of changes in metal-loaded biomass, it could be concluded that N–H was involved in binding the metal. The same results were shown in the case of Pleurotus ostreatus (Amna et al. 2011). The presence of siliceous from diatomaceous earth, in algal waste and composite material, can justify for the absorbance peak at 791.72 cm−1 (Si–C) and is shifted to the 785.29 cm−1. The bands present below 800 cm−1 are finger print zone of phosphate and sulfur functional groups and N containing bioligands. The significant changes in the wave numbers of the specific peaks suggested that hydroxyl, amide, bounded – OH, bounded –NH, C=H stretching vibrations, amide N-H bending vibrations, and C=O of carboxylic acid groups could be predominantly involved in the biosorption of Pb+2 onto C. fastigiata. Similar results were reported for the biosorption of different heavy metals on various algal species (Sibel et al.

SEM is a useful technique in the study of both the natural sorbent morphology and its modification derived from sorbate interactions. The electron interactions with the atoms of the sample produce signals that contain information about topography, morphology, and composition of the sample surface. It is evident from analysis that the surface areas of algal plant biomasses are uneven, heterogeneous with pores on the surfaces. SEM image of native biomass shows the number of pores with different diameters and different pore areas (Fig. 10a, b). Furthermore, pores facilitate the (Tong et al. 2011) good possibility for metal ions to be adsorbed. At ×500 magnification, the shiny needles of the biosorbents were focused, where an uneven surface texture along with a lot of irregular surface format was observed, pores were almost closed, and pore area also reduced. After Pb2+ biosorption (Fig. 10), the biopolymer surfaces showed a completely different morphology. Further irregular spikes are also accumulated on the surface, indicating the biosorption of metal by algal biomass powder. The similar condition of reduction in pore size and pore area were reported earlier on Zea Mays powder (Pritee and Shalini 2009). X-ray diffraction XRD patterns of macroalgae C. fastigiata before and after biosorption were depicted in the figure (Fig. 11), and it is indicating poor crystallinity of pure biomass (Chun-Shui et al. 2009; Adan et al. 2011). Furthermore, the shift in 2 and dspacing values were observed in Pb (II)-loaded biomass. From these observations, it could be concluded that there was a change in crystallinity of biomass C. fastigiata after the biosorption.

Conclusions Pb (II) biosorption for C. fastigiata is strongly affected by parameters such as contact time, initial metal ion concentration, pH, temperature, biosorbent dosage, and biosorbent particle size. The biosorption capacity (qe) of Pb (II) is 16.109 mg/g on C. fastigiata biomass. The particle size of 0.074 mm is best in adsorption of Pb (II) by C. fastigiata biomass. The favorable temperature is 25 °C, and an optimum pH for Pb (II)


biosorption was found to be 5.0. Freundlich isotherm model proved to be good fits for the experimental data of Pb (II) biosorption on C. fastigiata. Free energy change (ΔG°) with negative sign reflects the feasibility and spontaneous nature of the process. The enthalpy and entropy changes were calculated as −23.496 kJ mol−1 and −64.26498 J mol−1 K−1, respectively. The reaction is exothermic and randomness decreases at solid– liquid interface. The SEM studies showed Pb (II) biosorption on selective grains. The FTIR spectra indicated bands corresponding to – OH, COO−, -CH, C=C, C=S, and –C-C- groups. The XRD pattern of the Caulerpa fastigiata was mostly amorphous in nature. Hence, algae C. fastigiata is a promising biosorbent to remove Pb (II) metal ions from wastewaters. Acknowledgments One of the authors Mrs. B. Sarada is thankful to UGC SAP/DSA Phase-III for permission to use the AAS. K. Kishore Kumar thankfully acknowledge the financial assistance from the UGC, New Delhi (Dr. D. S. Kothari post-doctoral fellowship no. F.4-2/2006 (BSR)/13-176/2008) and also grateful to Prof. M.N.V. Prasad for providing the platform at University of Hyderabad.

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