Evolution of tubular copper sulfide nanostructures from copper(I)–metal organic precursor: a superior platform for the removal of Hg(II) and Pb(II) ions

Abstract

A wet-chemical reaction strategy has been adopted for the exclusive production of porous copper sulfide using an aqueous solution of copper chloride (CuCl₂·2H₂O), which acts as a precursor salt, and thioacetamide (TAA) with prolonged standing at room temperature (∼25 °C). The mixed phase copper(I) sulfide with tubular porous morphology has been derived from rod-shaped, white-colored Cu(I) metal organic complex through an auto degenerative hollowing. The importance of chloride ions for obtaining mixed phase copper sulfide has been proved unequivocally. Porous copper sulfide showed outstanding removal efficiency towards two toxic heavy metal ions, i.e. Hg(II) and Pb(II), which certifies the strong interaction between these metal ions and S²⁻ based on the SHAB (soft hard acid base) principle, and a cation exchange mechanism emerges out at the tubular surface as confirmed by various spectroscopic techniques. The as-synthesized product shows removal capacity of 3096 and 2787 mg g⁻¹ for Hg(II) and Pb(II), respectively. Thus, mixed phase copper(I) sulfide stands to be a highly effective substrate for environmental abatement application.

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Introduction

With the development of industries, indiscriminate disposal of organic pollutants and heavy metal ions have amplified the probability of water contamination as well as environmental pollution. Nowadays, heavy metal contamination in water has become a dangerous threat due to its negative impact on the environment and public health.^1,2 Heavy metal ions such as Hg(II) and Pb(II) ions are highly toxic water pollutants because of their easy accumulation in the living systems and non-biodegradable nature; moreover, they have some serious side effects and can cause permanent damage to various organs even at very low concentrations.^3,4 Therefore, various traditional techniques, such as chemical precipitation,⁵ adsorption,⁶ electrochemical method,⁷ membrane filtration,⁸ ion exchange⁹ and biological treatment,¹⁰ have been employed for the removal of such toxic metal ions.

Compared to the traditional adsorbents (such as activated carbon,¹¹ graphene,^12–14 and ion exchanged zeolites¹⁵) and various metal oxide nanomaterials (such as CeO₂, MgO, Al₂O₃, silica hollow sphere, and Fe₃O₄),^16,17 ion exchangers are highly desirable because of their cost-effectiveness and technical uses.¹⁸ Over the years, traditional ion exchangers (such as oxidic inorganic clay^19,20 or zeolites²¹) have dominated the water remediation process. Because of the presence of oxidic framework in the abovementioned traditional ion-exchangers and their low binding capability towards heavy soft metal ions, a new ion-exchanger has been developed with thiol group functionalized oxidic material to improve removal efficiency.^22,23 These materials show improved exchange affinity only for Hg(II) ion and not for other heavy metal ions. Moreover, researchers have developed a sulfur functionalized soft basic framework for the removal of other soft heavy metal ions.^24,25

In this study, we have synthesized efficient sulfur containing porous adsorbent, which works exceptionally well for binding Hg(II), Pb(II) ion in its porous interior. Herein, soft–soft interaction between sulfur donor and toxic metal ions play an important role for admirably high binding efficiency, leading to environmental remediation.

Copper sulfide has attracted much attention over the years for its various unique properties,²⁶ for which it has been applied in numerous fields such as optical filters, room temperature ammonia gas sensing, lithium rechargeable batteries, and photocatalysis.^27–30 Herein, we have demonstrated a solution phase synthetic protocol for porous copper sulfide by mixing an aqueous solution of CuCl₂ and thioacetamide (TAA) under ambient conditions at room temperature after keeping the reaction mixture for several days. In contrast to the earlier report,³¹ a mixture of chalcocite (Cu₂S) and covellite (CuS) is produced in our case instead of CuS. The porous nature of the as-synthesized material helps in achieving a high adsorption phenomenon for Hg(II) and Pb(II) ions, whereas the facial ion exchange mechanism provides for their efficient removal capacity. Thus, the proposed strategy for copper sulfide results in a low-cost material as an effective adsorbent for both Hg(II) and Pb(II) removal from aqueous solutions.

Experimental details

Materials and instruments

The related information is presented as S1 in the ESI.†

Synthesis of copper sulfide

In a typical procedure, solutions A and B were prepared by dissolving 4 mmol of thioacetamide (TAA) and 4 mmol of copper chloride (CuCl₂·2H₂O) in 100 mL water in two separate beakers. Then, solution B was added dropwise to solution A in a beaker with stirring. A white gelatinous product appeared slowly, and the precipitation was completed in about 24 h. The white precipitate transformed quantitatively into a dark green colored product on standing for 14 days. Then, the precipitate was removed from the bottom of the beaker and washed several times with distilled water and then with absolute alcohol and dried in vacuum at room temperature.

Adsorption kinetics

The adsorption kinetics was studied individually with the single metal ion in question employing 10 mg of the as-prepared copper sulfide in 200 mL 10⁻³ M Hg(II) or Pb(II) solution at pH ∼2.8 and 4.7, respectively. After a pre-determined time interval, a part of the reaction mixture was centrifuged. To determine the concentration of left over Hg(II) ion in 4 mL of centrifugate, 0.1 mL 10⁻³ M diphenyl carbazide³² was added. Then, the absorbance was measured using a UV-vis spectrophotometer at λ_max = 521 nm. In a similar way, 0.5 mL 10⁻³ M xylenol orange³³ was added into a 3 mL aliquot for absorbance measurement at λ_max = 579 nm for Pb(II) ion, and adsorption capacity was calculated using the following equation:¹⁴

Adsorption capacity = (C₀ − C_f) × V × M_M(II)/W_s (1)

where C₀ and C_f are the initial and final concentration of aqueous metal ion solution in M, respectively, V (L) is total volume, M_M(II) is the atomic weight of adsorbate metal ion in mg and W_s (g) is the weight of copper sulfide. The C_f value is evaluated from a standard graph of the corresponding metal ions (for details see Fig. S1, ESI†).

Adsorption isotherm

The adsorption isotherms were studied by conducting a set of adsorption experiments with the as-prepared copper sulfide and corresponding metal ions in solution. Copper sulfide (10 mg) was added individually to 200 mL aqueous solution of Hg(II) with seven variable concentrations ranging from 2 × 10⁻⁴ to 2 × 10⁻³ M and stirred for 3 h to reach equilibrium, and then C_f value in each case was determined as per the abovementioned procedure. For Pb(II) ion, the same procedure was followed using the same concentration range of Pb(II) for six different sets of aqueous solutions.

Results and discussion

In the reaction mixture, the addition of TAA into CuCl₂ aqueous solution caused its gradual color change from blue to colorless, which suggests the reduction of Cu(II) to Cu(I) at room temperature. After 2–3 minutes of mixing, the resulting reaction mixture produced a white gelatinous precipitate, which slowly converted into an insoluble greenish product as indicated in Scheme 1.

Scheme 1 Schematic of stepwise formation of copper sulfide at room temperature.

The X-ray diffraction pattern of the tubular greenish product indicates the crystalline mixed phase copper sulfide, i.e. chalcocite and covellite (Fig. 1a). It should be noted that most of the diffraction peaks match with JCPDS data file of chalcocite (Cu₂S) [72-1071], whereas few peaks match with the JCPDS data file of covellite (CuS) [79-2321]. Note that in both these phases, the valency of copper is (+1). Previous studies have reported that covellite (CuS) is actually a composition of (Cu¹⁺)₃(S₂²⁻)(S¹⁻)³⁴ or (Cu¹⁺)₃(S₂⁻)(S²⁻).³⁵ Here, in both these cases, the valency of Cu and S are (+1) and (−1), respectively. Recently, Xie et al. also proved by various techniques that the valency of Cu is (+1) in covellite.³⁶ Note that Cu₂S adopts a tetragonal P4₃2₁2 structure with lattice parameter a = 3.996 Å, c = 11.28 Å, whereas CuS has a close resemblance to hexagonal P6₃/mnc CuS with lattice parameter a = 3.788 Å, c = 16.33 Å. For a complete analytical feedback on the crystallographic nature of our as-synthesized copper sulfide sample, we have performed Rietveld refinement analysis. Herein, the tetragonal Cu₂S and hexagonal CuS were used as the starting model for two-phase Rietveld refinement. The Rietveld refinement (Fig. S2†) shows estimated matching in peak positions of these two phases with the XRD pattern of the as-synthesized copper sulfide with little differences in intensity profiles between the experimental and theoretical. Finally, a summary of peak positions (with tetragonal Cu₂S and hexagonal CuS phases) has been incorporated in Table S1† for the sake of clarification. In the XRD pattern, few low intense peaks are present, which show a similarity with orthogonal Cu₂S [JCPDS no. 02-1272]. The comparatively smaller size of the crystallites is the basis of broader diffraction peaks than the standard graph. Even after storing the adsorbent material for one month, the XRD pattern shows nearly the same peaks of mixed phase copper sulfide, which indicates that the product is highly stable in nature (please see Fig. S3, ESI†). Again for further study of the composition, the as-prepared product was characterized using XPS analysis, where the peak positioned at 932.1 eV and 952.1 eV are attributed to Cu 2p_3/2 and Cu 2p_1/2 electronic state of Cu(I), respectively, as indicated in Fig. 1b. One small peak at ∼933 eV (Fig. 1c), corresponding to Cu(II) 2p_3/2, indicates that the presence of Cu(II) is negligible compared to predominant Cu(I). Thus both XRD and XPS analyses confirmed the presence of Cu(I) state in the as-synthesized copper sulphide predominantly.

Fig. 1 (a) XRD pattern, (b and c) wide and narrow range XPS spectra.

Morphology of the green product was at first investigated using a Field Emission Scanning Electron Microscope (FESEM). The FESEM image in Fig. 2a confirmed the synthesis of porous tubular product after 14 days. Note that the length of a tube was 20–30 μm and the diameter of the tube was 1.5–2 μm. Open ends of tubes were clearly seen from the FESEM image in Fig. 2b. The 600–700 nm thick walls of the tube were built up using conjointly stacked microflowers, which are a congregation of nanoflakes with thicknesses of 15–25 nm (Fig. 2c). For the further characterization of morphology, TEM and HRTEM analyses have been conducted with the as-prepared product. The TEM image in Fig. 2d represents a tubular structure having micrometer size long with thick walls. Prolong ultrasonication of tubular structure leads to microflowers and nanoflakes of ∼1 μm diameter as shown in Fig. 2e. The selected area electron diffraction (SAED) pattern in Fig. 2f shows the well resolved diffraction spots, indicating its crystalline nature. The HRTEM image (Fig. 2g) of a typical flake illustrates clear lattice spacing of 0.275 nm due to the (111) plane (2θ = 32.64°) of Cu₂S (JCPDS no. 72-1071). Thus, HRTEM analysis confirmed that the difference in intensity of experimental and theoretical XRD patterns is caused by the preferred orientation of the crystal planes.

Fig. 2 (a–c) FESEM images in different magnification, (d and e) TEM images, (f) SAED pattern and (g) HRTEM image of the as-synthesized product.

The investigation of the porous nature and specific surface area of the as-synthesized copper sulfide was carried out using Brunauer–Emmett–Teller (BET) gas sorptometry. The nitrogen adsorption–desorption isotherm of this porous material is shown in Fig. 3a. This isotherm could be categorized as type-IV with a clear H₁-type hysteresis loop in the 0.5–1.0 p/p₀ range, indicating the mesoporous nature of the material.^37,38 The material shows N₂ gas uptake capacity of 336.11 cc g⁻¹, and the calculated specific surface area is found to be 62.308 m² g⁻¹ with a total pore volume of 5.182 × 10⁻¹ cc g⁻¹. The average pore diameter of the sample is observed to be 2.73 nm in accordance with the Barrett–Joyner–Halenda (BJH) plots calculation. The pore size distribution graph is shown in Fig. 3b, and the pore diameter supports the mesoporous nature of the sample. Therefore, the surface area of the compound (62.302 m² g⁻¹) is related with the small pores (mesopores) on the sample wall and not the micro-sized tubes confirmed by this analysis.

Fig. 3 (a) Nitrogen adsorption and desorption isotherm and (b) pore size distribution plot of the as-synthesized porous copper sulfide.

FTIR analyses were performed to investigate the bonding nature of the intermediates and also the final product. In the metal organic complex, the bonding between Cu(I) and TAA is driven by electron donation from sulfur to copper(I), whereas no signature of the stretching vibration of N–H and C–N bonds was identified in the FTIR spectrum of the final product (Fig. S4†).

To investigate the formation mechanism of the copper sulfide tubular structure with hierarchical porous wall in detail, the intermediates were examined using various spectroscopic techniques (Fig. S5†). It has been reported³¹ that CuCl₂ in solution yields a white colored Cu(I) metal organic complex with the addition of TAA. On standing at room temperature for 14 days, rod-like, white-colored, metal organic complex transformed into porous tubular copper sulfide via an auto degenerative hollowing process. In contrast to the previous report, here mixed phase copper sulfide i.e., chalcocite and covellite, are produced simultaneously from the Cu(I) metal organic complex. The complete transformation from Cu(I) complex to Cu(ii)-sulfide was blocked by performing the reaction at room temperature under heating conditions. Herein, the counter anion in the precursor salt and ligand–metal concentration ratio plays an important role for the porous copper sulfide synthesis (detailed growth mechanism is discussed in ESI†).

Heavy metal ions removal

The removal of toxic heavy metal ions (Hg(II) and Pb(II)) from water by hierarchical porous copper sulfide material was investigated using UV-vis absorbance spectroscopy as the metal ions are well known to form colored complexes with respective organic ligands.

Kinetic study of the adsorption process enlightens us about the whole mechanism of the adsorption process. In this case, experimental data of adsorption kinetics of both the metal ions (Fig. 4a and d) are in good agreement with the pseudo second order³⁹ equation:

t/q_t = 1/(k₂q_e²) + t/q_e (2)

where q_e and q_t are the adsorption capacity in mg g⁻¹ at equilibrium and time t (min), and k₂ in min⁻¹ is the rate constant of the pseudo second order adsorption process. The correlation coefficient (R²) value is 0.9999, and the calculated q_e value is nearly equal to the experimental value, which suggests that the adsorption process follows the pseudo second order model. Note that the other obtained kinetic parameters are summarized in Table 1.

Fig. 4 Pseudo-second-order kinetic plots (a and d), Langmuir isotherm (b and e) plots, adsorption isotherm (c and f) plots of Hg(II) and Pb(II) obtained using copper sulfide.

Table 1 Adsorption parameters of Pb(II) and Hg(II) adsorption

Material	Pseudo-second-order			Langmuir isotherm
	K2 (g mg^−1 min^−1)	qe,cal (mg g^−1)	R^2	qe,exp (mg g^−1)	KL (L mg^−1)	qm (mg g^−1)	R^2
Pb(II)	2.577 × 10^−4	2715	0.9999	2649	0.4335	2787	0.9999
Hg(II)	3.425 × 10^−6	2961	0.9999	2916	0.0930	3096	0.9978

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The adsorption mechanism of Hg(II) and Pb(II) onto copper sulfide could be easily understood by the study of adsorption isotherms shown in Fig. 4b and e. Two types of adsorption isotherms are generally used to describe the mechanism of adsorption. The Langmuir adsorption isotherm model⁴⁰ assumes that adsorption takes place on a homogeneous surface and is expressed by the following equation:

C_e/q_e = 1/(K_Lq_m) + C_e/q_m (3)

where q_e is the adsorption capacity (mg g⁻¹) at equilibrium, C_e is the concentration (mg L⁻¹) of Hg(II) or Pb(II) ions at equilibrium condition, K_L is a constant and is related to the energy of adsorption (L mg⁻¹) and q_m is the Langmuir monolayer adsorption capacity (mg g⁻¹).

In this case, the experimental data are in an excellent agreement with the Langmuir isotherm (Table 1). From the graph, the value of the maximum adsorption capacity (q_m) of Hg(II) by copper sulfide is calculated and it has a value very close to q_e. Moreover, the same process was followed for the q_m calculation of Pb(II) ion adsorption by copper sulfide and the value is nearly equal to the q_e value. Thus, the result confirms that the adsorption process followed the Langmuir model, i.e. monolayer adsorption. Maximum adsorption capacity of the as-synthesized copper sulfide for both Hg(II) and Pb(II) was calculated from the plot of q_e (mg g⁻¹) against C_e (mg L⁻¹) in each case (Fig. 4c and f). Multiple sets of varied initial concentration of the metal ions were mixed with the same amount of as-synthesized adsorbent (0.05 g L⁻¹) and stirred for 3 h to reach equilibrium. After complete adsorption, each solution was centrifuged and the equilibrium concentration (C_e) of each set was analyzed by UV-vis spectroscopy as described previously. Similarly, adsorption capacities at equilibrium (q_e) of each set for both the metal ions were calculated and plotted against their corresponding equilibrium concentration to obtain the maximum adsorption capacities. In this case, the maximum adsorption capacity is found to be 3096 and 2787 mg g⁻¹ for Hg(II) and Pb(II), respectively; moreover, as shown in Tables 2 and 3, the results are much higher than those reported previously,^41–48 including those of carbon based compounds, which suggests that our synthesized material is a practically applicable adsorbent for the removal of these toxic metal ions.

Table 2 Maximum adsorption capacity of Hg(II) ion by various adsorbents

Adsorbent	qe (mg g^−1)	References
MCM-41	1000	41
PMO-ICS-75	1800	42
PPy-RGO	980	43
MNPs-DTC	47.870	44
MNPs	59.453	44
BTESPTS	2710	45
Copper sulfide	3096	In this work

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Table 3 Maximum adsorption capacity of Pb(II) ion by various adsorbents

Adsorbent	qe (mg g^−1)	References
EDTA-GO	479 ± 46	12
RGO/PAM	1000	33
MgO	1980	46
Zinc silicate	210	47
γ-AlOOH	124.2	48
Copper sulfide	2787	In this work

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Quantitative study of Hg(II) and Pb(II) ions removal from water

In order to understand the excellent removal efficiency, we performed further investigation with only copper sulfide after the removal of Hg(II) and Pb(II) metal ions. Herein, the metal ions removal reaction occurred via the ion exchange pathway. We have monitored the change in Hg(II) or Pb(II) concentration with time (Fig. 5). It should be noted that with time the concentration of heavy metal ions decreased with increase in Cu(II) ion concentration in the solution. Initially, the exchange processes for both cases were found to be very fast, as presented in Fig. 5, and after that they became much slower, which may be due to the deficiency of active sites at the later stage. The gradual release of Cu(II) ions in solution with time supports the fact that the removal occurs through an ion exchange pathway predominantly rather than any physisorption or chemisorption mechanism. The exchanged Cu(I) may be oxidized to Cu(II) by the oxygen dissolved in water.

Fig. 5 Decrease in (a) Hg(II) and (b) Pb(II) ions concentration as a function of contact time.

After the removal of Hg(II), the copper sulfide adsorbent is converted to HgS as confirmed by XRD and elemental area mapping analysis (Fig. 6). The sizes (ionic radii of S²⁻, Cu(I), Hg(II) and Pb(II) are 170, 91, 116 and 133 pm, respectively) of the corresponding ions help to understand the formation of HgS or PbS based on the SHAB principle. Tabulated formation energies⁴⁹ for a number of binary metal chalcogenides support the thermodynamic feasibility of exchange reaction between Hg(II) or Pb(II) and copper sulfide with a negative ΔG_reaction value. The value of ΔG_reaction in the case of HgS and PbS formation reaction from copper sulfide are −21.7 and −63.5 kJ mol⁻¹, respectively.

Fig. 6 (a) XRD pattern and (b) area mapping of Hg(II) adsorbed copper sulfide.

Now heavy metal adsorbed copper sulfide interestingly depicts a small amount of Cu(I) sulfide species in the XRD spectra. This clearly indicates a fair competition of the incoming heavy metal ion with Cu(I) sulfide. The XRD result revealed that the fast interaction may have occurred as follows:

Cu(I)S (solid) + Hg(II) (soln.) → HgS (solid) + Cu(II) (soln.)

Therefore, our synthesized material shows high removal capacity via adsorption of metal ions on its porous structure followed the ion-exchange mechanism on its surface. Numerous reports on metal ions removal have been based on ion-exchange mechanism.^9,49–51 Note that Brock et al.⁵⁰ and Yan et al.⁵¹ have reported efficient removal of Pb(II) and Hg(II) ions respectively from polluted water using metal sulphides. In our study removal of toxic metal ions by copper sulphide follows the same pathway.

Fig. S11 (ESI†) also corroborates the abovementioned mechanism for Pb(II) removal. With conversion of Hg(II) and Pb(II) into corresponding sulfides, a small amount of copper ion as Cu(II) leaches out into the solution (∼22% in case of Hg(II) ion adsorption and ∼18% for Pb(II) ion adsorption). The maximum contamination level of mercury, lead and copper are 0.002, 0.015, 1.3 ppm, respectively, in drinking water as per the regulation of the United States Environmental Protection Agency (EPA). This information indicates that the toxicity of Cu(II) is insignificant compared to highly toxic mercury and lead. It is worth mentioning that in both the cases the removal capacities are exceptionally high from a catalytic amount of the adsorbent. A similar adsorption experiment was compared with non-porous, spherical Cu(II) sulfide synthesized by mixing TAA and CuSO₄ for the removal of Hg(II) and Pb(II). In both these cases, the adsorption capacities are much lower than that of tubular porous copper sulfide after 3 h of adsorption (see Fig. S12, ESI†). Though this is a chemical exchange reaction, our materials follow the Langmuir isotherm similar to the previous report of Song et al.⁴⁶ This result confirms that the excellent and admirably high adsorption behavior of copper sulfide is not only due to its porous nature but also the strong interaction between S²⁻ and heavy metal ions, which plays the most important role.

Conclusion

In summary, a template-free, low-cost, room temperature synthetic procedure has been reported for the production of porous, highly stable, tubular nanostructure of copper sulfide in an aqueous medium, which provides exceptionally high removal capacity towards Hg(II) and Pb(II) ions with a maximum capacity of 3096 and 2787 mg g⁻¹, respectively. The exchange reaction mechanism between Cu(I) and Hg(II) ions has been proposed based on the SHAB principle. Both the porous nature (adsorption) and exchange reaction are the main factors for this very high adsorption behavior of copper sulfide towards heavy metal ions in comparison to carbonaceous materials. Thus, copper sulfide can be a good alternative to carbon-based classical adsorbents for wastewater treatment.

Acknowledgements

We are thankful to CSIR, New Delhi; BRNS, Mumbai; and Indian Institute of Technology, Kharagpur, for financial and instrumental support.

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Footnote

† Electronic supplementary information (ESI): Description of materials and instruments, UV-vis spectra for standard graph, FTIR spectrum, elemental area mapping, FESEM, XRD analysis of the products and Pb adsorbed copper(I) sulfide, and adsorption data of heavy metal removal by Cu(II) sulfide. See DOI: 10.1039/c4ra09999k

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