Electrochemical sensor based on a thin film of organokaolinite material modified glassy carbon electrode (GCE) and application to the simultaneous sensitive detection of Pb²⁺ and Cu²⁺ ions in contaminated water

Abstract

A pure natural kaolinite (K) was modified by intercalation and grafting into the interlayer space of 1-(2-hydroxyethyl)piperazine (HEP). The resulting organo-inorganic hybrid material (referred as KHEP) was characterised using X-ray diffraction (XRD), Fourier transform infrared (FTIR) and ²⁹Si nuclear magnetic resonance (NMR) techniques. The electrochemical characterisation of K and KHEP materials was carried out by studying the electrochemical behaviour using cyclic voltammetry of [Fe(CN)₆]³⁻ ions on the surface of a glassy carbon electrode (GCE) modified with a film of each of these materials (GCE/K and GCE/KHEP). The GCE/KHEP organokaolinite film electrode was successfully applied for the simultaneous detection of Pb(II) and Cu(II) in contaminated media. The peak currents of Pb(II) and Cu(II) obtained under the same optimal conditions on GCE/KHEP were much more intense than those obtained on GCE/K and bare GCE. Under optimal conditions, the influence of the concentration of Cu²⁺ and Pb²⁺ ions on the peak currents was studied for several concentration ranges. The obtained linear calibration curves were then used to calculate the different limits of detection (LOD) on the basis of a signal-to-noise ratio of 3. Thus, in simultaneous detection and for the concentration range from 0.02 to 0.12 µmol L⁻¹ (pH accumulation medium 5.0) and for each of the ions analysed, LODs of 1.942 nmol L⁻¹ for Cu(II) and a LOD of 1.072 nmol L⁻¹ for Pb(II) were obtained. The GCE/KHEP sensor developed in this work was successfully applied for the simultaneous detection of Pb(II) and Cu(II) in real samples.

Graphical abstract

ASSWV signal corresponding to the simultaneous detection of Pb²⁺ and Cu²⁺ ions recorded on unmodified GCE (black curve), GCE/K; (red curve) and GCE/KHEP (blue curve) in 0.2 mol L⁻¹ NaNO₃ (with pH adjusted at 2.5) after 7 min of accumulation in aqueous medium (pH 5.0) of each ion at the concentration at 2.0 µmol L⁻¹. Other experimental conditions are as follows: t_elec: 80 s, E_elec: − 0.90 V, Potential amplitude: 0.05 V, Step potential (V): 0.005 V and Frequency: 50 Hz.

1 Introduction

The increasing proliferation of mining activities throughout the world, in conjunction with the poor management of waste water from electronic industries (batteries, accumulators) and the accidental or deliberate discharge into the environment of untreated waste from the paint industry, all contribute to heavy metal contamination and pollution of the environment [1, 2]. Heavy metals are highly toxic to humans, and some of them are even toxic in trace amounts. The toxicity of heavy metals is directly linked to their high reactivity with living matter and, above all, their low biodegradability [2,3,4,5,6,7]. Consumed even in small doses through contaminated water or contaminated foods such as fish and vegetables, heavy metals have the capacity to accumulate progressively in living organisms through bioaccumulation processes, causing a number of diseases such as anaemia, cancer, high blood pressure, neurological disorders, certain congenital malformations and genetic mutations [8,9,10]. Lead and copper are typical examples of these heavy metals. The World Health Organisation (WHO) has set the maximum acceptable concentration in drinking water at 4.82 × 10⁻⁸ mol L⁻¹ for lead and 3.14 × 10⁻⁵ mol L⁻¹ for copper [11,12,13,14,15].

Detecting and controlling the quantities of these two heavy metals in the various compartments of the environment, and in particular in contaminated aqueous media, is therefore a major challenge for analytical chemists if they are to preserve a healthy environment. Conventional methods for detecting heavy metals, such as atomic adsorption spectrometry, atomic fluorescence spectrometry and plasma-coupled mass spectroscopy [15,16,17,18], have shown their limits, mainly due to the high cost of equipment and the maintenance involved. These conventional techniques also have long analysis times. Electrochemical methods used in recent years for the detection and electroanalysis of heavy metals in contaminated environments have proved to be more interesting, with equipment that is much less expensive, easy to use and maintain and, above all, shorter analysis times. The efficiency, sensitivity and selectivity of these analysis methods depend mainly on the nature of the working electrode.

Typical working electrodes include glassy carbon electrode (GCE) and carbon paste electrode (CPE). In order to improve the sensitivity and selectivity of these conventional working electrodes for detecting heavy metals, and consequently improve electrochemical analysis techniques, several types of natural or synthetic materials have been developed for their modification [7, 10, 19,20,21,22,23,24,25,26,27,28,29]. These materials include resins, zeolites, silicas [29,30,31,32,33,34], hydroxyapatites [27], increasingly geopolymers [28] and above all clays materials because of the nature of their surface functional groups, their good adsorption capacity and their microporous structure [2, 3, 19, 24, 25, 35, 36]. For this last family of electrode materials developed extensively in recent years, one can cite natural smectites or smectites modified by grafting of amine [7, 23, 37,38,39,40] or thiol [3, 7, 23, 40,41,42] groups, which have a high affinity with certain heavy metals by forming coordination bonds with them. Modified electrodes based on modified clays have shown good sensitivity and selectivity for detecting heavy metal ions in aqueous solution [24, 37,38,39, 41, 42].

Kaolinite belongs to another large family of clay minerals, in particular the family of 1/1-type phyllite clays. Contrary to smectites, this clay mineral is much more available in nature, especially in Cameroon (Central Africa) where it possesses many deposits. However, its use in electrochemistry as an electrode material and especially in electroanalysis is more recent [39, 41, 43,44,45,46,47]. In a very recent work by our research team, we successfully modified a natural kaolinite by grafting an amine group and the organokaolinite obtained was applied to the development of a new electrode that we used with great success for the sensitive and selective detection of lead under certain experimental conditions in artificial contaminated aqueous media and in real contaminated aqueous media [39]. With this sensor, a detection limit of the order of 10⁻⁹ mol L⁻¹ was determined [39]. In addition to this work, we also found that the organokaolinite-based sensor we had used in this recent work [39], under different experimental conditions, was also sensitive to the electrochemical detection of Cu(II) in aqueous media. This made it possible to understand that, the amine groups present in the film of this organokaolinite could have either a high affinity with Cu(II) or a high affinity with Pb(II) in aqueous solution, depending on experimental conditions.

Considering that lead and copper are generally found together and at the same time in the same contaminated and polluted aqueous media, we thought it would be interesting, especially in order to reduce analysis time and the number of manipulations, to study the possibility of simultaneous electrochemical detection of these two heavy metals in a contaminated aqueous medium. This approach of simultaneous voltammetric detection of several heavy metal ions in the same contamination medium has attracted a great deal of interest from electrochemical analysts in recent years because it reduces the quantity of reagents used and the number of experiments compared with individual detection. In the literature, there are some previous works on the simultaneous electrochemical detection of several heavy metal ions in contaminated aqueous media using conventional electrodes modified with organosmectite materials [2, 24, 38, 48]. According to our bibliographic research, no work on the simultaneous voltammetric detection of heavy metals has yet been carried out on a Glassy Carbon Electrode (GCE) modified with an organokaolinite film. In this work, we propose to use the organokaolinite synthesised in a previous and recent work [39], by intercalation and grafting of 1-(2-hydroxyethyl)piperazine in a pure natural kaolinite, for the implementation of an electrochemical sensor applied to the simultaneous detection of lead and copper in a contaminated aqueous medium and in real contaminated water.

2 Experimental

2.1 Raw kaolinite material, reagents, solutions and solvents

The natural clay used in this work is a reference kaolinite, well-crystallized, known by the commercial name KGa-1b [39]. The bio-organic molecule used for the chemical modification of pure kaolinite (K) was 1-(2-hydroxyethyl)piperazine (HEP) (99%) supplied by Sigma-Aldrich. The water used for all the experiments was of ultrapure analytical quality obtained using a water deioniser (18 mΩ cm⁻¹) which was filtered through a millipore milliQ water purification system. KCl was purchased from Sigma-Aldrich. K₃Fe(CN)₆ (≥ 99% Prolabo) was reagent grade and used as received.

For the electrochemical characterisation of pure kaolinite and organokaolinite films, the supporting electrolyte used throughout this study was KCl solution (0.10 mol L⁻¹). The supporting electrolyte used for the simultaneous electrochemical detection of copper and lead was prepared at 0.20 mol L⁻¹ from NaNO₃ purchased from Sigma-Aldrich. To adjust the pH of the various solutions used in experiments, small controlled quantities of concentrated solutions of NaOH (BDH) and HNO₃ (70%, Normapur) were added to the supporting electrolyte. The Pb²⁺ solution analysed was prepared from a 4.82 mol L⁻¹ concentrated Pb²⁺ standard solution. The Cu²⁺ solution was obtained by dissolving Cu(CH₃COO)₂ (Sigma-Aldrich). Dimethylsulfoxide (DMSO) was purchased from Fisher, dioxane (99%) and isopropanol (99%) were purchased from Sigma-Adrich. All other chemicals and solvents used for experiments in this work were of analytical grade.

2.2 Preparation of organokaolinite

The starting kaolinite KGa-1b, before its chemical modification, was purified following a procedure already well described in previous works [39, 47, 49,50,51,52]. After purification, the kaolinite obtained was symbolised by K. The organokaolinite synthesis procedure used in this work was well described in a previous work published by Ngassa et al. [39]. The synthesis procedure involved two main stages. The first step consisted of intercalating DMSO molecules in the interlayer space of pure kaolinite [39, 43,44,45,46,47, 49,50,51], and the resulting material is referred to as K-DMSO. The second step consisted of replacing the DMSO molecules in the interlayer space by intercalation and grafting of 1-(2-hydroxyethyl)piperazine molecules as shown in Scheme 1. In practice, this synthesis can be summarised as follows: 20 g of HEP was placed in a 100 mL round-bottom flask then heated until a viscous liquid was obtained, then the mixture was refluxed for 24 h at a constant temperature of 170 °C. The temperature of the synthesis medium was again gradually increased to 250 °C and kept constant under reflux for 24 h. The aim of this second increase in the temperature of the synthesis medium being to improve the grafting reaction of 1-(2-hydroxyethyl)piperazine molecules. The solid residue obtained at the end of this process was first washed with isopropanol (2 × 60 mL) and then dried at room temperature. In a second step, the solid obtained and dried in the first wash was dispersed again in 100 mL of deionised water and stirred for 24 h. The solid residue from this second wash was recovered by centrifugation and washed with isopropanol (60 mL). The organokaolinite obtained at the end of this synthesis was dried at room temperature and denoted KHEP.

Scheme 1

Steps involved in the modification of natural kaolinite (K): intercalation of dimethylsulfoxide (DMSO) between the layers of pure kaolinite leading to K-DMSO, then intercalation followed by grafting of 1-(2-hydroxyethyl)piperazine (HEP) molecules within the inter lamellar space of the kaolinite

2.3 Characterisation techniques

The powdered X-ray diffraction (XRD) patterns of pure kaolinite and organokaolinite were recorded at room temperature using a Rigaku Ultima IV Diffractometer (Rigaku, Tokyo, Japan) with Cu Kα radiation (λ = 1.5406 Å) using a generator voltage of 45 kV and current of 40 mA with a scan rate of 2.0 deg.min⁻¹. All XRD diffractograms were recorded on glass slides.

FTIR-ATR spectra were registered at room temperature on a Nicolet 6700 FTIR, in the spectral range from 4500 to 650 cm⁻¹ with a resolution of 4 cm⁻¹. A spectrometer equipped with the Opus/Mentor software provides an intuitive interface and facilitates the analysis of spectra.

Solid state ²⁹Si NMR CP/MAS spectra were collected on a Bruker AVANCE 200 NMR spectrometer spinning at 4500 Hz operating at 39 MHz. The ²⁹Si chemical shifts were referenced to TMS at 0 ppm using the high frequency signal of tetrakis(trimethyl)silane at 29.9 ppm as a secondary standard.

2.4 Preparation of the working electrode, electrochemical device and procedures

2.4.1 Preparation of the working electrode

The working electrode used for electrochemical experiments was a glassy carbon electrode (GCE) modified with a thin film of natural clay or organoclay prepared by the drop coating technique. This film electrode preparation technique is described in previous works [37, 39, 41, 43,44,45,46, 48] and proceeds as follows: the GCE bare was first polished with an aqueous dispersion of 0.05 μm alumina, then rinsed with deionised water, passed through an ultrasonic bath for 10 min and left to dry at room temperature. The kaolinite or organokaolinite film was deposited on the surface of the GCE by drop coating. Thus, 15 µL aliquots of an aqueous suspension of clay material (4 g L⁻¹) which were carefully deposited by drop coating on the active surface of the GCE (3 mm in diameter). The natural kaolinite thin film modified GCE denoted GCE/K or organokaolinite thin film modified GCE denoted GCE/KHEP were stored at room temperature for about 4 h to ensure complete drying before use for electrochemical experiments as working electrodes. The electrochemical cell also contained a platinum plate as counter electrode and an Ag/AgCl (KCl saturated as reference electrode. It is important to notice that all the potential values for electrochemical experiments in this work are referenced to this reference electrode.

2.4.2 Electrochemical device and procedures

The procedure for simultaneous electrochemical detection of Pb²⁺ and Cu²⁺ ions by stripping voltammetry is described in previous works [35,36,37,38,39, 41, 42, 47]. Briefly, this procedure involved two mains successive steps. The first step corresponds to the accumulation of Pb²⁺ and Cu²⁺ ions under open circuit conditions (zero potential at working electrode) on the surface of the working electrode: the working electrode (GCE/K or GCE/KHEP) was immersed in a constantly stirring aqueous solution containing Pb²⁺ and Cu²⁺ ions (accumulation medium) at precise concentrations for a given duration (accumulation time). The second stage corresponded to voltammetric detection itself, which was then achieved by Anodic Stripping Square Wave Voltammetry mode (ASSWV). After a given accumulation time, the working electrode was rinsed with deionised water and immediately immersed in another medium, solution of NaNO₃ concentrated 0.2 mol L⁻¹ at pH adjusted at 2.5 (detection medium). A sufficient electrolysis potential (E_elec = E_p) was then imposed on the working electrode for a given time (T_elec electrolysis time) to reduce the previously accumulated heavy metal ions, followed by reoxidation during which the electrochemical detection current was recorded. After each simultaneous ˝accumulation-detection˝ sequence of Cu²⁺ and Pb²⁺ ions, the surface of the working electrode was renewed by simply shaking in the detection medium until obtaining a flat voltammetric response before applying a new sequence of accumulation-detection. To avoid high contamination of the detection medium with Pb²⁺ and Cu²⁺ ions due to desorption, the detection solution was constantly renewed after several accumulation-detection series. The electrochemical characterisation of the natural kaolinite or organokaolinite film deposited on the surface of GCE was carried out by studying the electrochemical behaviour of [Fe(CN)₆]³⁻ ions. Thus, the working electrode (clay thin film electrode) was immersed in an aqueous solution of [Fe(CN)₆]³⁻ ions concentrated 10⁻³ mol L⁻¹ at a given pH (KCl support electrolyte concentrated 0.1 mol L⁻¹) and then scanned by multisweep Cyclic Voltammetry (CV) between 0.5 and − 0.3 V until a stable signal was obtained, corresponding to saturation of the film. Comparison of the evolution of the multicyclic voltammetry scan signals and the peak currents at saturation recorded on different electrodes (GCE, GCE/K and GCE/KHEP) allowed us to deduce information about the nature of the physicochemical properties of the surface of organokaolinite and pure kaolinite.

2.4.3 Procedure of electrochemical simultaneous detection of pb(II) and Cu(II) in natural water

For the application of the electrochemical sensor implemented in this work in a real environment, three types of natural water were chosen. Spring water and well water, both collected in the city of Yaoundé (Cameroon) in the Nkolbisson quarter and a tap water that of the Cameroonian company CAMWATER in charge of the production and distribution of drinking water. These water samples collected in real environments were then considered and referred to as the natural reference water samples. In practice, the natural reference water sample collected (natural reference solution) is used here as the accumulation medium. Firstly, the GCE/KHEP working electrode was immersed in 50 mL of natural reference water, stirred at a constant speed. After 10 min of accumulation, the electrode was removed and rinsed, then immersed in the detection medium where the electrochemical signal is recorded. Secondly, known quantities of Pb²⁺ ions and Cu²⁺ ions were successively added to the previous reference natural water sample and the electrochemical response corresponding to each concentration of added analytes was recorded. The peak current intensities of Pb(II) and Cu(II) measured were then used to plot the calibration curves corresponding to each type of natural water. In a third step, and finally, known concentrations of Pb²⁺ ions, C₁, and Cu²⁺ ions, C₂, were added to another sample of natural water and the electrochemical signal, of simultaneous detection, was recorded and the measured peak current intensities were projected onto the calibration curves obtained for the reference solutions and the equivalent concentrations were determined. The concentrations found in the test solution from the calibration curves, denoted C₁’ and C₂’ respectively were then compared with the introduced concentrations C₁ and C₂ and a percentage of recovery of the quantities determined by the electrochemical method used is then estimated.

3 Results and discussion

3.1 Structural and physicochemical characterisation of pure kaolinite and organokaolinite

3.1.1 XRD structural characterisation

Figure 1 shows the XRD patterns of pure Kaolinite, K-DMSO and KHEP materials, recorded for 2θ ranging from 2° to 20°. In the scanned 2θ range, the XRD diffractogram obtained for pure kaolinite (Fig. 1a) shows an intense and well-defined peak for a value of 2θ equal to 12.3°, which corresponds to an interlayer distance d₀₀₁ = 7.1 Å. This result is in perfect agreement with those obtained in previous work and is characteristic of kaolinite clay mineral [39, 41, 43,44,45,46,47,48,49,50,51,52]. On the XRD diffractogram of K-DMSO material (Fig. 1b), a shift of the intense peak corresponding to the d₀₀₁ line towards a smaller value of the angle 2θ was observed, compared with that of pure kaolinite and corresponding to a d₀₀₁ = 11.1 Å, i.e. an increase of 4.0 Å in the inter-reticular distance attributable to the intercalation of DMSO molecules in the interlayer space of pure kaolinite [39, 41, 46, 47, 50, 51]. The XRD pattern of K-DMSO also showed a very weak peak corresponding to the d₀₀₁ = 7.1 Å basal spacing of pure kaolinite, indicating the presence of a very small quantity of pure kaolinite that was not modified during DMSO intercalation. The XRD pattern of organokaolinite (KHEP) after two washings is presented in Fig. 1c. This diffractogram showed two sharp reflection peaks. The first peak is the characteristic d₀₀₁ of kaolinite (7.1 Å), and the second peak (d = 11.2 Å) is attributed to the intercalation of HEP in the interlayer space of pure kaolinite. With respect to the d₀₀₁ of starting kaolinite, the previous d₀₀₁ value of 11.2 Å for an organokaolinite KHEP indicate to an expansion of 4.1 Å. This result demonstrates and confirms the intercalation and grafting of HEP molecule within the interlayer space of the natural clay. This expansion of the d₀₀₁ ray of organokaolinite by a value of 4.1 Å compared with that of the initial pure kaolinite, corresponds approximately to a monolayer of HEP molecules arrangement, almost vertically (chair conformation) in the interlayer space of kaolinite (Scheme 1) [39].

Fig. 1

Powder X-ray diffraction patterns of a Pure Kaolinite, b K-DMSO and c KHEP materials

3.1.2 FTIR spectroscopy characterisation

Fourier Transform Infrared (FTIR) spectroscopy is a technique for characterising compounds and materials, which allow the functional groups, present to be identified. FTIR spectra of K, K-DMSO and organokaolinite KHEP are presented in Fig. 2. The wavelength range from 4000 to 3600 cm⁻¹ on the FTIR spectra of pure kaolinite, K-DMSO and KHEP corresponds to the –OH stretching region. The absorption bands of pure kaolinite (Fig. 2a) at 3692, 3668 and 3651 cm⁻¹ correspond to interlayer hydroxyl stretching while the band observed at 3620 cm⁻¹ corresponds to the stretching vibrations of the internal hydroxyl group [41, 45,46,47, 50, 51]. In the FTIR spectrum of K-DMSO (Fig. 2b), the disappearance of the absorption bands at 3668 and 3651 cm⁻¹ is observed as well as the appearance of new bands at 3535, 3500 and 3664 cm⁻¹ [39, 41, 45,46,47, 50]. The New bands which appear at 3535 and 3500 cm⁻¹ in the FTIR spectrum of K-DMSO corresponds to hydrogen bonding interactions between the hydroxyl groups of the inner surface and the clay mineral and interlayer DMSO [44,45,46,47, 50, 51]. The stretching bands at 3020 and 2933 cm⁻¹ were also observed in the FTIR spectrum of K-DMSO which are attributed to –CH₃ groups of DMSO molecules [39, 41,42,43, 45,46,47,48,49,50,51]. In the case of KHEP (Fig. 2c), two bands were observed at 2825 and 2954 cm⁻¹ corresponding to the symmetric and antisymmetric stretching modes of –CH₂– groups in HEP, indicating it is present in the grafted material [37, 39, 51]. The characteristic band of the CN stretching vibration of HEP is present at 1470 cm⁻¹ [42, 53]. The absorption bands centred at 3439 and 1645 cm⁻¹ observed in the FTIR spectrum of KHEP are attributable to the bending vibrations of adsorbed water molecules within the organokaolinite [39]. Also noteworthy is the total disappearance of the characteristic DMSO absorption bands (3535 and 3500 cm⁻¹) in the FTIR spectrum of organokaolinite KHEP, indicating that during the synthesis, all the DMSO molecules previously intercalated in the interlayer of pure kaolinite were completely replaced or substituted by HEP molecules.

Fig. 2

FTIR spectra (4000–700 cm⁻¹) of a pure Kaolinite (K), b DMSO intercalated pure Kaolinite (K-DMSO) and c Kaolinite modified by grafting of 1-(2-Hydroxyethyl)piperazine (KHEP)

3.1.3 ²⁹Si CP/MAS NMR spectroscopic characterisation

The ²⁹Si MAS NMR spectra of pure kaolinite, kaolinite modified by intercalation of DMSO and organokaolinite are shown in Fig. 3. The ²⁹Si MAS NMR spectrum of pure kaolinite showed two signals of equal intensities well separated at − 91.5 and − 90.9 ppm which can be attributed to crystallographically equivalent Si atoms of the tetrahedral sheet (Fig. 3a) [44, 54, 55]. After the intercalation of DMSO, the signal is shifted to lower frequencies, at − 92.7 ppm, accounting for the keying of one of the DMSO methyl groups in the siloxane rings of the tetrahedral sheets [53, 56]. This slight shift in the signal from − 91.5 for pure kaolinite to the low frequencies of − 92.7 ppm for KDMSO illustrates the stability of the silica sheet of kaolinite material [57]. The two low-intensity peaks at − 91.5 and − 90.9 ppm observed on the ²⁹Si MAS NMR spectra of K-DMSO correspond to the natural kaolinite residue that did not undergo any modification during the intercalation of the DMSO molecules. When the DMSO molecules are displaced by the intercalation and grafting of the 1-(2-hydroxyethyl)piperazine molecules, we notice that the ²⁹Si NMR spectra obtained (Fig. 3b) show a well-defined signal at − 91.4 ppm which corresponds to the Si atoms of the tetrahedral sheet having formed the Si–O bond during grafting and another relatively broad signal envelope around − 90.9 ppm and of low intensity which certainly corresponds to the Si atoms of the tetrahedral sheet which establish hydrogen bonds with the amine group of the grafted 1-(2-hydroxyethyl)piperazine molecule. The total disappearance of the signal at − 92.7 ppm on the ²⁹Si NMR spectra of the KHEP material (Fig. 3c) clearly confirms that all the DMSO molecules were replaced during the synthesis of the organokaolinite.

Fig. 3

²⁹Si MAS NMR spectra of a pure Kaolinite (K), b Kaolinite intercalated with DMSO (K-DMSO) and c Kaolinite grafted with 1-(2-Hydroxyethylpiperazine) (KHEP)

3.2 Electrochemical characterisation of pure kaolinite and organokaolinite

Several previous works have shown that the study of certain reference redox systems such as [Fe(CN)₆]³⁻ and [Ru(NH₃)₆]³⁺ by the electrochemical method using cyclic voltammetry and multicyclic scanning can be used to evaluate the ion exchange properties of a thin film layer of a material deposited on the surface of a conventional GCE [37, 45, 57,58,59,60]. In practice, examination of the evolution of the intensities of the electrochemical responses obtained by multicyclic scanning provides an indication of those interactions between the surface of the modified electrode and the redox system under study that are favourable or unfavourable. Furthermore, by examining the differences between the intensities of the currents of successive peaks in cyclic voltammograms, it is possible to deduce useful information about the diffusion kinetics of the analytes in the film deposited on the GCE. As part of this work, in order to evaluate the physicochemical surface properties and ion exchange properties of the pure kaolinite K and KHEP organokaolinite film, the electrochemical behaviour of [Fe(CN)₆]³⁻ ions was studied at the surface of a conventional glassy carbon electrode modified by the film of this materials and using the cyclic voltammetry technique. Figure 4 shows the cyclic voltammetry signals at [Fe(CN)₆]³⁻ ions saturation (10⁻³ mol L⁻¹) recorded at a scan rate of 50 mV s⁻¹ using the bare GCE unmodified, GCE coated with pure kaolinite thin film (GCE/K) and GCE coated with organokaolinite thin film (GCE/KHEP) in analysis medium pH 7.0 (Fig. 4a) and in analysis medium pH 2.0 (Fig. 4b). These results show that for all three types of electrode, the electrochemical system is reversible, corresponding to the monoelectronic transformation of the [Fe(CN)₆]³⁻/[Fe(CN)₆]⁴⁻ system. Whatever the pH value of the analysis medium, the intensity of the electrochemical signal recorded on bare GCE remains unchanged despite several cyclic voltammetry scans, indicating a constant concentration of [Fe(CN)₆]³⁻ ions at the surface of this electrode. However, it can be noted that the peak current of the saturation signal obtained on unmodified GCE at pH 2.0 is 1.5 times more intense than the signal recorded under the same conditions on bare GCE at pH 7.0, indicating a higher concentration of [Fe(CN)₆]³⁻ ions at the GCE electrode at this pH value. This could be explained by significant protonation of the carboxylate groups on the GCE surface, resulting in more favourable interactions with [Fe(CN)₆]³⁻ ions, an increase in its concentration at the surface of electrode and consequently a more intense peak current. The electrochemical response recorded by cyclic voltammetry up to saturation on GCE/K in the analysis medium at pH 7.0 is 2.1 times less intense than the signal obtained on bare GCE under the same conditions. This indicates that the pure kaolinite film allows very little adsorption and diffusion of ferricyanide ions until to the reactive surface of the GCE. This result indicate that the pure kaolinite film behaves somewhat like a barrier for [Fe(CN)₆]³⁻ ions at pH of analysis medium equal 7.0 [37, 58,59,60]. The peak current at saturation (Ip_s), for [Fe(CN)₆]³⁻ ions, recorded on GCE/K at pH 2.0 is twice as intense as that obtained at pH 7.0 on the same electrode. This could also be explained by a light protonation of the –OH groups and oxygen groups on the external surface and internal surface of pure kaolinite at pH 2.0, which would permit a little more favourable interactions with [Fe(CN)₆]³⁻ ions. It is in the case of cyclic voltammetry signals recorded on GCE/KHEP in the analysis medium pH 7 and in the analysis medium pH 2.0 that the difference in saturation peak current intensity is most spectacular. The saturation peak current of the cyclic voltammetry signal of [Fe(CN)₆]³⁻ ions obtained in the pH 2.0 medium is 8.1 times more intense in oxidation reaction and 10.2 times more intense in reduction reaction than the saturation signal intensity recorded on the GCE/KHEP electrode at pH 7.0, indicating a high concentration of [Fe(CN)₆]³⁻ ions adsorbed on the surface of the GCE/KHEP electrode in the analysis medium pH 2.0. This could certainly be explained by strong protonation at pH 2.0 of the amine groups (–NH₂⁺–) of the 1-(2-hydroxyethyl)piperazine molecules grafted into the organokaolinite, which would then lead to a strong accumulation of [Fe(CN)₆]³⁻ ions by very favourable electrostatic interactions. Multicyclic scanning voltammetry signals recorded under the same experimental conditions and on the GCE/KHEP electrode for different acidic pH values of the analysis medium and presented in Fig. 5a (pH 4.0); Fig. 5b (pH 3.0) and Fig. 5c (pH 2.0) show that the accumulation of [Fe(CN)₆]³⁻ ions in the KHEP organokaolinite film occurs progressively until saturation of the film is obtained, resulting in a constant signal intensity. This progressive accumulation of [Fe(CN)₆]³⁻ ions could be explained by a progressive protonation of the amine groups in the film of the KHEP material (–NH₂⁺–), itself due to a progressive diffusion and fixation of H⁺ ions in this film. In the particular case of the multicyclic voltammetry signals recorded in analysis medium pH 2.0, it can be seen that the progressive increase peak current intensity is accompanied by a gradual shift in peak potential towards more positive values, which could be explained by a difference in kinetics between the protonation process of the amine groups of the KHEP material and the accumulation process of [Fe(CN)₆]³⁻ ions by favourable electrostatic interactions. Analysis of the saturation VC signals recorded under the same conditions in the different media pH 7.0, 4.0, 3.0 and 2.0 and presented in Fig. 5d shows that the intensity of the saturation peak current (Ip_s) increases with decreasing pH, i.e. with increasing acidity of the medium. This certainly indicates that the more acidic the analysis medium, the greater the protonation of the amine groups in the film of the KHEP material, and the greater the quantity of ferricyanide ions accumulated [37, 58, 59]. Analysis of the signals presented in Fig. 5d also shows a shift in peak potential values towards the more positive values with decreasing pH, which is in agreement with the results obtained in previous work [37, 59]. To yield more information on the electrochemical behaviour of [Fe(CN)₆]³⁻ ions at the GCE/KHEP electrode, the effect of increasing the potential scan rate (V_b) on the cyclic voltammetric peak current was also investigated. The cyclic voltammetric responses of [Fe(CN)₆]³⁻ ions obtained under the same conditions as in Fig. 5 are presented in Fig. SI 01. Examination of this result showed that, the peak current intensity (Ip) increased with V_b in the range from 25 to 200 mV s⁻¹ and a plot of oxidation peak current Ip_ox versus the square root of V_b (Fig. SI01b_(i)) and a plot of reduction peak current Ip_red versus the square root of V_b (Fig. SI01b_(ii) ) exhibited a linear dependence, Ip_ox = 1.401 V_b^1/2 + 3.061 (R² = 0.993) in the case of oxidation peak current and Ip_red = − 1.478 V_b^1/2 − 4.688 (R² = 0.990) in the case of reduction peak current. This finding reveals that the charge transfer process at the organokaolinite modified electrode is controlled by diffusion [59, 60]. The same study of the influence of the scanning speed on the saturation peak current of [Fe(CN)₆]³⁻ ions was carried out for a pH of the analysis medium equal to 3.0. As in the previous case, the results of the voltammograms obtained and presented in Fig. SI 02 also showed that, the peak current intensity (Ip) increased with V_b in the range from 25 to 200 mV s⁻¹ and a plot of oxidation peak current Ip_ox versus the square root of V_b (Fig. SI02b_(i)) and reduction peak current Ip_red versus the square root of V_b) (Fig. S02b_(ii)) exhibited a linear dependence, Ip_ox = 0.938 V_b^1/2 + 7.570 (R² = 0.994) in the case of oxidation peak current and Ip_red = − 1.322 V_b^1/2 − 8.163 (R² = 0.992) in the case of reduction peak current. This linear relation between Ip and square root of scans rate V_b^1/2, reveals that even at pH of analysis medium equal 3.0, the charge transfer process at the organokaolinite modified electrode is controlled by diffusion [59, 60].

Fig. 4

Superposition of steady-state cyclic voltammetry curves recorded at 50 mV s⁻¹ in 0.1 mol L⁻¹ KCl + 1.0 mmol L⁻¹ [Fe(CN)₆]³⁻ between + 0.65 V and – 0.30 V potential range, using the bare GCE ; GCE coated with pure kaolinite (GCE/K) and GCE coated with KHEP organokaolinite at a pH 7.0 and b pH 2.0

Fig. 5

Multisweep cyclic voltammograms recorded at 50 mV s⁻¹ in 0.1 M KCl + 1.0 mmol L⁻¹ [Fe(CN)₆]³⁻ between + 0.65 V and – 0.30 V potential range, using the bare GCE coated with KHEP organokaolinite at a pH 4.0; b pH 3.0 and c pH 2.0. Superposition of steady-state Cyclic Voltammograms curves of [Fe(CN)₆]³⁻ ions recorded on the GCE/KHEP with different pH of detection medium (i) 7.0 ; (ii) 4.0 ; (iii) 3.0 and (iv) 2.0

3.3 Electrochemical application of the organokaolinite to the simultaneous detection of lead(II) and copper(II)

3.3.1 Preliminary studies

In order to consider the application of the sensor based organokaolinite GCE/KHEP, for the simultaneous detection of Pb(II) and Cu(II) in the same medium, it is important first to carry out some preliminary studies that we consider fundamental. These preliminary studies include assessing the ability of the GCE/KHEP sensor to give an electrochemical response with well-defined peak currents that are distinct from each other and correspond to each of the heavy metal ions analysed. A set of preliminary investigations were performed to define basic conditions for the electrochemical simultaneous detection of Pb²⁺ and Cu²⁺. Figure 6 presents the Anodic Stripping Square Wave Voltammograms (ASSWV) recorded on GCE/KHEP from an aqueous solutions containing Pb²⁺ only concentred 2.0 µmol L⁻¹ (curve a in Fig. 6a), an aqueous solutions containing Cu²⁺ only concentrated 2.0 µmol L⁻¹ (curve a in Fig. 6b), an aqueous solution containing a mixture of Pb²⁺ 2.0 µmol L⁻¹ and Cu²⁺ ions concentrated 2.0 µmol L⁻¹ (curve a in Fig. 6c) after 7 min accumulation time (pH 5.0) for each case. All signals were recorded under the same conditions in detection solution of NaNO₃ concentrated 0.2 mol L⁻¹ with pH adjusted at 2.5. In individual detection, the voltammetric response shows for Pb(II) (curve a in Fig, 6a), an intense and well defined peak current centred − 0.55 V and for Cu(II) (curve a in Fig. 6b), a well-defined peak current centred 0.15 V. The simultaneous detection signal (curve a in Fig. 6c) gives two well-defined peak currents centred at -0.55 V and 0.15 V which certainly correspond to the SWV peak currents of Pb(II) and Cu(II) respectively. These results presented on Fig. 6c(i) shows that the GCE/KHEP sensor can be successfully applied to the simultaneous detection of Pb(II) and Cu(II) heavy metals ions in a given contaminated aqueous solution. The peak current intensity comparison shows that the GCE/KHEP sensor under the experimental conditions in Fig. 6 is more sensitive to the electrochemical detection of Pb(II) than electrochemical detection of Cu(II). Under the same experimental conditions, we note that in electrochemical simultaneous detection, the peak current of Pb(II) decreases by 30% and the peak current of Cu(II) decreases by 20% compared with their peak currents obtained in individual detection. This could be explained by competition between the two ions for binding sites of KHEP film during simultaneous detection [38]. The signals in Fig. 6j, k, l, m and n correspond to the desorption of the analytes Pb(II) and Cu(II) after detection. Analysis of these signals shows that the Pb²⁺ ions, whatever the type of detection, are completely and efficiently desorbed from the electrode surface after one a simple shaking of the detection medium resulting in a Pb(II) peak current of zero intensity (Curve b in Fig. 6a and Curve b in Fig. 6c). On the other hand, for Cu(II), we observe a peak current that decreases progressively following multiple shaking of the detection medium until a flat signal is obtained after five successive shakings. This indicates a more difficult desorption of Cu²⁺ ions from the KHEP film, which takes place slowly and progressively in the detection medium. The level of stability of a signal given by an electrochemical sensor strongly determines its application for the electroanalysis of chemical species in a real environment. It provides a means of assessing the reliability of the electrochemical signals given by a sensor and also of demonstrating the performance of the sensor implemented. The stability of a result obtained under the same experimental conditions is a fundamental performance criterion in the field of analytical chemistry. The stability of signal consequently the stability of the sensor based on thin film organokaolinite GCE/KHEP was studied. Thus, a series of ten electrochemical responses for the simultaneous detection of Pb²⁺ and Cu²⁺ ions were recorded by ASSWV on the surface of the same GCE/KHEP electrode under the same experimental accumulation and detection conditions and at regular 30 min time intervals (results not shown). No significant decrease in the electrochemical response between these ten successive measurements was observed. For this series of ten measurements, the relative standard deviation (RSD) of the anodic ASSWV peak current was found to be 1.5% for Pb(II) and 2.5% for Cu(II), this gives a stability rate of 98.5% and 97.5% of the electrochemical response of the Pb²⁺ and Cu²⁺ ions respectively recorded on the GCE/KHEP sensor. These results indicate that thin film of KHEP deposited on the GCE are also quite stable and can be used for several successive measurements [37, 39, 60]. The stability of the signal for simultaneous detection of Pb²⁺ and Cu²⁺ ions on the surface of GCE/KHEP under the same conditions over the long term was assessed. For the same electrode and under the same experimental accumulation and detection conditions, the square wave voltammetry signal for simultaneous detection of Pb(II) and Cu(II) was recorded every day at the same time for ten days. During these experiments, the GCE/KHEP electrode was stored in a 0.1 mol L⁻¹ KCl (pH 7.0) when not in use. The relative standard deviation (RSD) for this series of ten ASSWV signals (results not show) was less than 5% (3.5%), this gives a stability rate of 96.5%. These results demonstrate that the stability of square wave voltammetric response obtained of the GCE/KHEP sensor is good and the KHEP thin film on the surface of glassy carbon electrode is also very stable at the long term [27, 37, 39].

Fig. 6

Anodic stripping square wave voltammograms recorded on GCE/KHEP in 0.2 mol L⁻¹ NaNO₃ detection medium pH 2.5 after 7 min accumulation from a an aqueous 2.0 µmol L⁻¹ Pb²⁺ solution at pH 5.0; b an aqueous 2.0 µmol L⁻¹ Cu²⁺ solution at pH 5.0 c an aqueous a mixture of 2.0 µmol L⁻¹ Pb²⁺ and 2.0 µmol L⁻¹ Cu²⁺ solution at pH 5.0 (i) ASSWV response obtained directly after accumulation step, (j) SWV response recorded after first desorption in detection medium (k) second (l) third (m) fourth and (n) fifth desorption. Other experimental conditions are as follows: amplitude: 0.05 V, step potential (V): 0.005 V, frequency: 50 Hz, t_elec: 80 s and E_elec: − 0.90 V

3.3.2 Electrochemical response of simultaneous detection of Pb²⁺ and Cu² on GCE modified by the pure kaolinite and organokaolinite.

The main objective of this work was to modify the surface of a conventional glassy carbon electrode with an organokaolinite film and to apply the resulting sensor to the sensitive and efficient simultaneous electrochemical detection of Pb²⁺ and Cu²⁺ in a contaminated environment. To demonstrate the value of chemically modifying kaolinite prior to its application for the implementation of an electrochemical sensor, we have decided to carry out a comparative study of the simultaneous detection ASSWV signal on GCE modified with a pure kaolinite film and the signal obtained on GCE modified with an organokaolinite film. Figure 7 shows typical ASSWV curves of simultaneous detection of Pb²⁺ and Cu²⁺ ions recorded on bare GCE (Fig. 7a), GCE modified by thin film of pure kaolinite (GCE/K) (Fig. 7b) and signal recorded on GCE modified by thin film of organokaolinite (GCE/KHEP) (Fig. 7c) in NaNO₃ medium (0.2 mol L⁻¹ pH adjusted at 2.5), after 7 min accumulation time at open circuit in aqueous medium containing each ions at 2.0 µmol L⁻¹. The results obtained and presented in Fig. 7 show that the peak current intensities depend strongly on the nature of the material film deposited on the GCE surface. The peak current intensities of Pb²⁺ and Cu²⁺ on the simultaneous detection signal are 50.5 µA and 40.7 µA on GCE/K and 207.5 µA and 163.4 µA on GCE/KHEP, respectively. A comparative analysis of the ASSWV electrochemical responses shows that, the Pb(II) peak current obtained on GCE/KHEP is more than 4.1-fold more intense than the Pb(II) peak current obtained on GCE modified with pure kaolinite (GCE/K). The Cu(II) peak current obtained on GCE/KHEP is 4.0 times more intense than that obtained on GCE modified with pure kaolinite (GCE/K). This observation clearly shows that, the organokaolinite film accumulates much more Pb²⁺ and Cu²⁺ ions than the pure kaolinite film (K). Indeed, on the pure kaolinite film, adsorption of the heavy metal ions analysed occurs only by interactions with the surface hydroxyl groups, whereas on the organokaolinite film adsorption occurs not only by interactions with the surface hydroxyl groups of the kaolinite structure but also by complexation reaction with the amine groups of the HEP molecules which are intercalated and grafted in the interlayer space of organokaolinite. There are several similar previous works in the literature that sufficiently demonstrate the high affinity of Pb²⁺ and Cu²⁺ ions with the amine groups with which they form coordination or complexation bonds [30, 37,38,39, 41, 49].

Fig. 7

ASSWV signal corresponding to the simultaneous detection of the Pb²⁺ and Cu²⁺ ions recorded on a Bare GCE; b GCE/K ; c GCE/KHEP in 0.2 mol L⁻¹ NaNO₃ medium (detection medium pH 2.5) after 7 min of accumulation in deionized aqueous medium (pH 5.0) containing each ion at a concentration of 2.0 µmol L⁻¹. Other experimental conditions are as follows: amplitude: 0.05 V, step potential (V): 0.005 V, frequency: 50 Hz, t_elec: 80 s and E_elec: − 0.90 V

3.3.3 Optimization of parameters affecting the detection step

The procedure for the simultaneous electroanalysis of Pb(II) and Cu(II) involves two main steps [37,38,39, 48]. An open-circuit accumulation step and detection step in a suitable medium. After adsorption of the Pb²⁺ and Cu²⁺ ions onto the film of the KHEP material during the accumulation step, they must first be desorbed from the various binding sites in the detection medium and then react with the surface of the GCE electrode, during which the corresponding electrochemical response is recorded by ASSWV. The influence of the nature of the detection medium and its pH was evaluated. The curves giving the evolution of the peak currents of Pb²⁺ ions and Cu²⁺ ions as a function of the pH of the detection medium are shown in Fig. SI 03. In the case of very acidic detection media with a pH of less than 2.5, a poor stability of the electrochemical response was observed, linked to very strong protonation of the amine groups in the KHEP material film during the desorption step, making it very difficult to regenerate the material film electrode for a new accumulation and detection. For detection medium pH values above 2.5, signals with low the intensity peak currents of electrochemical signal obtained decrease. This could be explained by least reduction by ion exchange of Pb²⁺ and Cu²⁺ ions in this analysis medium, which is less rich in H⁺ ions. Remember that, Anodic Stripping Square Wave Voltammetry (ASSWV) it is method was used to the simultaneous electrochemical detection of Pb(II) and Cu(II). In fact, this electrochemical method of analysis includes a sub-step of pre-concentration of the analytes by electrolysis (reduction) using a suitable electrolysis potential (E_elec) imposed on the working electrode for a given time (T_elec) to reduce the maximum amount of heavy metal ions accumulated previously. In order to obtain intense and stable electrochemical signals under the same experimental conditions, 0.2 mol L⁻¹ NaNO₃ pH adjusted at 2.5 was selected as the optimal pH of the detection medium for subsequent experiments [39]. Table 1 summarises the optimised of parameters affecting detection step for the simultaneous electrochemical detection of lead and copper ions that were used in this work. Under the experimental conditions shown in Fig. 7, the electrolysis time was studied between 0 and 150 s and the electrolysis potential was studied between 0 and − 1.1 V. The curves obtained are included as supplementary information for two values of pH of accumulation solution (Fig. SI 04 for accumulation solution pH 5.0 and Fig. SI 05 for accumulation solution pH 6.0). The results in Table 1 show that − 0.9 V and 80 s are respectively the optimal electrolysis potential and the optimal electrolysis time that were used in further experiments. It is important to note that other parameters linked to the electrochemical method used (ASSWV) and which affect the detection step were selected according to a previous study by our group [39]. Their values ​​were kept constant. These include potential amplitude 0.05 V, step potential 0.005 V and frequency 50 Hz.

Table 1 Some optimum parameters for the simultaneous voltammetric detection of lead and copper ions concentration, at each of them it is 2.0 µmol L^-1, using glassy carbon electrode (GCE) modified by a thin film of organoclay material GCE/KHEP by Anodic stripping square wave voltammetry (ASSWV) method

3.3.4 Optimization of parameters affecting the accumulation step

Several parameters govern the quantity of Pb²⁺ and Cu²⁺ ions accumulated on the surface of the GCE/KHEP working electrode during the accumulation step. Table 1 also summarises the optimised parameters that affecting the accumulation step for the simultaneous detection of lead and copper ions. The influence of amount KHEP organokaolinite in the dispersion used to prepare thin film electrode, was studied between 2.0 and 10.0 g L⁻¹. Analysis of the curve obtained from this study (Fig. SI 06) shows that the intensity of the electrochemical signal increases with the concentration of the organoclay suspension, reaching a maximum at 4.0 g L⁻¹and then starting to decrease. The greater the quantity of KHEP material at the electrode, the greater the number of binding sites and the greater the quantity of accumulated Pb²⁺ and Cu²⁺ ions, resulting in a more intense signal [32, 37, 60]. For concentrations in suspension greater than 4.0 g L⁻¹, the film becomes increasingly dense, making diffusion of Pb²⁺ and Cu²⁺ ions more difficult, which then leads to a decrease in the quantity of accumulated ions, hence the decrease in peak current observed. 4.0 g L⁻¹ was chosen as the optimal amount of organokaolinite in a suspension for further experiments. Considering that the different binding sites for Pb²⁺ and Cu²⁺ ions can also bind protons, thereby creating a competitive situation, we decided, using the preselected optimal parameters (E_elec = − 0.9 V, t_elec = 80 s), to study the influence of the pH of the accumulation medium on the electrochemical response of simultaneous detection. Figure 8 shows the square wave voltammograms of simultaneous detection of Pb²⁺ and Cu²⁺ recorded under the same conditions after 5 min of accumulation time in an accumulation medium with each ion concentrated to 2.0 × 10⁻⁶ mol L⁻¹ for the pH equal 4.0 ; 5.0 and 6.0. Analysis of these three signals shows remarkable differences in peak current. At pH 4.0, a very low Pb(II) peak current (13.6 µA) and a very high Cu(II) peak current (223.5 µA) are observed, indicating a certain good selectivity of the GCE/KHEP sensor to almost exclusively bind Cu²⁺ ions for this pH value of the accumulation medium. At pH 5.0, the electrochemical response obtained shows two intense and well-defined peak currents corresponding to the two ions detected, i.e. 205.6 µA for Pb(II) and 175.5 µA for Cu(II), which suggests that for this pH value of the accumulation medium, the sensor adsorbs quantitatively and simultaneously both Pb²⁺ and Cu²⁺ ions. It was concluded that at pH of the accumulation medium equal to 5.0, the GCE/KHEP sensor has no particular affinity for either of the two ions analysed. At pH 6.0, the electrochemical response obtained showed two signals, one attributable to Pb(II) and the other to Cu(II). However, examination of the signal showed that the peak current for Pb²⁺ ions (275.7 µA) is 3.7-fold more intense than that for Cu²⁺ ions (73.5 µA), demonstrating that for this pH accumulation value, the GCE/KHEP sensor has a greater affinity to adsorb Pb²⁺ ions. The GCE/KHEP sensor therefore appears to be more selective for the detection of Pb(II) at this pH of the accumulation medium. Figure 9 shows the evolution of the peak current of Pb(II) and Cu(II), obtained from simultaneous detection after 5 min of accumulation in 2.0 × 10⁻⁶ mol L⁻¹ medium, as a function of the pH of the accumulation medium for a pH range from 2.0 to 10.0. The analysis of this curve shows that for low pH values of the accumulation medium below 3.0, the peak currents of Cu²⁺ and Pb²⁺ ions are of low intensity, indicating low accumulation of these heavy metal ions on the surface of the GCE/KHEP electrode in this pH range, certainly due to strong protonation of the amine groups (–NH₂⁺–) involving electrostatic repulsion of the Cu²⁺ and Pb²⁺ ions with the electrode surface [33, 37,38,39, 48]. From pH 4.0 onwards for Cu(II) and from pH 5.0 onwards for Pb(II), there is a very sharp and significant increase in the intensity of the peak current, indicating a strong accumulation of heavy metal ions on the various binding sites, in this case on the amine groups, which are then more available in the non-protonated form (–NH₂⁺–). Above pH 4.0, the Cu(II) peak current decreases, whereas the lead peak continues to increase in intensity, reaching its maximum at pH 6.0. This could be explained by the onset of precipitation of Cu²⁺ ions into Cu(OH)₂ reducing their ability to complex with amine groups. Between pH 4.0 and 5.0 for copper and between pH 6.0 and 7.0 for lead, a smaller variation in peak current intensities is observed, indicating that the accumulation process of Pb²⁺ and Cu²⁺ ions is strongly controlled by a complexation mechanism with amine groups [37, 39, 48]. Between pH 5.0 and pH 6.0, the peak current of Cu(II) decreased sharply and then gradually decreased until pH 9.0. In contrast, the evolution of the Pb(II) peak current did not show a sudden decrease but rather a gradual decrease up to pH 9.0, probably due to the gradual precipitation of Pb²⁺ ions to Pb(OH)₂ in the accumulation medium. The pH 5.0 accumulation medium was chosen for the efficient and sensitive simultaneous detection of Cu²⁺ and Pb²⁺ ions in the other experiments. The influence of the accumulation time on the electrochemical response was studied between 0 and 16 min for the medium containing Pb²⁺ and Cu²⁺ ions at a concentration of 2.0 × 10⁻⁶ mol L⁻¹ and between 0 and 24 min for the more dilute media containing Pb²⁺ and Cu²⁺ ions at a concentration of 2.0 × 10⁻⁷ mol L⁻¹. The voltammograms obtained from the case where the concentration of analytes in accumulation medium is 2.0 × 10⁻⁶ mol L⁻¹ are shown in Fig. SI07. The plot of peak current intensity as a function of accumulation time (Inset in Fig. SI07) showed that the peak current increase with the accumulation time as observed in previous works [27, 28, 37, 38, 40, 42, 48, 60]. For the least dilute medium (2.0 × 10⁻⁶ mol L⁻¹), the film of KHEP material starts to saturate in Pb²⁺ and Cu²⁺ ions after 10 min of accumulation time. For the more dilute medium containing 4.0 × 10⁻⁷ mol L⁻¹ and 2.0 × 10⁻⁷ mol L⁻¹ (Fig. SI 08 and Fig. SI09, respectively in Supplementary Information), longer accumulation time 16 min is required to obtain a stable signal, indicating film saturation [37, 38, 42, 48]. An accumulation time of 16 min was chosen as the accumulation time for any further accumulation and electrochemical detection experiments in accumulation media more diluted than 2.0 × 10⁻⁶ mol L⁻¹.

Fig. 8

ASSWV recorded for simultaneous detection of Pb²⁺ and Cu²⁺ in 0.2 mol L⁻¹ NaNO₃ (pH 2.5) at GCE/KHEP electrode after 7 min of accumulation time in deionized water different pH 6.0; 5.0 and 4.0 with each ion at the concentration of 2.0 µmol L⁻¹. Other experimental conditions are as follows: amplitude: 0.05 V, step potential (V): 0.005 V, frequency: 50 Hz, t_elec: 80 s and E_elec: − 0.90 V

Fig. 9

Effect of the pH of the accumulation medium on the ASSWV peak current of 2.0 µmol L⁻¹ of the Pb²⁺ and Cu²⁺ ions recorded on GCE/KHEP in 0.2 mol L⁻¹ NaNO₃ medium (pH 2.5) after 7 min of accumulation in deionized water. Other experimental conditions are as follows: amplitude: 0.05 V, step potential (V): 0.005 V, frequency: 50 Hz, t_elec: 80 s and E_elec: − 0.90 V

3.3.5 Influence of Pb²⁺ and Cu²⁺ concentration on the electrochemical response of simultaneous detection and calibration graph.

The electrochemical theory show that, the intensity of the voltammetry peak current of chemical electroactive specie, is proportional to the amount of this specie adsorbed on the surface of working electrode during the accumulation step. Thus, using the optimised experimental conditions of accumulation step and detection step obtained from previous section (Table 1), the influence of Pb(II) and Cu(II) concentration on the ASSWV response of electrochemical simultaneous detection of Pb²⁺ and Cu²⁺ was studied for GCE/KHEP for different pH of accumulation medium. Figure 10a show the typical ASSWV of simultaneous detection of Pb²⁺ and Cu²⁺ signal recorded in 0.2 mol L⁻¹ NaNO₃ (detection medium pH adjusted at 2.5) at GCE/KHEP electrode after 10 min of accumulation time in deionized water (pH of accumulation medium was adjusted at 5.0) for a concentration range from 0.2 to 1.0 µmol L⁻¹ for each of the two ions analysed. The analysis of the voltammograms obtained shows that the peak current intensity increases with Pb²⁺and Cu²⁺concentration in the accumulation medium. As shown by Fig. 10b, a linear relationship was obtained for the peak current (Ip) and concentration of Pb²⁺ and concentration of Cu²⁺, which obeyed to the two following equations: Ip(µA) = 96.502 [Pb²⁺] (µmol L⁻¹) − 0.422 (R² = 0.993) for Pb(II) and Ip(µA) = 66.255 [Cu²⁺] (µmol L⁻¹) + 0.172 (R² = 0.995) for Cu(II). Under optimal conditions of accumulation step and detection step, the influence of the concentration of Pb²⁺ and Cu²⁺ ions on the electrochemical response of the simultaneous detection of Cu(II) was also investigated for the case where the pH of the accumulation medium was equal to 6.0. The voltammograms recorded for concentrations of each of the Pb²⁺ and Cu²⁺ ions ranging from 0.15 to 1.0 µmol L⁻¹ after 10 min of open-circuit accumulation are shown in Fig. 11. Analysis of these voltammograms shows that the peak current intensities increase with the concentration of Pb²⁺ and Cu²⁺ ions in the accumulation medium. As the previous case, the calibration curve inset in Fig. 11 shown that, a linear relationship is observed between the peak current (Ip) and concentration of Pb²⁺ and concentration of Cu²⁺ according to the following equation: Ip(µA) = 120.190 [Pb^2+] (µmol L⁻¹) − 1.228 (R² = 0.998) for Pb(II) and Ip(µA) = 41.071 [Cu²⁺] (µmol L⁻¹) − 0.082 (R² = 0.995) for Cu(II). The influence of analyte concentration was also studied for a more dilute accumulation solution with a concentration range from 2.0 × 10⁻⁸ mol L⁻¹ and 12.0 × 10⁻⁸ mol L⁻¹. The voltammograms obtained for this study are presented as supplementary information in Fig. SI10. As in the two previous cases, analysis of the voltammograms obtained indicates that the peak current intensities increase linearly with the concentration of Pb²⁺ and Cu²⁺ ions in the accumulation medium as shown by the following equations: Ip(µA) = 186.464 [Pb^2+] (µmol L⁻¹) + 1.046 (R² = 0.998) for Pb(II) and Ip(µA) = 134.732 [Cu²⁺] (µmol L⁻¹) + 1.240 (R² = 0.996) for Cu(II). In order simultaneous detection and to ensure a better control of interference between the two heavy metal ions analysed, we decided to maintain the concentration of Pb²⁺ constant at 2.0 µmol L⁻¹ in the pH 6.0 accumulation medium and to vary the concentration of Cu²⁺ ions between 0.2 and 2.0 µmol L⁻¹ with steps of 0.4 µmol L⁻¹. The voltammograms obtained after 10 min of accumulation and at circuit are shown in Fig. 12. The intensity of the Pb(II) peak current shows very little variation for all the voltammograms recorded. This observation indicates that, the peak current of Pb(II) is not affect by the presence of Cu(II) in the accumulation medium for this range of concentration. On the other hand, the intensity of the Cu(II) peak current increases with its concentration in the accumulation medium. The calibration graph inset in Fig. 12 show a linear correlation between the peak intensity current and concentration of Cu²⁺ ions in accumulation medium (Ip(µA) = 30.612 [Cu²⁺] (µmol L⁻¹) − 0.076) with a correlation coefficient of 0.998. Similarly, the concentration of copper was fixed in the accumulation solution and the concentration of Pb(II) varied between 1.0 × 10⁻⁷ and 10.0 × 10⁻⁷ mol L⁻¹. The voltammograms obtained for this study and for two pH values of the accumulation medium pH 5.0 and 6.0 are shown in Supplementary Information respectively in Fig. SI11 and Fig. SI12. Examination of the voltammograms shows that the peak current of Pb(II) increases linearly with concentration according respectively to the equations Ip(µA) = 101.708 [Pb²⁺] (µmol L⁻¹) − 1.469 (R² = 0.996) and Ip(µA) = 125.878 [Pb²⁺] (µmol L⁻¹) + 1.173 (R² = 0.997). On the other hand, the Cu(II) peak current decreases progressively as the amount of Pb(II) in the accumulation medium increases, demonstrating once again that the KHEP sensor has a greater affinity for Pb²⁺ ions, which competitively displace Cu²⁺ ions to occupy its binding sites. The detection limits values calculate on the basis of a signal-to-noise ratio equal to 3, for each of the cases are reported in Table 2. These detection limits are for the most part comparable to those obtained in previous work with electrochemical sensors based on other clay materials such as natural or chemically modified smectites [37, 38, 48], chemically modified kaolinite [39, 41], or other geo-materials such as geopolymers [28] or natural or chemically modified hydroxyapatites [27]. Table 2 also shows the sensitivity values of the GCE/KHEP sensor and intercepts point of calibration graph, according to the pH of the accumulation medium. Table 3 shows the comparison of the performance of various clay materials modified electrodes, towards the voltammetric detection of Pb²⁺ and Cu²⁺ ions with organokaolinite clay material modified glassy carbon electrode build in this work.

Fig. 10

a ASSWV curves recorded for simultaneous detection of Pb²⁺ and Cu²⁺ ions, under optimised conditions in 0.2 mol L⁻¹ NaNO₃ (pH 2.5) at GCE/KHEP after 10 min of accumulation in deionized water (pH accumulation medium 5.0), as a function of the concentration of each ion, varied from 0.1 to 1.0 µmol L⁻¹ by step increasing 0.2 µmol L⁻¹. b Superposition of the calibration curves obtained for Pb²⁺ and Cu²⁺ ions

Fig. 11

ASSWV curves recorded for simultaneous detection of Pb²⁺ and Cu²⁺ ions under optimised conditions, in 0.2 mol L⁻¹ NaNO₃ (pH 2.5) at GCE/KHEP after 10 min of accumulation in deionized water (pH accumulation medium 6.0), as a function of the concentration of each ion varied from 0.15 to 0.90 µmol L⁻¹ by a step increasing of 0.15 µmol L⁻¹. Inset shows the calibration curves obtained for Pb²⁺ and Cu²⁺ ions

Fig. 12

Simultaneous ASSWV detection of Pb²⁺ and Cu²⁺ ions signal recorded at GCE/KHEP under optimised conditions, in 0.2 mol L⁻¹ NaNO₃ solution pH 2.5 after 10 min of accumulation in aqueous medium (pH accumulation medium 6.0) containing Pb²⁺ with concentration has been fixed at 2.0 µmol L⁻¹ and Cu²⁺ with the concentration range from 0.2 to 2.0 µmol L⁻¹ by step increasing 0.4 µmol L⁻¹. Inset shows the corresponding calibration graph

Table 2 The slope, intercept and the detection limits obtained by simultaneous detection of Cu²⁺ and Pb²⁺ ions after 10 min of accumulation in deionized aqueous medium for different pH and different concentration ranges

Table 3 Comparison of the performance of various clay materials modified electrodes, towards the voltammetric detection of Pb²⁺ and Cu²⁺ ions and other heavy metal ions such as Hg²⁺ and Cd²⁺

Table 4 Simultaneous determination of Pb²⁺ and Cu²⁺ ions after 10 min of accumulation in different real water samples

3.4 Interference study

The influence of the presence of other ionic species in the accumulation medium on the electrochemical response of simultaneous detection of Pb²⁺ and Cu²⁺ was studied. The percentages of variation in the anodic peaks currents of Pb²⁺ and Cu²⁺ ions for the simultaneous electrochemical detection, each ion concentrated 2.0 µmol L⁻¹ recorded in the presence of others ionic species introduced in relation to Pb²⁺ and Cu²⁺ concentration are presented in Table SI01. The results obtained and reported in Table SI01 shows that certain ions, even at low relative concentrations, have a considerable effect on the peak current of Pb²⁺ and Cu²⁺. Thus, when Hg²⁺ ions are present in the accumulation medium at a concentration 10 times greater than that of the analytes, the intensity of the Pb(II) peak current decreases by 0.5% and the intensity of the Cu(II) peak current by 45%. When Ag⁺ is present in the accumulation medium at an amount equal to 1 times the concentration of analytes, we observe a drop in the intensity of the lead peak current of 24.3% and a drop in the copper peak current of 43.5%. When the accumulation medium contains Zn²⁺ ions at a concentration 20 times greater than that of the analytes, we observe a decrease of less than 5% in the intensity of the Pb²⁺ peak current and a decrease in the intensity of the Cu²⁺ peak current of almost 10%. Generally speaking, we can see that the Pb(II) peak current is only slightly influenced by the presence of ions (Hg^2+, Ag⁺, Zn²⁺ and Ni²⁺), whereas the intensity of the Cu(II) peak current is greatly affected by the presence of the same ions in the accumulation medium. This can be explained by the affinity that these ions also have with the amine group (-NH-) of KHEP. There is therefore genuine competition between Cu²⁺ ions and these ions for binding sites by complexation. The presence of Cd²⁺ ions in the accumulation medium at a concentration 10 times higher than that of the analytes had very little effect on the Cu(II) peak current and almost no effect on the Pb(II) peak current. A 50-fold excess of Fe³⁺ and Fe²⁺, a 100-fold Mg²⁺ and a 1000-fold excess of Na⁺, K⁺, Cl⁻, SO₄²⁻ and NO₃⁻ in the accumulation medium were found to have no considerable effect on the Pb(II) and Cu(II) electrochemical peak current.

3.4.1 Analytical application of the simultaneous method for detection of Pb²⁺ and Cu²⁺ to the real samples analysis

The electrochemical simultaneous determination of Pb²⁺ and Cu²⁺ in laboratory tap water, spring water and well water and was performed using the method already described in Sect. 2.4.2. Fig. SI13a, and Fig. SI14a show in (a) the square-wave voltammograms of simultaneous detection of Pb(II) and Cu(II) recorded in NaNO₃ medium (pH adjusted at 2.5) after 10 min accumulation, respectively in laboratory well water and tap water (i) without addition of analytes Pb²⁺ and Cu²⁺ (ii-vii) with addition of analytes for concentrations, each ranging from 1.0 × 10⁻⁷ and 10.0 × 10⁻⁷ mol L⁻¹ and in (b) the different calibration lines derived from them. The calibration graph obtained in different types of natural water, presented in Fig. SI13b for the case of well water and Fig. SI14b for the case of tap water showed a linear correlation between the peak intensity current and concentration of analytes (Tables SI02 and S03). Analysis of voltammograms obtained for different case shows that the first ASSWV signal has a peak current centred on − 0.3 V and corresponding to the copper current. This result shows that the chosen natural water is already contaminated with copper. On the basis of the different calibration curves drawn, the initial concentration of copper in real water is 3.2 × 10⁻⁷ mol L⁻¹ in tap water, 3.8 × 10⁻⁷ mol L⁻¹ in well water and 3.4 × 10⁻⁷ mol L⁻¹ in spring water. However, it is important to notice that all these initial concentrations of copper in natural water are well below 3.14 × 10⁻⁵ mol L⁻¹, which is limit concentration according WHO standard. For each quantity of Pb²⁺ and Cu²⁺ ions added to a sample of real water, analysis was done five times using the GCE/KHEP sensor and the results obtained are listed in Table 4. By spiking the laboratory tap water with 0.4 µmol.L⁻¹ of Pb(II) and 0.4 µmol L⁻¹ of Cu(II), a mean recovery of 0.390 µmol L⁻¹ and 0.394 µmol L⁻¹ was found respectively, corresponding to 97.50% of the added Pb(II), 98.50% of the added Cu(II) and a RSD of 1.85% and 2.25% respectively. Spring water collected in Yaoundé town (Nkolbisson quarter) was spiked with 0.4 µmol L⁻¹ Pb(II) and 0.4 µmol L⁻¹ Cu(II) a mean recovery of 0.391 µmol L⁻¹ and 0.385 µmol L⁻¹ was found respectively, corresponding to 97.75% of the added Pb(II), 96.25% of the added Cu(II) and a RSD of 2.55% and 2.86% respectively. As in the two previous environments, we also prepared a solution of Pb²⁺ and Cu²⁺ ions at a concentration of 0.4 µmol L⁻¹ in well water, and using the GCE/KHEP sensor implemented in this work, and the ASSWV method, we were able to detect a concentration of 0.396 µmol L⁻¹ for Pb(II) and 0.389 µmol L⁻¹ for Cu(II), equivalent to a recovery rate of 98.75% for Pb(II) and 97.25 for Cu(II), a RSD of 2.72% and 2.94% respectively. These results show that the voltammetric method based GCE/KHEP sensor developed in this work can be successfully applied to simultaneous analysis of Pb²⁺ and Cu²⁺ ions in real water samples polluted by this heavy metal. These results also showed that, it is possible effectively to use GCE/KHEP sensor, for the simultaneous electrochemical determination of Cu²⁺ and Pb²⁺ in a natural environment.

4 Conclusion

The aim of this work was to prepare an organokaolinite and apply it as electrode material to the simultaneous detection of Pb(II) and Cu(II). The organokaolinite was synthesised by intercalation and grafting of 1-(2-hydroxyethylpiperazine) (HEP) to the interlayer space of kaolinite. To better understand the structure of organokaolinite and ensure that the synthesis was successful, XRD, FTIR and ²⁷Si NMR characterisations were carried out. The results show that the kaolinite was successfully modified. The XRD results show an increase in the interlayer distance from 7.1 to 11.2 Å for organokaolinite (KHEP). The results obtained from the FTIR characterisation show the presence of new bands and new adsorption peaks in the spectrum of the KHEP material attributable to the intercalated HEP molecules. A glassy carbon electrode (GCE) was further modified with a film of the kaolinite (referred as GCE/K), then with the film of organokaolinite (noted GCE/KHEP) and used to study the electrochemical behaviour of [Fe(CN)₆]³⁻ ions by cyclic voltammetry. Multicyclic scanning voltammetry signals recorded in analysis medium pH 2.0 show a progressive accumulation of [Fe(CN)₆]³⁻ due certainly by strong protonation of the amine groups (–NH–) of the 1-(2-hydroxyethyl)piperazine molecules grafted into the organokaolinite. The ASSWV electrochemical signal of simultaneous detection of Pb(II) and Cu(II) recorded under the same conditions of accumulation and detection on GCE/KHEP is much more intense than that obtained on GCE/K and that obtained on unmodified GCE. The parameters that influence the accumulation and detection steps have been optimised. The sensitivity of the GCE/KHEP sensor for either of the two ions analysed varies greatly with the pH of the accumulation medium. This enabled us to determine different limits of detection on the basis of the signal-to-noise ratio out of 3. Thus, in simultaneous detection and for the concentration range from 0.02 to 0.12 µmol L⁻¹ (pH accumulation medium 5.0) and for each of the ions analysed, we obtained a LOD of 1.942 nmol L⁻¹ for copper and a LOD of 1.072 nmol L⁻¹ for lead. For the concentration range from 0.15 µmol L⁻¹ to 0.90 µmol L⁻¹ (pH accumulation medium 6.0), we obtained a LOD of 24.562 nmol L⁻¹ for copper and a LOD of 9.214 nmol L⁻¹ for lead. The GCE/KHEP sensor developed in this work was successfully applied to the simultaneous detection of Cu(II) and Pb(II) in real environments (tap water, spring water and well water).