Introduction

Removal of heavy metal ions from various types of waters and waste waters is an essential problem in environmental protection and has a result on human health. On the other hand, stricter requirements concerning the quality of waste waters lead to more refined and complex chemical processes used in their purification. Therefore, the search for new methods and novel materials for heavy metal ions removal, particularly of As(V) and Cr(VI) ions, is very important [1–8]. Some of them are hybrid ion exchangers (HIX) based on nanosized metal oxides (NMOs) such as ferric, manganese, aluminum, titanium, magnesium, and cerium oxides. For example, the addition of hydrated iron oxides to the polymers (PS and DVB, less frequently polyacrylic, with a macroporous or gel matrix, containing cation or anion exchange functional groups) leads to the obtaining of adsorbents characterized by very good adsorption, mechanical, and hydraulic properties [9, 10]. The quality of the sorbents obtained depends on: (i) the quantity, structure, crystallinity, and degree of dispersion of the iron oxides introduced; (ii) physical form of the base polymer; and (iii) the type of functional groups involved in the preparation of hybrid polymers as well as influence on the process of adsorption. For example, Qiu et al. [11] found that the hybrid sorbent based on the hydrous iron oxide and neutral polystyrene matrix (PS) without an attached functional group (HFO–PS) and based on the negatively charged HFO–PS–SO3 − matrix had totally different adsorption properties toward Cu(II) ions in the presence of sulfate(VI) ions. The adsorption of Cu(II) on HFO–PS was promoted in sulfate solutions; however, when influenced by covalently attached sulfonic acid groups, Cu(II) adsorption on HFO–PS–SO3 − was inhibited by the added sulfates.

As described by Sarkar et al. [12], the first step in the preparation of HIX based on iron oxides and polystyrene–divinylbenzene copolymers (PS–DVB) includes loading of Fe(III) onto the functional groups of ion exchanger, then desorption and simultaneous precipitation of Fe(III) hydroxides using a solution containing NaCl and NaOH, and finally rinsing and washing the resin with an ethanol–water mixture followed by a mild thermal treatment. The high concentration of the functional groups allowed high and fairly uniform loading of hydrous iron oxide particles (~9–12 % of Fe by mass) within the polymeric matrix. Such ions as Cl−, SO4 2−, and HCO3 − exhibit only weak affinities for hydrous iron oxides and are characterized by a smaller selectivity. Cumbal and SenGupta [5] reported the preparation of two classes of polymer-supported nanoparticles containing iron for As(III,V) removal: (i) hydrated Fe(III) oxides dispersed in an ion exchange resin, and (ii) polymeric particles with magnetic properties.

Increasing the amount of iron oxide introduced into the polymer increases the adsorption capacity of the sorbent, but after crossing saturation of the polymer matrix followed by the blocking of pores, a significant decrease in adsorption capacity is observed. Typically, the iron oxide content in the copolymer is equal to 100–200 mg Fe g−1 of sorbent. The best adsorption properties of hybrid polymers are found for those in which the inner surfaces of the pores are coated with a thin layer of iron oxides. This effect can be achieved by proper selection of the polymer matrix, the amount of iron placed in a single cycle of synthesis, and by performing appropriate treatment. These factors also affect the degree of dispersion of the introduced iron oxides. The introduction of smaller agglomerates provides a greater surface area and thus a more effective adsorbent.

In this article, the HIX Purolite Arsen Xnp (Purolite) and Lewatit FO36 (Lanxess) were used to remove Cu(II), Zn(II), Cd(II), and Pb(II) ions in the presence of ethylenediamine-N,N′-disuccinic acid (EDDS) which belongs to a group of readily biodegradable complexing agents. The metals for the evaluation were chosen due to their occurrence in waste water streams. Optimal conditions of adsorption were based on the speciation distribution of the studied heavy metal complexes with EDDS. The adsorption properties of the above-mentioned sorbents were also evaluated in the case of arsenic and chromium.

Investigations were carried out using the static method, based on which adsorption parameters were calculated in single versus multicomponent systems at different pH values and temperatures as well as in the presence of interfering ions. To compare the surface morphologies of the studied ion exchangers before and after the adsorption process, scans using atomic force microscope (AFM) were also recorded, and infrared spectroscopy (ATR/FT-IR) was applied to establish the adsorption mechanism.

Materials and methods

Purolite Arsen X np also known as FerriIX™ A33E is a strongly acidic macroporous cation exchanger based on a polystyrene-divinylbenzene skeleton with attached sulfonic acid groups and containing hydrated iron oxide nanoparticles. The resin has the bead size of 0.3–1.2 mm; adsorption capacity 1.56 meq L−1; and iron oxide content not exceeding 42 %. It is also characterized by an enormous capacity for As(V) [13].

Lewatit FO36 is a weakly basic, macroporous anion exchanger also based on PS–DVB and containing the weak amine functional groups and iron oxide nanoparticles. It has the bead size 0.34–0.38 mm, adsorption capacity 1.5 eq L−1 (30 BV h−1, 0.1 mg L−1 As(V), 6 mg L−1 SiO2, and 60 mg L−1 PO4 3−), and iron oxide content not exceeding 23 % [14]. The following salts were used: CuCl2·2H2O, ZnCl2, Cd(NO3)2·4H2O, and Pb(NO3)2 as a source of metal ions, and the S,S-isomer of EDDS (Innospec Speciality Chemicals, UK) used as supplied to react with these metal ions in solution without pH adjustment. In addition, a Cr(VI) working stock solution was prepared by dissolving K2Cr2O7 in deionized water. Experiments were carried out in multicomponent systems containing Cr(VI), Cu(II), and Zn(II) ions, with or without the complexing agent (EDDS) as well as in the presence of accompanying ions Cl−, NO3 −, and SO4 2−. In the investigations mixtures of Cd(II) and Pb(II), with and without EDDS were also used. In these systems, metal ions are complexed in the form of [M(edds)]2− ions, whereas, Cr(VI) is not complexed.

Batch kinetic experiments were conducted in 250-mL conical flasks containing ion exchanger (2 g) to which a solution of metal ions (200 mL) was added, with and without EDDS and with the initial pH values of 4.5 and 10.0. The flasks were shaked on a rotary shaker at 180 rpm. Batch studies were carried out to investigate the effects of phase contact time (1–180 min), initial metal concentration (1–5 mmol L−1), solution pH (4–12), temperature (20–60 °C), and the presence of other ions (Cl−, NO3 −, SO4 2−). Solution pH was adjusted with 0.1-mol L−1 HNO3 or 0.1-mol L−1 NaOH.

A laboratory shaker Elpin type 357, (Elpin-Plus, Poland) was used for agitation. The pH values were measured with a PHM 84 pH meter (Radiometer, Denmark) with glass REF 451 and calomel pHG 201-8 electrodes. The accuracy of the pH meter is ±0.01 pH units. Concentrations of the heavy metals were measured with an AAS spectrometer SpectrAA 240FS (Varian, Australia). Measurements were performed using the following parameters: wavelength Cu(II): 324.70 nm, Zn(II): 213.90 nm, Cd(II): 228.80 nm, and Pb(II): 217.00 nm with slit width = 0.50 nm. Cr(VI) ions in the effluent were determined colorimetrically using 1,5-diphenyl-carbazide method at 546 nm with a Specord M42 (Carl Zeiss, Jena) UV–Vis spectrophotometer. The morphology of the hybrid sorbent samples was analyzed with AFM Digital Instruments (USA) NanoScope III. The infrared spectra were recorded by a Nicolet-380 Fourier transform infrared spectrometer using the ATR method. All experiments were made in triplicate, and the results were taken as the average value.

The amounts of metal ions adsorbed at time “t” of hybrid sorbents (q t , mg g−1) were calculated from the mass balance equation:

$$ q_{t} = {{\left( {c_{0} - c_{t} } \right)V} \mathord{\left/ {\vphantom {{\left( {c_{0} - c_{t} } \right)V} m}} \right. \kern-0pt} m}, $$

where c 0 is the initial concentration of a metal ion (mg L−1); c t is the concentration of metal ion at time t (mg L−1), V is the volume of the solution (L); and m is the mass of HIX (g).

The equilibrium data were analyzed using the Langmuir and Freundlich isotherms and the characteristic parameters for each isotherm determined. The linear form of the Langmuir model can be expressed as [15]

$$ {{c_{\text{e}} } \mathord{\left/ {\vphantom {{c_{\text{e}} } {q_{\text{e}} = {{c_{\text{e}} } \mathord{\left/ {\vphantom {{c_{\text{e}} } {q_{0} + {1 \mathord{\left/ {\vphantom {1 {K_{\text{L}} }}} \right. \kern-0pt} {K_{\text{L}} }}q_{0} }}} \right. \kern-0pt} {q_{0} + {1 \mathord{\left/ {\vphantom {1 {K_{\text{L}} }}} \right. \kern-0pt} {K_{\text{L}} }}q_{0} }}}}} \right. \kern-0pt} {q_{\text{e}} = {{c_{\text{e}} } \mathord{\left/ {\vphantom {{c_{\text{e}} } {q_{0} + {1 \mathord{\left/ {\vphantom {1 {K_{\text{L}} }}} \right. \kern-0pt} {K_{\text{L}} }}q_{0} }}} \right. \kern-0pt} {q_{0} + {1 \mathord{\left/ {\vphantom {1 {K_{\text{L}} }}} \right. \kern-0pt} {K_{\text{L}} }}q_{0} }}}}, $$

where c e is the equilibrium concentration of metal ions (mg L−1) and q e is the amount of the metal ions adsorbed on the hybrid sorbent (mg g−1), and q 0 and K L are the Langmuir constants related to the adsorption capacity (mg g−1) and the equilibrium constant (L g−1), respectively.

The adsorption equilibrium data were also applied to the Freundlich model [16]:

$$ { \log }\;q_{\text{e}} = { \log }\;K_{\text{F}} + \left( {{1 \mathord{\left/ {\vphantom {1 n}} \right. \kern-0pt} n}} \right){ \log }\;c_{\text{e}} , $$

where K F and n are the Freundlich constants related to the adsorption capacity and the adsorption intensity, respectively.

The pseudo-first-order model (PF-order) was presented by Lagergren and expressed as follows [17]:

$${\text{log(}}q_{e} - q_{t} ) = {\text{log}}q_{e} - (k_{1} /2.303)t $$

where q t is the amount of metal ions adsorbed on the hybrid sorbents at time t, and k 1 is the rate constant (min−1). The rate constant, k 1, was obtained from the slope of linear plots of log(q e − q t ) against t.

The adsorption data were also analyzed in terms of the pseudo-second order mechanism (PS-order) [18, 19]:

$$ {t \mathord{\left/ {\vphantom {t {q_{t} }}} \right. \kern-0pt} {q_{t} }} = \left( {{1 \mathord{\left/ {\vphantom {1 h}} \right. \kern-0pt} h}} \right) + \left( {{1 \mathord{\left/ {\vphantom {1 {q_{\text{e}} }}} \right. \kern-0pt} {q_{\text{e}} }}} \right)t, $$

and the initial rate of adsorption h is

$$ h = k_{2} q_{\text{e}}^{2} , $$

where k 2 is the rate constant of PS-order adsorption (g(mg min)−1), and h is the initial rate of adsorption (mg(g min)−1).

The constants q e, h, and k 2 can be determined from the plots of t/q t against t.

Results and discussion

The study showed that the hybrid sorbents, including Purolite Arsen Xnp and Lewatit FO36, can be used to remove Cu(II), Zn(II), Cd(II), Pb(II), and Cr(VI) under appropriate conditions. The effectiveness of the adsorption process depends on the initial concentration of solution, pH, phase contact time, addition of a complexing agent EDDS and presence of accompanying ions.

It was found that the optimal pH values for the adsorption of Cu(II), Zn(II), Cd(II), Pb(II), and Cr(VI) ions on Purolite Arsen Xnp and Lewatit FO36 are above 4.5 (Fig. 1).

Fig. 1
figure 1

Effect of pH on the adsorption of Cu(II) and Cr(VI) on Purolite Arsen Xnp and Lewatit FO36

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Under these conditions, Cu(II), Zn(II), Cd(II), and Pb(II) ions are coordinated to the goethite phase. It should be mentioned that, for example, Cu(II) at pH between 2 and 6.5 occurs predominantly as the aqueous Cu2+ cation. With increasing pH, the following hydroxoions occur Cu(OH)+, Cu(OH)2, Cu(OH) −3 , and Cu(OH) 2−4 . At pH values >7, the major hydrolysis product is Cu2(OH) 2+2 . Therefore, at pH values >4.5, Cu(II) is likely to sorb as the aqueous cation. Under analogous conditions, other ions are also in the form of cations or hydroxoions. Numerous studies show that the Cu(II) ions are sorbed on goethite as the surface complexes ≡FOCu+ and ≡FOCuOH or a combination of both [20, 21].

However, as complexing agents are present, these are adsorbed specifically on metal oxides and hydroxides to form inner sphere complexes. It is also known that they are responsible for increase, decrease, or no change in the adsorption of metal ions, for example on soils. Some authors confirm that the adsorption of metals in systems containing organic ligands increases at pH values below the adsorbent point of zero charge (PZC) and increases at higher pH [22]. However, other authors showed that the addition of organic substances decreases the adsorption even under acid conditions below the PZC [23].

In the present article, ethylenediamine-N,N′-disuccinic acid (H4edds) consists of four isomers: S,S- (25 %), R,R- (25 %), R,S- (50 %), and S,R- (50 %); however, only the S,S-isomer of EDDS is readily biodegradable. It was proved that 83 % of it is converted to CO2 within 20 days [24]. The other isomers are either partly or completely nonbiodegradable [25, 26]. It follows from the EDDS speciation profile at pH 10 that the predominant form is the anion (edds)4− (Fig. 2).

Fig. 2
figure 2

Speciation profile of EDDS versus pH

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EDDS has four ionisable protons with the following protonation constants:

$$ \begin{aligned} {\text{edds}}^{ 4- } + {\text{H}}^{ + } \rightleftarrows & {\text{Hedds}}^{ 3- } ,\quad & { \log }\;K_{\text{p1}} = 9. 70 \\ {\text{Hedds}}^{ 3- } + {\text{H}}^{ + } \rightleftarrows & {\text{H}}_{ 2} {\text{edds}}^{ 2- } ,\quad & { \log }\;K_{\text{p2}} = 7.0 1\\ {\text{H}}_{ 2} {\text{edds}}^{ 2- } + {\text{H}}^{ + } \rightleftarrows & {\text{H}}_{ 3} {\text{edds}}^{ - } ,\quad & { \log }\;K_{\text{p3}} = 3. 9 2\\ {\text{H}}_{ 3} {\text{edds}}^{ - } + {\text{H}}^{ + } \rightleftarrows & {\text{H}}_{ 4} {\text{edds}},\quad & { \log }\;K_{\text{p4}} = 3.0 3\\ \end{aligned} $$

and (Σ log K pi  = 23.66, i = 1–4) [27, 28].

According to the literature data, EDDS showed good complexing properties toward heavy metal ions, and the reaction is characterized by the formation of stable 1:1 metal to ligand complexes as the major species:

$$ {\text{M}}^{ 2+ } + {\text{edds}}^{ 4- } \rightleftarrows \left[ {{\text{M}}\left( {\text{edds}} \right)} \right]^{ 2- } . $$

For the Cu(II), Zn(II), Cd(II), and Pb(II) complexes with EDDS, the stability constants are as follows: 13.1, 13.4, 16.4, and 12.7, respectively.

According to the literature data, metal capacity on the goethite surface increases with metal electronegativity: Cu(II) > Pb(II) > Cd(II) [29]. In addition, the behavior of the metals in the three-component system consisting of metal solution; organic ligand and hybrid sorbent can be controlled by several processes: (i) the adsorption of the free metal cations on the sorbent surface, (ii) the adsorption of the metal cations by the sorbent surface modified by interaction with the anions of dissolved EDDS, (iii) the formation of metal complexes with EDDS in solution, (iv) the adsorption of free and charged metal complexes with EDDS on the sorbent surface, and finally (v) the interaction of anionic complexes with the functional groups of the hybrid sorbents.

To establish which of the above-mentioned steps has decisive influence on the sorption process in the next stage, the effect of the phase contact time on the effectiveness of Cu(II), Zn(II), Cd(II), Pb(II), and Cr(VI) ions’ adsorption at different initial concentrations (1–5 mmol L−1) was studied, and the examples of results for Purolite Arsen Xnp are presented in Fig. 3.

Fig. 3
figure 3

Effect of the phase contact time on the adsorption of Cr(VI), Cu(II), Zn(II), Cd(II), and Pb(II) at different initial concentrations on Purolite Arsen Xnp

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It was found that in the case of Cr(VI) adsorption, the equilibrium on Purolite Arsen Xnp was established within 60 min, in the case of Cu(II) and Zn(II) within 30–40 min. (at low concentrations), and for Cd(II) and Pb(II) even <30 min. For Lewatit FO36, the equilibrium times were almost analogous. Moreover, it was found that the effectiveness of adsorption for these ions on Purolite ArsenXnp was higher than on Lewatit FO36.

Based on the kinetic analysis, it was found that the rate constants k 1 decreased with increase in the initial metal ion concentrations (data not presented for the concentrations 3 and 5 mmol L−1). The correlation coefficients were high, ranging from 0.8253 to 0.9284 for Purolite Arsen Xnp and from 0.9562 to 0.9982 for Lewatit FO 36. However, the calculated equilibrium capacities (q 1) according to the Lagergren pseudo-first order rate expression were in disagreement with the experimental capacities (q e,exp) for solutions with initial concentrations 1–5 mmol L−1. Better results were obtained using the pseudo-second order equation (Table 1).

Table 1 Kinetic parameters for Cr(II), Cu(II), Zn(II), Cd(II), and Pb(II) ion sorption on Purolite Arsen Xnp and Lewatit FO36 at initial concentration 1 mmol L−1
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In addition, studies on the effect of ethylenediamine-N,N′-disuccinic acid (EDDS) addition clearly showed that the influence of complexing agent is different depending on the system (Figs. 4, 5). For Purolite ArsenXnp, addition of EDDS had a negative impact on the adsorption of Cu(II) and Pb(II), however, for Cd(II) an increase of the removal efficiency was observed. In the case of Lewatit FO36, much better adsorption in the presence of EDDS was found for Cu(II), Zn(II), and Cd(II), while it had no effect on Cr(VI) adsorption. As for Pb(II) ions, their removal from waste waters decreases. The increase in the metal adsorption is probably related to the formation of stable negatively charged M(II)-EDDS complexes of the type M[(Hedds)]3− or [M(edds)]2−. These complexes increase the adsorption capacity of the surface due to the involvement of positively charged sorption sites into the adsorption process. Their adsorption on the hybrid sorbents surface can occur via the metal–surface interaction and through the organic ligand [20, 30].

$$ \begin{aligned} \equiv & {\text{Fe}}{-}{\text{OH}} + {\text{M}}^{ 2+ } + {\text{edds}}^{{{\text{n}} - }} + {\text{H}}^{ + } \rightleftarrows \equiv {\text{Fe}}{-}{\text{edds}}{-}{\text{M}}^{{({\text{n}} - 3) - }} + {\text{H}}_{ 2} {\text{O}} \\ \equiv & {\text{Fe}}{-}{\text{OH}} + {\text{M}}^{ 2+ } + {\text{edds}}^{{{\text{n}} - }} \rightleftarrows \equiv {\text{Fe}}{-}{\text{O}}{-}{\text{M}}{-}{\text{edds}}^{{({\text{n}} - 1) - }} + {\text{H}}^{ + } \\ \end{aligned} $$

where M2+ is the metal cation and edds4− is the ligand.

Fig. 4
figure 4

Comparison of sorption properties of Purolite Arsen Xnp toward Cr(VI), Cu(II), Zn(II), Cd(II), and Pb(II) ions without and with the addition of EDDS

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Fig. 5
figure 5

Comparison of sorption properties of Lewatit FO36 toward Cr(VI), Cu(II), Zn(II), Cd(II), and Pb(II) ions without and with the addition of EDDS

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Moreover, especially for Lewatit FO36 with weakly basic functional groups, the anion exchange process should be considered, and this mechanism is probably responsible for the enhanced adsorption of the negatively charged complexes:

$$ 2 {\text{RCl}} + \left[ {{\text{M}}\left( {\text{edds}} \right)} \right]^{ 2- } \rightleftarrows {\text{R}}_{ 2} \left[ {{\text{M}}\left( {\text{edds}} \right)} \right] + 2 {\text{Cl}}^{ - } . $$

As was shown by Ali and Dzombak [31], sulfate ligands can dramatically enhance the adsorption of heavy metals on iron oxides by modifying the electrostatic environment at the interface or by forming metal-sulfate ternary surface complexes like ≡Fe–OHCuSO4. As was mentioned earlier, Qiu et al. [11] also found that the Cu(II) adsorption on a hybrid sorbent based on the PS–SO3 − matrix is strongly inhibited by the presence of the SO4 2− ligand. Therefore, the effects of accompanying ions such as Cl−, NO3 −, and SO4 2− were studied. From the obtained results, it should be noted that the effects of accompanying ions are negligible; however, the addition of the complexing agent EDDS to the system in the presence of Cl−, NO3 −, and SO4 2− ions improves the sorption process parameters for both hybrid sorbents. This is mainly noticeable in the case of Zn(II), where the addition of EDDS significantly improves the sorption capacity from 14.43 to 21.50 mg g−1 for Purolite ArsenXnp as well as from 14.43 to 21.50 mg g−1 for Lewatit FO36 (initial concentration 3 mmol L−1). In the case of Cr(VI), this increase is negligible, and in the case Cu(II) ions the sorption capacity decreases from 18.46 to 9.22 mg g−1 in the same conditions (Figs. 6, 7).

Fig. 6
figure 6

Comparison of sorption properties of Purolite Arsen Xnp toward Cr(VI), Cu(II), and Zn(II) ions in the presence of Cl−, NO3 −, SO4 2− ions, and EDDS, Cl−, NO3 −, SO4 2− ions

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Fig. 7
figure 7

Comparison of sorption properties of Lewatit FO36 toward Cr(VI), Cu(II), and Zn(II) ions in the presence of Cl−, NO3 −, SO4 2− ions, and EDDS, Cl−, NO3 −, SO4 2− ions

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In addition, it is found that the sorption capacity of the sorbents slightly increased when the temperature of heavy metal solutions increased from 20 to 60 °C. The sorption of the above complexes could be described by the Langmuir isotherm with high correlation coefficients ranging from 0.9986 to 1.000, whereas for the Freundlich isotherm they range from 0.8345 to 0.9871. The values of the separation factor, R L, indicate the nature of the adsorption process with: R L > 1 unfavorable; R L = 1 linear; 0 < R L < 1 favorable; and R L = 0 irreversible. In the present study, these values indicate that the adsorption process is favorable for both the Purolite ArsenXnp and Lewatit FO36 samples. The maximum capacities of adsorption of Cr(VI), Cu(II), Zn(II), Cd(II), and Pb(II) on Purolite Arsen Xnp are equal to 25.55, 21.47, 10.11, 28.99, and 81.85 mg g−1, respectively, and they increase with the increasing temperature. Meanwhile, the Langmuir equilibrium constants (K L) (L g−1) for Purolite Arsen Xnp decrease as follows: Cr(VI): 0.240–0.017; Cu(II): 0.185–0.014; Zn(II): 0.122–0.0087; Cd(II): 0.221–0.013; and Pb(II): 0.324–0.025. The Langmuir sorption isotherms of Cd(II) and Pb(II) on Lewatit FO36 and Purolite Arsen Xnp are presented in Fig. 8. Moreover, the Langmuir and Freundlich parameters of sorption isotherms of Cd(II) and Pb(II) on hybrid sorbents are presented in Table 2. The comparison between the determined sorption capacities of Purolite ArsenXnp and Lewatit FO36 toward heavy metal ions is presented in Table 3.

Fig. 8
figure 8

The Langmuir sorption isotherms of Cd(II) and Pb(II) on Lewatit FO36 and Purolite Arsen Xnp (1.0 × 10−3–2.5 × 10−2 M, shaking time 24 h, shaking speed 180 rpm, temperature 295 K)

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Table 2 The Langmuir and Freundlich parameters for adsorption of Cd(II) and Pb(II) on Lewatit FO36 and Purolite Arsen Xnp (1.0 × 10−3–2.5 × 10−2 M, shaking time 24 h, shaking speed 180 rpm, temperature 295 K)
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Table 3 Comparison of adsorption capacities for heavy metal ions onto different sorbents containing hydrous iron oxides
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It can be observed that the adsorption capacity of both ion exchangers is equal and also higher than that of some other sorption materials reported in the literature. For example, Lee et al. [32] found that sorption capacity of the iron oxide nanoparticles immobilized sand for Cd(II) 1.26 mg g−1 and for Pb(II) 2.08 mg g−1, respectively. For the used ion exchangers Lewatit FO36 and Purolite Arsen Xnp, these values are equal to 15.93, 3.02, 17.14, and 8.20 mg g−1, respectively. However, it should be mentioned that in the case of applicability of EDDS, the mechanism of sorption is quite different and connected with the presence of anionic complexes (ion exchange). In the literature data, there is no information concerning the applicability of ethylenediaminedisuccinic acid as a complexing agent for heavy metal ion removal on HIX based on the hydrous iron oxides. EDDS was only used for heavy metal ion removal in the anionic complexes on anion exchangers, as shown in [39–41]. The preliminary results for heavy metal ion removal on Lewatit FO36 were also presented in our previous article [42]. Such biodegradable chelating agents as EDDS are commonly used to enhance the heavy metal extraction from the contaminated soils. EDDS causes a much lower potential metal leaching than EDTA, and it is less toxic to soil microorganisms [43].

As was previously mentioned, the sorption of anionic complexes of Cu(II), Zn(II), Cd(II), and Pb(II) with EDDS is connected with the ion exchange process. This mechanism was confirmed by the ATR/FT-IR analysis. The exemplary results for Purolite Arsen Xnp and Lewatit FO36 are presented in Fig. 9. In the recorded spectra, the stretching vibrations at 3290 and 3170 cm−1 are connected with the presence of hydroxyl and amine groups (Lewatit FO36). Other characteristic bands appear at the wavelengths 1636 cm−1 and 1478 cm−1. These are related to the presence of the NH deformation vibration and that of symmetric and asymmetric ν asCOO− and ν sCOO−. In addition, the bands at 1031 and 1086 cm−1 are connected with the stretching vibrations of CO group. Changing the pH of the solution causes a slight shift of various bands, Ø most noticeably in the case with Purolite Arsenic Xnp.

Fig. 9
figure 9

FT-IR images of the HIX Purolite Arsen Xnp and Lewatit FO36 before and after the adsorption process

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The AFM images illustrate the changes in the surface of the HIX Purolite Arsen Xnp used in the studies before and after sorption (Fig. 10). No significant changes in the surface morphology were observed before and after the sorption. Based on the observations, it was found that Cr(VI) ions present in the solution penetrate inside the ion exchanger pores according to the reaction:

$$ 2 {\text{R}}-{\text{N}}^{ + } \left( {{\text{CH}}_{ 3} } \right)_{ 3} {\text{Cl}}^{ - } + {\text{ Cr}}_{ 2} {\text{O}}_{ 7}^{ 2- } \rightleftarrows [{\text{R}} - {\text{N}}^{ + } \left( {{\text{CH}}_{ 3} } \right)_{ 3} \left] {_{ 2} } \right[{\text{Cr}}_{ 2} {\text{O}}_{ 7} ]^{ 2- } + {\text{ 2Cl}}^{ - } . $$
Fig. 10
figure 10

AFM scans of Purolite Arsen Xnp before (a) and after (b) sorption of Cr(VI)

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Desorption of hybrid sorbents proceeds easily on alkalization of the adsorbent surface. In an alkaline environment, the dominant form on the surface of iron oxides are ≡Fe–O− groups, and they adsorb anions in solution. Some ions may, however, remain inside the polymer by reducing the effective sorption capacity by 10–20 %. It has been shown in subsequent regeneration cycles that this effect gets smaller. Regeneration is carried out with a mixture of NaOH and NaCl or NaOH (4–10 %). Depending on conditions and sorption properties of the hybrid material, in the regeneration process, 90–98 % ions are removed. In order to increase the efficiency of regeneration, this process is sometimes carried out at elevated temperatures.

Conclusions

Based on the results obtained, both Purolite Arsen Xnp and Lewatit FO36 may find use in removing heavy metal ions such as Cu(II), Zn(II), Cd(II), and Pb(II) in the presence of EDDS. The interactions of metal ions with the hybrid sorbents are complex, probably simultaneously dominated by adsorption and ion exchange. The pH dependence of ion exchange may suggest that the metal ions are adsorbed according to an ion exchange mechanism. The effect of sulfate ligands on metal adsorption to hybrid sorbents is largely dependent on the surface properties of these materials.

On the basis of the studies, it can be stated that the effectiveness of sorption for Cr(VI), Cu(II), Zn(II), Cd(II), as well as Pb(II) ions depends on the conditions under which the process is carried out. The main parameters affecting sorption are initial concentration of the solution, pH, and phase contact time. Temperature has only a slight influence. With respect to effectiveness of the removal of Cr(VI), Cu(II), Zn(II), Cd(II), and Pb(II), the hybrid sorbents can be arranged as Purolite Arsen Xnp > Lewatit FO36. It should also be emphasized that the best results are obtained for Cr(VI) ions. Therefore, they can be applied in systems of drinking and industrial water purification.