A comprehensive review on LDH-based catalysts to activate persulfates for the degradation of organic pollutants

Abstract

Degradation of organic contaminants into less toxic substances is the best option to remove these compounds rather than using conventional techniques. The sulfate radical-based-advanced oxidation process is an effective strategy that degrades organic contaminants by activating peroxymonosulfate (PMS). Such a strategy generates singlet oxygen (¹O₂), hydroxyl ($^ \bullet \!{{{\mathrm{OH}}}}$), and sulfate (${{{\mathrm{SO}}}}_4^{ \bullet\! - }$) radicals. ${{{\mathrm{SO}}}}_4^{ \bullet \!- }$ is distinguished by its high oxidation selectivity and activity toward the degradation of organic contaminates compared to other radicals. Various catalysts are employed in PMS activation including layered doubled hydroxides (LDHs), which are characterized by their facile synthesis and high catalytic activity. This review article is the first attempt to compile the recent progress in the degradation of common organic pollutants including aromatic compounds, pharmaceutical residues, and dyes via the PMS activation using LDH-based catalysts. The degradation pathways, reaction parameters’ influence, stability of LDHs, and comparisons between different LDH-based catalysts are investigated in this work.

Introduction

Recently, with the augmentation of urbanization and industrialization to meet human needs, a humongous amount of toxic wastes has been released into the environment^1,2,3,4,5,6. Therefore, water pollution takes the priority worldwide over other issues because of the continual discharge of industrial waste containing different organic dyes, agrochemicals, personal care products, pharmaceuticals products, and aromatic compounds into water bodies^7,8,9,10,11. Such detrimental contaminants constitute a major threat to the environment and human health owing to their low biodegradability, stability, and high toxicity^{12,13,14,15,16,17,18}.

Several biological, physical, and chemical treatment techniques have been adopted to remove these wastes from water bodies. Biological processes such as bioreactors, biologically activated sludge, microbiological treatments, and lagoons possess many positive aspects including getting rid of biodegradable organic matter along with a significant capability of removing biological oxygen demand, and suspended solids^19,20. On the other hand, they reveal several drawbacks such as slow rate of degradation, low biodegradability of dyes, and the possible formation of biological sludge^21,22. While physical processes include membrane filtration, which is remarked as quite a straightforward process and convenient method, yet it is of a high operational cost^23,24. Furthermore, electrodialysis is characterized by its high separation efficiency, but it is expensive because of the energy requirement^25,26. Moreover, chemical treatments such as coagulation-flocculation depend on adding chemicals to convert suspended solid into sludge. Despite being a cost-effective technique, it results in the formation of secondary pollutants, requiring a further treatment^27,28,29. Therefore, it could be concluded that these techniques are not capable of completely removing the organic residuals from wastewater^{30,31,32,33,34,35}.

Over the past few decades, advanced oxidation processes (AOPs) have acquired a noticeable attention by virtue of their potent oxidizing ability, rapid reaction rate, simplicity, great efficiency, and extensive applications for removing the organic contaminants^{36,37,38,39,40,41}. Simply, AOPs exploit the high reactive oxygen species (ROS) such as ${{{\mathrm{SO}}}}_4^{ \bullet - }$ and $^ \bullet \!{{{\mathrm{OH}}}}$ radicals to degrade the organic residuals into less hazardous compounds or completely mineralize them into carbon dioxide (CO₂) and water vapor (H₂O)^42,43. Nonetheless, $^ \bullet \!{{{\mathrm{OH}}}}$-based AOPs comprising photo-Fenton, electro-Fenton, photo-sono-Fenton, photo-electro-Fenton, and sono-Fenton suffer some limitations including the necessary acidic conditions (pH 2–4) and the instability of hydrogen peroxide^44,45,46. In the mid-2000′ s the progression of research resulted in arousing a sub-category of AOPs which is the sulfate radical-based advanced oxidation processes (SR-AOPs)⁴⁷.

Recently, SR-AOPs have been considered one of the most effective methods in the degradation of organic residuals owing to their potent oxidizing ability and high efficiency. Besides, ${{{\mathrm{SO}}}}_4^{ \bullet\! - }$ in comparison with $^ \bullet \!{{{\mathrm{OH}}}}$ affords diverse merits. For instance, ${{{\mathrm{SO}}}}_4^{ \bullet \!- }$ acquires higher redox potential (E_o = 2.5–3.1 V) than $^ \bullet \!{{{\mathrm{OH}}}}$ (E_o = 2.8 V). Interestingly, ${{{\mathrm{SO}}}}_4^{ \bullet\! - }$ possesses magnificent oxidation selectivity and the ability to react with distinct organic compounds with a superior rate constant (range from 10⁶ to 10⁹ M⁻¹ s⁻¹) and a wide functioning pH range⁴⁸. In addition, the longer half-lifetime of ${{{\mathrm{SO}}}}_4^{ \bullet \!- }$ (30–40 μs) compared to $^ \bullet \! {\rm{OH}}$(10^{− 3} μs) is considered another advantage of ${{{\mathrm{SO}}}}_4^{ \bullet \!- }$⁴⁹. For producing ${{{\mathrm{SO}}}}_4^{ \bullet \!- }$, a convenient oxidant is energized such as PMS that has attained the researchers’ interest owing to its remarkable features; i) PMS can be obtained in a solid form that could be easily carried, stored and transported; ii) PMS can generate a high yield of ROS; iii) Numerous materials could be employed in the activation of PMS⁵⁰.

Ghanbari et al.⁵¹ and Wang et al.⁵² reported that the direct reaction of PMS on its own, even though being a thermodynamically potent oxidant, is sluggish when used with most of the toxic pollutants. Thus, an appropriate method is required for PMS activation and generation of sulfate radicals. PMS activation can be done by applying different approaches like ultrasound (US), ultraviolet (UV), heat, microwave, alkali, and transition metals. However, the activation using transition metals is more favorable as it is more economical in comparison with energy-based techniques such as UV, US, microwave, and heat⁵³. Activation by alkali extends the duration of the treatment process as it needs pH adjustment before and after treatment, rendering the process more complex. Therefore, homogenous and heterogeneous catalysis have been employed to counteract the aforementioned limitations for PMS activation where the catalysts are either dissolved metal ions or solids, respectively⁵⁴.

Notable homogenous catalysis exhibits rapid reaction and higher efficiency particularly at pH < 7. However, it causes secondary pollution through the leaching of their component metal ions into water, resulting in the difficulty of recovering the applied catalyst from treated water and raising the overall cost of the treatment process. Additionally, it could be resulting in detrimental human health effects. On the other hand, heterogeneous catalysis offers a lot of advantages including the catalyst’s easy recovery, durability in various conditions^47,55, applicability over a wide pH range (2–9), and also it does not require a secondary treatment⁵⁶.

The activation of PMS by transition metals is represented in Eqs. (1–4). Initially, the catalyst activates PMS, generating ROS such as ${{{\mathrm{SO}}}}_4^{ \bullet \!- }$ and $^ \bullet \! {\rm{OH}}$ radicals, in addition to ¹O₂ engenders from the electron-transfer mechanism. The transition metal ions (Mⁿ⁺) initiate the reaction through generation ${{{\mathrm{SO}}}}_4^{ \bullet \!- }$, resulting in the oxidation of the metal to Mⁿ⁺¹. Interestingly, the weak redox potential of peroxymonosulfate radical (${{{\mathrm{HSO}}}}_5^ -$/${{{\mathrm{SO}}}}_5^{ \bullet\! - }$) (E_o = 1.1 V) facilitates the recovery of Mⁿ⁺¹ via the electron transfer. Moreover, the presence of two or more convenient transition metals leads to the synergistic enhancement of PMS activation^57,58.

$${{{\mathrm{M}}}}^{2 + } + {{{\mathrm{HSO}}}}_5^ - \to {{{\mathrm{SO}}}}_4^{ \bullet \!- } + {{{\mathrm{M}}}}^{3 + }$$

(1)

$${{{\mathrm{M}}}}^{3 + } + {{{\mathrm{HSO}}}}_5^ - \to {{{\mathrm{SO}}}}_5^{ \bullet \!- } + {{{\mathrm{M}}}}^{2 + }$$

(2)

$${{{\mathrm{SO}}}}_4^{ \bullet \!- } + {{{\mathrm{H}}}}_2{{{\mathrm{O}}}} \to {{{\mathrm{SO}}}}_4^{2 - } + ^ \bullet\!\!{{{\mathrm{OH}}}} + {{{\mathrm{H}}}}^ +$$

(3)

$$2{{{\mathrm{SO}}}}_5^{ \bullet \!- } \to 2{{{\mathrm{SO}}}}_4^{ \bullet \!- } + ^1\!\!{{{\mathrm{O}}}}_2\,{{{\mathrm{or}}}}\,{{{\mathrm{S}}}}_2{{{\mathrm{O}}}}_8^{2 - } + ^1\!\!{{{\mathrm{O}}}}_2$$

(4)

LDH is an ionic clay two-dimensional (2D) material with a hydrotalcite-like structure, having an individual ion-exchange characteristic owing to its anionic interlayer and cationic metal layers⁵⁹. The structure of LDHs consists of divalent and trivalent cations that are neutralized by the commutable anions intercalated into the interlayer⁶⁰. LDHs general formula is exemplified as [M_1−x²⁺ M_x³⁺ (OH)₂]^x+(Aⁿ⁻)_x/n • mH₂O, in which M²⁺ symbolizes the divalent metal cations while M³⁺ trivalent ones, Aⁿ⁻ shows the exchangeable anions among the layers (such as carbonates or nitrates), and x represents the molar ratio between metal ions (Fig. 1a)⁶¹. Moreover, LDHs are characterized by their outstanding chemical and physical properties, as well as their facile fabrication and cost-effectiveness⁶². Consequently, LDHs are applied for bountiful applications such as catalysis, batteries, cosmetics, adsorption, sensors, etc.^63,64,65. In particular, the abundant choices of M²⁺/M³⁺ render LDHs as efficient heterogeneous catalysts possessing a high PMS activation capacity for degrading toxic organic residuals from water bodies^66,67. Furthermore, the decomposition of LDHs via calcination from 400 to 600 °C forms mixed metal-oxides known as layered double oxides (LDOs). Interestingly, LDOs possess several advantages including extraordinary surface area, eco-friendly nature, easy preparation, and light-sensitive catalytic activity. In addition, LDOs have a higher catalytic activity in AOPs compared to LDHs due to their higher stability⁶⁸. Table 1 summarized a comparison between LDHs and LDOs.

Fig. 1: Chemical structure of LDHs and their common preparation methods and applications.

a The structure and bountiful applications of LDHs, and b Schematic illustration of the three superior preparation routes of LDHs.

Table 1 A comparison between the main characteristics of LDHs and LDOs.

LDHs have exhibited a promising performance in AOPs, which is most likely due to (a) LDHs pH-buffering extent acquiring the ability to retain the neutral pH levels; (b) LDH-based catalysts possess a steady structure that can boost the reusability and limits the metals’ leaching; (c) the simplicity of synthesizing LDHs and the capability of immobilizing distinct metal ions can enhance the catalytic functionality; (d) the presence of assorted ROS specifically ¹O₂ can diminish the negative influence of coexisting substrates as inorganic anions; (e) several LDH-based nanocomposites can be composed by paring LDHs or LDOs along with other materials such as carbonaceous and magnetic materials to enhance the catalytic performance, conductibility, and stability⁶⁹.

Various review articles have recently highlighted the synthesis of LDH-based catalysts and their applications in wastewater treatment. For instance, Yang et al.⁷⁰ investigated the controlling factors that affect the photodegradation of dyes using LDHs in 2016. Wu et al.⁷¹ has focused recently on the progression of the photocatalytic activities of LDH-based catalysts, in terms of energy and structural fabrication. In 2021, Kohantorabi et al.⁷² addressed the catalytic activity of many heterogeneous catalysts including metal/bimetallic oxides, supported noble/non-noble metals, carbonaceous-based materials, LDHs, metal-organic frameworks (MOFs), and zeolites towards the PMS activation. Moreover, Xie et al. highlighted the preparation and characterization approaches of LDHs as well as their application in AOPs⁶⁹.

Based on the authors’ knowledge, the PMS activation by LDH for the degradation of organic contaminants was briefly discussed as one of the AOPs. However, the recent increase in research papers that addressed this hot spot topic needs a thorough and up-to-date assessment. In this connection, our review aimed to summarize the recent progress in the degradation of common organic pollutants including aromatic compounds, pharmaceutical residues, and dyes via the PMS activation using LDH-based catalysts. Importantly, the review focused on the various preparation methods of LDH and the controlling factors on the quality and quantity of the produced LDH. Furthermore, the degradation mechanism pathways were discussed based on quenching tests and advanced analyses like X-ray absorption fine structure (XAFS) spectroscopy, X-ray photoelectron spectroscopy (XPS), electron spin resonance (ESR) spectroscopy, etc. Notably, the durability of LDH-based catalysts during the PMS activation was evaluated from leaching and reusability tests. Ultimately, in light of our in-depth investigations, we concluded the progress, challenges, and recommend ations for inspiring researchers to further developments.

Synthesis routes of LDHs

The LDH fabrication is fulfilled by the combination of a considerable range of divalent and trivalent cations with diverse interlayer anions. In preparation of LDHs, many factors must be taken into consideration, for instance, the sort of metallic cations and interlayer anion, the M²⁺/M³⁺ ratio, pH condition, route of assemblage, and atmosphere control. Notably, such parameters control the main characteristic of LDHs including surface area, crystallinity, morphology, and particle size. Furthermore, it is greatly noticed that the various physical features of the fabricated LDHs make them possess propitious abilities to be utilized in industrial fields⁷³. In the last few years, environmentally safe and cost-effective preparation techniques have been fostered including; co-precipitation, hydrothermal/solvothermal, urea hydrolysis, sol-gel, ion exchange, calcination-rehydration, etc.

Co-precipitation route

The co-precipitation method is one of the simplest, cost-effective, and most efficient techniques for fabricating of LDHs^74,75. In addition, such an approach provides a direct preparation of LDHs from the diversified M²⁺ and M³⁺ and a wide scale of anions from inorganic anions (viz., carbonate (${{{\mathrm{CO}}}}_3^{2 - }$), nitrate (${{{\mathrm{NO}}}}_3^ -$) and chloride (Cl⁻)) to organic molecules and biomolecules⁷⁶. The fabrication of LDHs via co-precipitation approaches involves the addition of anion solution into a solution from a proper ratio of M²⁺/M³⁺, then adding a base or urea hydrolysis for raising the solution pH (from 8 to 10) and prompting the precipitation of LDH⁷⁷. Ultimately, the reaction mixture is stirred for one day at room temperature, filtered, washed with deionized water several times, and dried using vacuum or conical funnel⁷⁵. While for attaining LDO, the synthesized LDHs are calcined at moderate temperature (from 450 to 600 °C), endowing them with plenty of active sites, high surface area, and uniform distributed composition⁷³. On the other hand, the calcination temperature is not permitted to be too high since mixed metal oxides could be sintered and transformed into spinel phase, and diminishing both of the pore volume and surface area⁷⁸.

Hydrothermal route

Unlike the co-precipitation route, LDHs’ synthesis via the hydrothermal route is an ideal route for the rapid procedure with a broad-scale industrialization potentiality. It is more commonly operated in autoclaves, generating autogenous pressure due to heating at temperatures ranging from 60 to 200 ^oC^79,80,81. In this method, an alkali solution encompassing the metal salts that will compose the LDH is stirred well before pouring into an autoclave and the solution is maintained at hydrothermal conditions for a specific duration that could vary between hours to many days^82,83. Then it is left for crystallization at room temperature, washing, and subsequently dried⁸⁴. A higher yield of the hydrothermally fabricated LHDs becomes conspicuous by augmenting the temperature and the time of the synthesis process. Interestingly, the hydrothermal approach possesses special merits such as producing LDHs with higher crystalline morphology compared to that fabricated by the co-precipitation method as well as averting the toxic residuals that could be negatively influencing humans and the environment (i.e. Cl⁻, OH⁻, and ${{{\mathrm{NO}}}}_3^ -$)^75,85.

Urea hydrolysis

Urea has been extensively applied as a precipitating agent for a plethora of metals into their metal hydroxides analogous as it is a weak Brønsted base (pK_b = 13.8). Thereby, the resultant slow hydrolysis rate could control the degree of supersaturation and precipitation. In general, the urea hydrolysis takes place via two reaction phases: (1) the yielding of ammonium cyanate (NH₄CNO) which is considered as a rate-determining step (Eq. 1); and (2) the fast hydrolysis of cyanate to form carbonate (Eq. 2)⁸⁶.

$${{{\mathrm{CO}}}}\left( {{{{\mathrm{NH}}}}_2} \right)_2 \to {{{\mathrm{NH}}}}_4{{{\mathrm{CNO}}}}$$

(5)

$${{{\mathrm{NH}}}}_4{{{\mathrm{CNO}}}} + 2{{{\mathrm{H}}}}_2{{{\mathrm{O}}}} \to 2{{{\mathrm{NH}}}}_4^ + + {{{\mathrm{CO}}}}_3^{2-}$$

(6)

Moreover, the rate of urea hydrolysis can be simply dominated by adjusting the reaction temperature, where it is noticeable that the rate constant is boosted 200 times by raising the process temperature from 60 to 100 °C⁸⁴. Notably, the urea hydrolysis route is distinguished by the easiness of rinsing the precipitate rather than the co-precipitation that needs reiterated rinsing to eliminate the alkali-metal ions⁸⁷. However, the urea hydrolysis approach suffers a major shortcoming, which is the presence of carbonate anions in the fabricated LDHs⁸⁸.

Anion-exchange route

The anion-exchange route is an indirect synthesis route as the used precursor in this technique is the LDH that has robust anion-exchange characteristics. The facility of monovalent ions exchange is arranged as follows: ${{{\mathrm{NO}}}}_3^ -$ > Br⁻ > Cl⁻ > F⁻ > OH⁻. Therefore, it is obvious that nitrate ions are displaced quite easier, hence they are the utmost proper precursors for the fabrication of LDHs via anion-exchange route⁸⁹. Furthermore, this technique involves the synthesis of LDHs with host anions using a co-precipitation approach, then exchanging with the guest anions. This approach is only applied when the co-precipitation approach is improper, like the unstable di- or tri-valent metal cations or anions in basic conditions⁹⁰. In general, the mechanism of the anion-exchange reaction can be expressed as follows:

$${{{\mathrm{LDH}}}}.{{{\mathrm{A}}}}^{{{{\mathrm{m}}}} - } + {{{\mathrm{X}}}}^{{{{\mathrm{n}}}} - } \to {{{\mathrm{LDH}}}}.{{{\mathrm{X}}}}^{{{{\mathrm{n}}}} - } + {{{\mathrm{A}}}}^{{{{\mathrm{m}}}} - }$$

(7)

Where A is the host anions while X expresses as the guest having negative charges of m and n, sequentially⁸⁵.

It is worth noting that the anion-exchange process is primarily governed by the electrostatic interaction between the brucite sheets and the exchanging anions. Moreover, the anion-exchange approach has strict conditions⁷⁷;

• (a) Depending on the chemical composition of the host LDH layers.

• (b) The convenient solvent fosters the anion-exchange route.

• (c) The pH medium should be higher than 4.0 since LDH could be destroyed at a lower pH level.

• (d) Anion-exchange requires high temperatures.

On the other hand, anion-exchange has some drawbacks, including (a) that for the anion-exchange reaction, the LDH precursor should be obtained first by another approach. Therefore, two steps are needed to reach the final product, which results in increasing the process’s complexity and cost. (b) Neutral species are not able to be intercalated into LDH through this method⁹¹.

Reconstruction (rehydration) route

The reconstruction route primarily depends on making use of the memory effect of LDHs, where calcination of the LDHs is accomplished to detach the interlayer hydroxyl groups, anions, and water, producing mixed metal oxides⁹². Then, the produced mixed metal oxides (MMOs) are rehydrated through their dispersion in the target anions solution, ensuing the reestablishment of LDHs while the anionic species are inserted in the interlayer part of LDHs⁹³. The temperature of calcination is the key parameter in this route since the increment in the calcination temperature directly diminishes the memory effect⁹⁴. It is noteworthy to mention that the LDHs’ calcination temperature should not be over 550; the elevated temperatures will collapse the LDH structure. Nevertheless, the low calcination temperature affects the crystal phase⁹⁵.

The preeminent merits of the reconstruction route are summarized in the following points;

• 1. Both anions and neutral species can be incorporated into LDH.

• 2. Due to the possibility to eliminate the incorporated anions within the calcination operation, thus LDH incorporated with any sort of anions can be exploited as the original reactant.

On the other hand, the reconstruction route suffers some drawbacks such as;

• (a) The recovery of LDH essentially relies on the calcined metal-oxides activity in the solutions of the reconstruction.

• (b) Time-consuming and laborious since it requires two steps or more of the entire process encompassing calcination in addition to reconstruction⁹¹.

Sol-gel route

Sol-gel approach has earned a great reputation owing to its cost-effectiveness and ability to produce LDHs with a superior purity⁹⁶. Generally speaking, a number of variables should be taken into account when fabricating LDHs using the sol-gel method, including adjusting the temperature, changing the composition of the host anions, and adding or removing the species of reactant. Additionally, the fabrication process of the sol-gel route proceeds as follows: the components are continuously stirred to create sol, and the reaction mixture is aged to allow for the phase transition from sol to gel⁹⁷. The solvent, pH, concentration of the metal precursors, and synthesis temperature are some of the parameters that can be adjusted to control the characteristics of the fabricated LDHs using this method. These parameters also affect the rate of hydrolysis and condensation of the metal precursor⁹⁸. LDHs produced by the sol-gel procedures have higher thermal stability and lower crystallinity compared to those formed by the co-precipitation route⁹³.

Overall, several studies concluded that co-precipitation, anion exchange, and reconstruction were superior to other LDH preparation methods. Schematic representations of these LDH synthesis methods are shown in Fig. 1b. It was determined by contrasting the three preparation techniques that co-precipitation stands out for its high yield, which can be up to three times that of other LDH preparation techniques. While anion exchange produces LDHs with a larger anion size, endowing them with further advantages in several pharmaceutical applications. Furthermore, reconstruction produces poor crystalline and partially intercalated LDHs⁹⁹.

Environmental applications of PMS–based LDH systems

Degradation of aromatic compounds

Fazli et al.¹⁰⁰ reported the photocatalytic activity of Fe₃O₄@CuCr-LDH nanocomposite towards the degradation of diethyl phthalate (DEP) through PMS activation. As shown in Fig. 2a, the fabrication of CuCr-LDH was successful because even in the purest CuCr-LDH or magnetic composite, the associated metal hydroxide (M-OH) absorption peaks were clearly visible below 1000 cm⁻¹. Since pH has the greatest influence on the PMS activation process, its impact was studied over a wide pH range from 3 to 12 (Fig. 2b). The optimal pH was determined to be 8 because there is plenty of H⁺ in a highly acidic medium that forms H-bonds with the peroxy bond of ${{{\mathrm{HSO}}}}_5^{2 - }$. Zeta potential measurements revealed that the pHPZC of Fe3O4@CuCr-LDH was 8.9, indicating that the surface was positively charged in both neutral and acidic media. Thence, the H-attached ${{{\mathrm{HSO}}}}_5^{2 - }$ could not reach to the positive surface of Fe₃O₄@CuCr-LDH, so the catalytic degradation of DEP was directly declined. Additionally, Jawad et al.‘s¹⁰¹ study looked into how high acidic medium (4.32 mg/L at pH 3) caused the leaching of Cu ions in CuOx@Co-LDH to peak, hindering the phenol’s degradation. Moreover, there was a significant diminution in the DEP degradation efficiency at a high basic medium, which may be ascribed to the interaction between PMS and ${{{\mathrm{SO}}}}_5^{2 - }$. As a result, ¹O₂ is formed, which possesses a lower oxidizing potential than ${{{\mathrm{SO}}}}_4^{ \bullet - }$ and $^ \bullet \!{{{\mathrm{OH}}}}$ radicals. On the other hand, the PMS concentration strongly controlled the degradation aptitude of DEP at which a synchronic amelioration in the DEP degradation efficiency was observed with raising the PMS concentration from 1 to 14 mM (Fig. 2c). This finding may be anticipated by the abundance of the reactive ${{{\mathrm{SO}}}}_4^{ \bullet - }$. Nonetheless, the increase in the PMS concentration from 8 to 14 mM caused a slight enhancement in the DEP, which is concomitant with sang et al. study¹⁰². This common behavior may be assigned to the self-quenching reaction between ${{{\mathrm{SO}}}}_4^{ \bullet - }$ and the extra PMS^103,104.

Fig. 2: Characterization of the chemical structure of LDHs-based composites and their participated metal in the PMS activation.

a Fourier transform infrared (FTIR) spectra, b Effect of pH and c Effect of PMS concentration on the degradation efficiency of DEP adapted with permission from ref. ¹⁰⁰, Copyright, 2021, Elsevier. d The schematic diagram of the PMS activation by BC-LDH for the removal of DEP, e Normalized Co K-edge, and f Ni K-edge X-ray absorption near edge structure (XANES) spectra of BC-LDH adapted with permission from ref. ¹⁰⁵, Copyright, 2022, Elsevier.

In another attempt, Yang et al.¹⁰⁵ reported a quite smart and economical approach to stabilize the adsorbed Co²⁺ and Ni²⁺ onto a biochar (BC) surface to fabricate BC-LDH composite, which was utilized for the catalytic degradation of DEP (Fig. 2d). BC-LDH showed enhanced catalytic behavior towards the PMS activation since the degradation efficiency of DEP reached 90% within 240 min. XPS spectra of BC-LDH before and after the catalytic degradation of DEP were used to interpret the degradation mechanism. XPS spectrum of Co2p after the degradation reaction revealed a slight decrease in the Co²⁺ proportion from 50.1 to 49.8%, accompanied by an increase in the Co³⁺ proportion from 48.9 to 50.2%. This observation was accredited to the participation of Co²⁺ in the PMS activation, resulting in partial oxidation of Co²⁺ to Co³⁺, which was in line with Li et al. observation¹⁰⁶. Furthermore, the C‒C proportion in the C1s spectrum increased from 21.8 to 29.3% compared to pure BC, indicating the oxidation of the aromatic moieties (C=C→C-C) during DEP oxidation¹⁰⁷. Moreover, Co and Ni K-edge XAFS spectroscopy of BC-LDH (Fig. 2e, f) revealed a slight shift of Co K-edge to the high energy after DEP degradation, reflecting the partial oxidation of Co²⁺ to Co³⁺. Contrariwise, there was no change in the Ni K-edge of BC-LDH after the catalytic reaction, signifying that Ni did not participate in the reaction.

In another investigation, Zhao et al.¹⁰⁸ highlighted the activation of PMS by Fe-Al LDH to efficiently degrade Bisphenol A (BPA). Notably, Mӧssbauer spectroscopy was used to explain the surface transformation of Fe-Al LDH after the catalytic reaction. It was observed from the Mӧssbauer spectrum of Fe-Al LDH that there were two species of Fe²⁺ (Fe–O–Fe and Fe-O-Al)^109,110. Unequivocally, the electronegativity of Fe²⁺ in Fe–O–Fe is lower than Fe-O-Al owing to the higher electron-withdrawing property of Fe than Al. Furthermore, the proportions of Fe–O–Fe and Fe-O-Al in Fe-Al LDH were 15% and 62%, respectively, with a small proportion of Fe³⁺ that may be due to the sample oxidation throughout the measurement. The Mӧssbauer spectrum of the transformed Fe-Al LDH after the BPA degradation showed that the proportion of Fe²⁺ in both species decreased with an increase in the formed Fe³⁺. This finding designated the participation of Fe²⁺ in both reactive species which may be attributed to the good conductivity of the whole metal layer. Moreover, to scrutinize the impact of time on the active species during the activation of PMS, Fe-Al LDH/PMS was examined by an in-situ electron paramagnetic resonance (EPR) test. It was observed that there were three active species ($^ \bullet \!{{{\mathrm{OH}}}}$, ${{{\mathrm{SO}}}}_4^{ \bullet \!- }$ and superoxide radical (${{{\mathrm{O}}}}_2^{ \bullet \!- }$)) and the distinctive signal peaks of $^ \bullet \! {\rm{OH}}$ were the dominant compared to ${{{\mathrm{SO}}}}_4^{ \bullet \!- }$ signals over time. In addition, the intensity of both active species increased with time and reached its peak at around 5 min. Then, the signals of $^ \bullet \!{{{\mathrm{OH}}}}$ and ${{{\mathrm{SO}}}}_4^{ \bullet \!- }$ disappeared after 20 min, but the ${{{\mathrm{O}}}}_2^{ \bullet \!- }$ signals were almost still with their high intensity. On the contrary, the high intensity of $^ \bullet \! {\rm{OH}}$ and ${{{\mathrm{SO}}}}_4^{ \bullet \!- }$ signals was observed for a long time in some catalyst/PMS systems, indicating that PMS could be slowly and continuously activated¹¹¹. Besides, a quenching test was conducted to define the dominant reactive radical in the BPA degradation. Several scavengers were utilized including methanol (MeOH) for $^ \bullet \!{{{\mathrm{OH}}}}$ and ${{{\mathrm{SO}}}}_4^{ \bullet\! - }$, isopropanol for $^ \bullet \!{{{\mathrm{OH}}}}$, p-benzoquinone (p-BQ) for ${{{\mathrm{O}}}}_2^{ \bullet\! - }$ and sodium azide (NaN₃) for ¹O₂. The results clarified that there was a decline in the degradation efficiency in the presence of methanol, reflecting the participation of both $^ \bullet \!{{{\mathrm{OH}}}}$ and ${{{\mathrm{SO}}}}_4^{ \bullet \!- }$ in the BPA degradation. While ${{{\mathrm{O}}}}_2^{ \bullet\! - }$ is the controlling reactive radical since the BPA degradation significantly dwindled in the presence of p-BQ. Consequently, the radical pathway of the BPA degradation was suggested based on these results and it was further confirmed when the quenching of ¹O₂ slightly decreased the degradation efficiency.

Continuing with BPA removal studies, Shahazad et al. successfully synthesized a hexagonal crystalline Cu-rGO LDH nanohybrid for activating PMS to degrade BPA. The scanning electron microscope (SEM) image (Fig. 3a) revealed the hexagonal morphology of Cu-CO₃ LDH. While Fig. 3b indicated that the hexagonal structure remained intact after the ion exchange and the Cu-NO₃ LDH formation. The Tyndall light scattering effect showed the dispersion of the exfoliated graphene oxide (GO) and Cu-LDH nanosheets in dimethyl formamide and formamide, respectively, as shown in Fig. 3c. The high-resolution transmission electron microscopy (HRTEM) of Cu-rGO LDH (Fig. 3d, e) showed the periodic assembly of reduced graphene oxide (rGO) and Cu-LDH. In addition, the selected area electron diffraction (SAED) of Cu-rGO LDH (Fig. 3f) revealed that the interlayer spacing between two LDH layers was 1.03 nm. The SAED of Cu-rGO LDH (Fig. 3g) inferred the 2D structure of both Cu-LDH and rGO, implying the 2D order of Cu-rGO LDH. Figure 3h elucidates the different pathways of the ¹O₂ formation. Furthermore, EPR (Fig. 3i, j) and chemical quenching tests confirmed that the BPA degradation followed the non-radical pathway through the PMS activation by Cu-rGO LDH. Also, it was proposed that inner-sphere complexation between Cu-rGO LDH and PMS may be responsible for boosting the PMS decomposition. The plausible mechanism of generating ¹O₂ is clarified in Eqs. 7–11. Besides, it was found that the PMS concentration declined after the interaction with Cu-rGO LDH in the absence of BPA. This observation signified the strong bond between PMS and Cu-rGO LDH, forming a complex that is responsible for the formation of ¹O₂ through ${{{\mathrm{O}}}}_2^{ \bullet \!- }$ (Eqs. 11–13). Such a finding was in line with Shahzad et al.¹¹² study that suggested the phenol degradation through the PMS activation by Fe-rGO LDH followed a ternary system as the electrons were transferred from phenol (donor) to PMS (acceptor) via Fe-rGO LDH (mediator). Subsequently, the formation of non-radical ROS agreed with the quenching test that inferred the domination of ¹O₂.

$${{{\mathrm{HSO}}}}_5^ - \to {{{\mathrm{H}}}}^ + + \,^ \bullet \!{{{\mathrm{SO}}}}_5^ - + {{{\mathrm{e}}}}$$

(7)

$$\,^ \bullet \!{{{\mathrm{SO}}}}_5^ - + \,^ \bullet \!{{{\mathrm{SO}}}}_5^ - \to 2{{{\mathrm{SO}}}}_4^{2 - } + \,^1\!{{{\mathrm{O}}}}_2$$

(8)

$$\left[ {{{{\mathrm{Cu}}}}^{2 + } {\hbox{-}} {{{\mathrm{OH}}}}} \right]^ + + {{{\mathrm{H}}}} {\hbox{-}} {{{\mathrm{O}}}} {\hbox{-}} {{{\mathrm{O}}}} {\hbox{-}} {{{\mathrm{SO}}}}_3^ - \to [{{{\mathrm{Cu}}}}^{2 + } {\hbox{-}} {{{\mathrm{O}}}} {\hbox{-}} {{{\mathrm{O}}}} {\hbox{-}} {{{\mathrm{SO}}}}_3^ - ] + {{{\mathrm{H}}}}_2{{{\mathrm{O}}}}$$

(9)

$$2[{{{\mathrm{Cu}}}}^{2 + } {\hbox{-}} {{{\mathrm{O}}}} {\hbox{-}} {{{\mathrm{O}}}} {\hbox{-}} {{{\mathrm{SO}}}}_3^ - ] + 3\,{{{\mathrm{H}}}}_2{{{\mathrm{O}}}} + {{{\mathrm{HSO}}}}_5^ - \to 2\left[ {{{{\mathrm{Cu}}}}^ + {\hbox{-}} {{{\mathrm{OH}}}}} \right] + 3{{{\mathrm{SO}}}}_4^{2 - } + 2{{{\mathrm{O}}}}_2^{ \bullet \!- } + 5{{{\mathrm{H}}}}^ +$$

(10)

$$[{{{\mathrm{Cu}}}}^{2 + } {\hbox{-}} {{{\mathrm{O}}}}{\hbox{-}}{{{\mathrm{O}}}} {\hbox{-}} {{{\mathrm{SO}}}}_3^ - ] + {{{\mathrm{O}}}}_2^{ \bullet \!- } + {{{\mathrm{OH}}}}^ - \to \left[ {{{{\mathrm{Cu}}}}^ + {\hbox{-}} {{{\mathrm{OH}}}}} \right] + {{{\mathrm{SO}}}}_4^{2 - } + 3^1{{{\mathrm{O}}}}_2$$

(11)

$$2{{{\mathrm{O}}}}_2^{ \bullet\! - } + 2{{{\mathrm{H}}}}_2{{{\mathrm{O}}}} \to 2^1{{{\mathrm{O}}}}_2 + {{{\mathrm{H}}}}_2{{{\mathrm{O}}}}_2 + 2{{{\mathrm{OH}}}}^ -$$

(12)

$$\,^1{{{\mathrm{O}}}}_2 + {{{\mathrm{BPA}}}} \to {{{\mathrm{Degradation}}}}$$

(13)

Fig. 3: Morphology characterization of LDHs and identification of the controlled ROS on the PMS activation.

a, b SEM images of Cu-rGO LDH, c Tyndall light scattering effect, d, e TEM of Cu-rGO LDH, f, g SAED of Cu-rGO LDH, h the mechanism of ¹O₂ formation in the Cu-rGO LDH/PMS system, i, j ESR spectra of TMP-¹O₂ and DMPO-${{{\mathrm{O}}}}_2^{ \bullet \!- }$ and k, l Fe Mössbauer spectra FeAl-LDH and DOM-LDH, and m the proposed degradation mechanism adapted with permission from ref. ¹¹⁵ Copyright, 2020, Elsevier.

Notably, the inner-sphere complexation was confirmed by recording a red shift in the distinguishing peak of PMS using the attenuated total Reflectance-Fourier transform infrared spectroscopy (ATR-FTIR) spectrum after the interaction between Cu-rGO LDH and PMS. Such a result was concomitant with Yang et al.¹¹³. Furthermore, the OH groups might be replaced when PMS bonded with the catalyst active sites during the complexation. In addition, the different pathways of the ¹O₂ generation could be attributed to the uncharged oxygen in ${{{\mathrm{HSO}}}}_5^ -$ involving in the PMS/Cu-rGO LDH inner-sphere complexation¹¹⁴.

In that respect, Ye et al.¹¹⁵ improved the degradation efficiency of BPA from 60% to 93% within 60 min using FeAl-LDH after the introduction of dissolved organic matter (DOM) into the reaction¹¹⁶. The SEM images displayed the platelets-like structure of FeAl-LDH with a size range of 20–60 nm, which tended to aggregate and disperse randomly, as it is typical of LDHs¹¹⁷. After the introduction of DOM, the morphology changed into a flower-like morphology with a diameter of 50 nm and a thickness of 3 nm. Furthermore, a significant enhancement in the specific surface area, pore width, and pore volume of DOM-LDH (70.67 m² g⁻¹, 11.84 nm, and 0.181 cm³ g⁻¹) was observed compared to the pure FeAl-LDH (56.18 m² g⁻¹, 1.725 nm, and 0.069 cm³ g⁻¹), suggesting a better interaction between DOM-LDH, PMS, and BPA. Mössbauer spectrum of FeAl-LDH (Fig. 3k) represented the trans-coordinated octahedral Fe (II) species with 32.9% and both cis and trans-coordinated octahedral Fe (III) with 67.1%, whereas the spectrum of DOM-LDH (Fig. 3l) demonstrated trans-coordinated octahedral Fe (II) species with 61.2% and both cis and trans-coordinated octahedral Fe (III) with 19.2%. This result denoted the rapid conversion of Fe (III) to Fe (II) via the reduction by humic acid-like compound, which accelerated the PMS activation. Therefore, the superior catalytic degradation efficacy was achieved as elucidated in Fig. 3m.

In another study, Hou et al.¹¹⁸ highlighted the PMS activation by CuFeAl-LDO to degrade tetrabromobisphenol A (TBBPA). CuFeAl-LDO was fabricated by the calcination of CuFeAl-LDH at 600 °C. SEM images revealed a significant change in the morphology of CuFeAl-LDH after the calcination at which it has nanosheet morphology, while CuFeAl-LDO particles are globular. The experimental results implied the efficient degradation behavior of CuFeAl-LDO since the degradation rate of TBBPA reached 99.91% within 60 min using 0.1 g CuFeAl-LDO, 0.5 mM PMS, 15 mg/L TBBPA, and pH 8.5. Additionally, the successful regeneration of CuFeAl-LDO was confirmed as the degradation efficiency of TBBPA decreased around 13% after the 5^th cycle. It is worth mentioning that this decrease in the catalytic performance of CuFeAl-LDO could be attributed to the loss or aggregation of the catalyst during the reaction. In addition, the possibility of producing an intermediate that could be adsorbed on the catalyst surface and cover the active sites resulted in the deactivation of CuFeAl-LDO^119,120. Moreover, the leaching of Fe and Cu ions from CuFeAl-LDO was detected by inductively coupled plasma mass spectrometry (ICP-MS). It was found that the leaching amount of Fe and Cu were 0.003 and 0.064 mg/L, respectively, which are lower than the allowed limits by the United States Environmental Protection Agency (USEPA).

In this perspective, Lu et al.¹²¹ fabricated CuCoFe-LDH to activate PMS for the degradation of nitrobenzene (NB). SEM image displayed the stacked thin slices of CuCoFe-LDH, while the SAED pattern revealed the planes (012) and (113) where d (012) = 0.2629 nm and d(113) = 0.1528 nm. It was found that the NB degradation by CuCoFe-LDH/PMS generated intermediate products at which ultra-performance liquid chromatography-time-of-flight (UPLC-TOF) analysis clarified the formation of hydroxyl-NB. Furthermore, time-of-flight mass spectrometry (TOF-MS) depicted the extremely low intensity of the NB signal and the presence of propanedioic acid (negative ion mode)¹²². Also, the presence of these intermediates was confirmed by the continuous decrease in the pH of the system during the NB degradation process. Based on headspace-gas chromatography/mass spectrometry (HS-GC-MS) and UPLC-TOF analysis, it was proposed that the degradation pathway of NB by CuCoFe-LDH/PMS was driven by the substitution reactions between $^ \bullet \!{{{\mathrm{OH}}}}$ radicals and NB This suggested degradation process confirmed the domination of $^ \bullet \!{{{\mathrm{OH}}}}$ radicals for the NB degradation process^123,124. Also, Ali et al. confirmed the formation of intermediates during the degradation of para hydroxyl benzoic acid (PHB) using FeMgAl-MoS4/PMS. The degradation pathway of PHB took place through decarboxylation or dehydration reactions and 2-hydroxypropanoic acid, 2-hydroxyacedic acid, and fumaric acid were generated. Eventually, these intermediate molecular acids were converted to CO₂ and H₂O⁶⁷.

In one more attempt, Zhu et al.¹²⁵ synthesized Ni₁₅Cu-LDHs and introduced oxygen vacancies (OVs) through Cu doping, which are anionic surface defects contributing to the positively charged surfaces of catalysts. Such OVs can enhance the adsorption of reactants, shorten the electron-transfer distances, and accelerate the degradation process without releasing radicals into the liquid phase¹²⁶. Consequently, Ni₁₅Cu-LDHs, in which OVs were introduced, achieved an outstanding degradation efficacy of almost 99.07% against methyl 4-hydroxybenzoate (MeP) within just 10 minutes. Interestingly, Ye et al.⁴⁹ achieved 100% degradation efficiency of dimethyl phthalate within 60 min using BC-modified CoFe-LDH since BC played the role of catalyst carrier and electron mediator which enhanced the physicochemical characteristics of CoFe-LDH and improved the redox cycle of transition metals.

In another study, Zhao et al.¹²⁷ reported the degradation efficiency of 2,4-dichlorophenoxyacetic acid (2,4-D) using Fe₃O₄–CsxWO₃/NiAl LDH Composite. It was figured out that the presence of co-existing anions such as ${{{\mathrm{CO}}}}_3^{2 - }$, bicarbonate (${{{\mathrm{HCO}}}}_3^ -$), and Cl⁻ reduced the degradation efficiency of 2,4-D from 90.53% to 69.80% and 26.81% by increasing the concentration of ${{{\mathrm{CO}}}}_3^{2 - }$ and ${{{\mathrm{HCO}}}}_3^ -$ from 0 to 16 ppm, respectively. This result could be ascribed to the significant disrupting of ${{{\mathrm{SO}}}}_4^{ \bullet - }$ and ${\!}^ \bullet \!{{{\mathrm{OH}}}}$ by ${{{\mathrm{CO}}}}_3^{2 - }$ and ${{{\mathrm{HCO}}}}_3^ -$ to produce a weak oxidant (${{{\mathrm{CO}}}}_3^{ \bullet - }$)¹²⁸. In contrast, the degradation efficiency enhanced from 90.53% to 99.14% by increasing the Cl⁻ concentration from 0 to 16 ppm in 180 min, which could be attributed to the enhancement of the ${{{\mathrm{SO}}}}_4^{ \bullet - }$ production by the chlorine free radicals¹²⁹. The reusability of Fe₃O₄–CsxWO₃/NiAl LDH was not that good as it decreased to 75% after four consecutive cycles. On the other hand, a quite promising regeneration ability of CuCoFe-LDH (93%) was obtained after five cycles of p-nitrophenol (P-NP) degradation, signifying a high degree of CuCoFe-LDH stability¹³⁰.

Notably, the catalytic degradation mechanism of 2,4-D by Fe₃O₄-Cs_xWO₃/NiAl-LDH/PMS was proposed as follows; Fe₃O₄-Cs_xWO₃/NiAl-LDH formed electron/hole pairs (e⁻/h⁺) as the produced e⁻ at the conduction band reacted with O₂ and generated ${{{\mathrm{O}}}}_2^{ \bullet - }$ (Eq. 14). Meanwhile, h⁺ at the valence band resulted in the formation of ${\!}^ \bullet \!{{{\mathrm{OH}}}}$ when reacting with water molecules (Eq. 15).

$${{{\mathrm{Fe}}}}_3{{{\mathrm{O}}}}_4 {\hbox{-}} {{{\mathrm{Cs}}}}_{{{\mathrm{x}}}}{{{\mathrm{WO}}}}_3/{{{\mathrm{NiAl}}}} {\hbox{-}} {{{\mathrm{LDH}}}}\left( {{{{\mathrm{e}}}}_{{{{\mathrm{CB}}}}}^ - } \right) + {{{\mathrm{O}}}}_2 \to {{{\mathrm{O}}}}_2^{ \bullet \!- }$$

(14)

$${{{\mathrm{Fe}}}}_3{{{\mathrm{O}}}}_4 {\hbox{-}} {{{\mathrm{Cs}}}}_{{{\mathrm{x}}}}{{{\mathrm{WO}}}}_3/{{{\mathrm{NiAl}}}} {\hbox{-}} {{{\mathrm{LDH}}}}( {{{\mathrm{h}}}}_{{{{\mathrm{VB}}}}}^+ ) + {{{\mathrm{H}}}}_2{{{\mathrm{O}}}} \to {\,}^{\bullet} \!{{{\mathrm{OH}}}}$$

(15)

Consequently, the generated ${{{\mathrm{O}}}}_2^{ \bullet\! - }$ began to decompose PMS and released ${{{\mathrm{SO}}}}_4^{ \bullet\! - }$ as explained in Eq. (16) which in turn initiated the degradation of 2,4-D with the aid of other free radicals. Additionally, the produced e⁻s were suggested to be responsible for releasing ${{{\mathrm{SO}}}}_4^{ \bullet \!- }$ and ${\!}^{\bullet} \!{\mathrm{OH}}$ from PMS activation as elucidated by Eq. (17). It should be noted that the intermediate products resulting from the degradation of 2,4-D were further degraded by ${{{\mathrm{SO}}}}_4^{ \bullet\! - }$, ${{{\mathrm{O}}}}_2^{ \bullet\! - }$, ${\!}^{\bullet} \!{\mathrm{OH}}$, and also h⁺ into CO₂ and H₂O (Eq. 18). However, the degradation performance of ${{{\mathrm{O}}}}_2^{ \bullet - }$ and h⁺ was not significant as it was confirmed by investigating the effect of the four reactive radicals (tertiary butyl alcohol (TBA), MeOH, potassium iodide (KI) and p-BQ), where TBA is a scavenger for ${\!}^{\bullet} \!{\mathrm{OH}}$, MeOH for both ${\!}^{\bullet} \!{\mathrm{OH}}$ and ${{{\mathrm{SO}}}}_4^{ \bullet \!- }$, KI for h⁺, and p-BQ for ${{{\mathrm{O}}}}_2^{ \bullet \!- }$. As a result, ${{{\mathrm{SO}}}}_4^{ \bullet\! - }$ and ${\!}^{\bullet}\!{\rm{OH}}$ were identified as the primary controlling radicals in the degradation of 2,4-D.

$${{{\mathrm{HSO}}}}_5^- + {{{\mathrm{O}}}}_2^{ \bullet\! - } \to {{{\mathrm{SO}}}}_4^{ \bullet\! - }$$

(16)

$${{{\mathrm{Fe}}}}_3{{{\mathrm{O}}}}_4 {\hbox{-}} {{{\mathrm{Cs}}}}_{{{\mathrm{x}}}}{{{\mathrm{WO}}}}_3/{{{\mathrm{NiAl}}}} {\hbox{-}} {{{\mathrm{LDH}}}}({{{\mathrm{e}}}}_{{{{\mathrm{CB}}}}}^ - ) + {{{\mathrm{HSO}}}}_5^- \to {{{\mathrm{SO}}}}_4^{ \bullet \!- } + \,^ \bullet \!{{{\mathrm{OH}}}}$$

(17)

$$2,4 {\hbox{-}} {{{\mathrm{D}}}} \to {{{\mathrm{Intermediates}}}} \to {{{\mathrm{CO}}}}_2 + {{{\mathrm{H}}}}_2{{{\mathrm{O}}}}$$

(18)

Interestingly, the synergistic effect of the catalyst (ZnCo−LDH@CeO₂), PMS, and light resulted in extraordinary degradation efficiency of phenol (99.8%), which was reported by Gao et al.¹³¹. Such a superb activity was assigned to the high content of cerium oxide (CeO₂) that resulted in the introduction of more active sites being exposed on the catalyst surface. Moreover, metal ions including Co²⁺ and Ce³⁺ showed excellent performance in the activation of PMS and generating ${{{\mathrm{SO}}}}_4^{ \bullet - }$ for the degradation of phenol in another study¹³². Likewise, Ye et al.¹³³ achieved 100% degradation of p-arsanilic acid (p-ASA) using CoMnFe-LDH/PMS and CoMnFe-LDO/PMS within 30 min. PMS activation was based on the direct electron transfer from adsorbed p-ASA on the surface of CoMnFe-LDO/PMS. While in the LDH system, the PMS activation depended on the coupling with surface hydroxyl groups in CoMnFe-LDH/PMS system. However, it was noticed that CoMnFe-LDH was shown to be superior to CoMnFe-LDO in terms of PMS activation owing to its significant affinity for PMS via surface $^ \bullet \!{{{\mathrm{OH}}}}$ complexation and multiple-metal redox cycle^49,134.

In one word, this section summarizes the recent progression in aromatic compounds’ degradation through the PMS activation using LDHs. In light of the aforementioned studies, the most used approaches for LDHs’ preparation are hydrothermal and co-precipitation methods. Furthermore, the fabricated LDHs are modified by forming composites with carbonaceous materials (BC, and GO) that can act as carriers and electron mediators to enhance the physicochemical characteristics of LDHs and improve the redox cycle of the transition metals. Also, mixing LDHs with metal oxides/sulfides (Fe₃O₄, CeO₂, WO₃, MoS₄) enhances the redox cycle, enriches the catalytic medium with electrons, and facilitates the catalyst separation and reusability by employing the magnetic metal oxides. Besides, the calcination of LDH to form LDO can be considered one of the most impeccable modification ways to LDHs since it could improve the morphology, surface area, and degradation activity of LDHs, along with decreasing the metal leaching during the catalytic reaction. However, such a modification technique is not commonly used. Therefore, further studies in this aspect are highly recommended.

Interestingly, LDH-based catalysts modified forms revealed an excellent degradation ability toward a number of aromatic compounds like DEP, BPA, TBBPA, NB, 2,4-D, and p-ASA since the degradation percent attained almost 100% in some cases. Although some LDH-based catalysts revealed good reusability, others showed a decline in the degradation percentage ranging from 15 to 20% after the 3^rd or the 4^th cycle. This finding implied the metals leaching during the catalytic reaction which is a serious challenge restricting LDHs’ applications in industry. In addition, the substantial metal leaching in a highly acidic medium (pH ≤ 3) was recorded in some studies to be more than four-fold of the acceptable limit by EPA.

More importantly, the degradation mechanism of aromatic compounds via the PMS activation followed two paths; a radical pathway dominated by ${{{\mathrm{SO}}}}_4^{ \bullet - }$ and $^ \bullet \!{{{\mathrm{OH}}}}$ as well as the contribution of ${{{\mathrm{O}}}}_2^{ \bullet - }$ in some cases. While in the non-radical pathway ${{{\mathrm{O}}}}_2^{ \bullet - }$ is the dominant ROS that forms a complex with PMS to generate ¹O₂. In this pathway, PMS acts as an acceptor, aromatic compounds as a donor, and LDHs as a mediator.

Unexpectedly, when the interfering ions’ effect on the degradation efficiency of the aromatic compounds was investigated, the presence of Cl⁻ ions improved the efficiency owing to their capability of enhancing the ${{{\mathrm{SO}}}}_4^{ \bullet - }$ production via the chlorine free radical. Such findings regarding the degradation of the noxious organic pollutants in wastewater via the PMS activation using LDH-based composites are summarized in Table 2.

Table 2 The process condition and degradation efficiency of organic pollutants through the PMS activation by LDH-based composites.

Degradation of pharmaceutical residues

In one study, Caoa et al.¹³⁵ fabricated AlCo-LDH through an unconventional approach using Co-ZIF owing to the high crystallinity and porosity of metal-organic frameworks (MOFs), rendering them a propitious precursor to synthesize LDHs. The as-fabricated AlCo-LDH was examined by X-ray diffraction (XRD) technique, revealing that the characteristic peaks of Co-ZIF completely changed and became identical to the amorphous Al(OH)₃. In addition, the decrease in the pH of the aluminum solution after the etching process is another evidence to confirm the hydrolysis of the metal ions as clarified by the following equations:

$${{{\mathrm{Co}}}}^{2 + } + 2{{{\mathrm{H}}}}_2{{{\mathrm{O}}}} \leftrightarrow {{{\mathrm{Co}}}}\left( {{{{\mathrm{OH}}}}} \right)_2\, + \,2{{{\mathrm{H}}}}^ +$$

(19)

$${{{\mathrm{Al}}}}^{3 + } + 3{{{\mathrm{H}}}}_2{{{\mathrm{O}}}} \leftrightarrow {{{\mathrm{Al}}}}\left( {{{{\mathrm{OH}}}}} \right)_3\, + \,3{{{\mathrm{H}}}}^ +$$

(20)

Besides, the low concentration of Co ions in AlCo-LDH also indicated the replacement of Co²⁺ by Al³⁺ in Co-ZIF to form AlCo-LDH. The catalytic performance of PMS, Co-ZIF-D/PMS, and AlCo-LDH/PMS were scrutinized by the degradation of tetracycline (TC). The degradation efficacy of TC was only 26.4% and 37.43% using PMS and Co-ZIF/PMS, respectively. However, the efficacy boosted to 96.1% when AlCo-LDH/PMS was used. The promising degradation efficiency of TC by AlCo-LDH/PMS system is most likely due to the high surface area of AlCo-LDH. The impact of different quenching reagents was examined, clarifying that the reactive species responsible for the degradation of TC were ${{{\mathrm{SO}}}}_4^{ \bullet - }$ and ¹O₂ species. These results were confirmed via EPR by detecting the characteristic peaks of DMPO-${{{\mathrm{SO}}}}_4^{ \bullet - }$ and TMP-¹O₂ upon introducing 5,5-dimethyl- pyrroline-oxide (DMPO) and 2,2,6,6-tetramethyl-4-piperidinol (TMP), respectively, into the reaction. Due to the high electronegativity of Co, it was concluded to be responsible for the electrons’ transfer to PMS and further generating the reactive species, which in turn attacked TC and degraded it into CO₂ and H₂O (Fig. 4a).

Fig. 4: The plausible degradation mechanisms of pharmaceutical residues by the LDHs/PMS system.

a The TC degradation by AlCo-LDH/PMS system, adapted with permission from ref. ¹³⁵, Copyright, 2022, Elsevier, b the IBP degradation by Carbon-ZnFe LDH/PMS system, adapted with permission from ref. ¹³⁶ Copyright, 2021, Elsevier, c the MTZ degradation using 3D Co₃Mn-LDH/rGA/PMS system, adapted with permission from ref. ¹³⁷ Copyright, 2021, Elsevier and d the possible mechanism of SMX and SCP degradation by ultrathin-NiAl-LDH/PMS system, adapted with permission from ref. ¹³⁸ Copyright, 2022, Elsevier.

Concomitantly with the above-mentioned study, Naderi et al.¹³⁶ concluded that ${{{\mathrm{SO}}}}_4^{ \bullet - }$ was majorly driving the degradation process of Ibuprofen (IBP) when Carbon-ZnFe LDH was used to activate PMS and generate sulfate free radicals, as shown in Fig. 4b. The role of LDH compound was emphasized in this study by showing that the degradation efficiency was just 24% using PMS compared to 60.6% upon adding ZnFe-LDH to the PMS solution. Such a substantial increase in the degradation efficacy was attributed to the hydrotalcite structure of ZnFe-LDH, which results in providing a high surface area and generating more degradation sites, along with the enhanced electron transfer owing to the presence of electron-rich Fe active sites. Also, it is worth mentioning that when the LDH compound was modified with carbon nanoparticles, the IBP degradation efficacy increased to 75.4%, and the bio-toxicity of the released wastewater was largely diminished to 15.8%. Such a result was accredited to the carbonaceous activating role in improving the rate of electron transfer.

In the same line, He et al.¹³⁷ highlighted the effective role of carbonaceous materials in improving the electron transfer process by scrutinizing the effect of three-dimensional (3D) graphene aerogel in enhancing the degradation efficiency of Metronidazole (MTZ) that reached 93.7% using 3D Co₃Mn-LDH/rGA as a PMS activator under visible light irradiation (Fig. 4c). Not only did 3D Co₃Mn-LDH/rGA perform as a PMS activator, but it also served as a photocatalyst, which was confirmed by recording a degradation efficiency of only 56.4% using Co₃Mn-LDH/rGA in the absence of light. Moreover, the electron transfer from 3D Co₃Mn-LDH/rGA to PMS was indicated by the oxidation of Co²⁺ to Co³⁺ and Mn³⁺ to Mn⁴⁺ with 12.2% and 13.6%, respectively. Also, when different radical scavenging experiments were conducted, it was revealed that $^ \bullet \!{{{\mathrm{OH}}}}$, ${{{\mathrm{SO}}}}_4^{ \bullet\! - }$, and ¹O₂ are the main species responsible for MTZ degradation.

In this perspective, Wang et al.¹³⁸ investigated the synergistic effect between the super-adsorption and the catalytic PMS degradation of sulfamethoxazole (SMX) and sulfachloropyridazine (SCP) using ultrathin-NiAl-LDH. Firstly, the presence of a positive charge on the surface of ultrathin-NiAl-LDH electrostatically attracted the negatively-charged SMX and SCP, and led to their super-adsorption on the surface of the LDH compound. Secondly, the PMS was activated by ultrathin-NiAl-LDH and resulted in the formation of $^ \bullet \!{{{\mathrm{OH}}}}$ and ${{{\mathrm{SO}}}}_4^{ \bullet \!- }$ that led to the degradation of both SMX and SCP into CO₂, H₂O, and other byproducts, as displayed in Fig. 4d. Additionally, it is worth mentioning that the super-adsorption effect of ultrathin-NiAl-LDH was responsible for minimizing the migration distance between the reactive radicals and the two antibiotic compounds that resulted in degradation efficiencies of 98.8% and 100% for SMX and SCP, respectively.

In another investigation, Wanga et al.¹³⁹ also adopted the etching process to fabricate FeCo-LDH from zeolitic imidazolate framework-67 (ZIF-67). As the Fe/Co ratio of 1.5 revealed the highest TC degradation efficiency, reaching 92%, it was reported that the molar ratio between Fe and Co significantly affected the catalytic performance of FeCo-LDH/PMS. This superb degradation aptitude of FeCo-LDH/PMS towards TC may be attributed to the contribution of both Fe and Co ions in producing ${{{\mathrm{SO}}}}_4^{ \bullet \!- }$ (Eqs. 21 and 22);

$${{{\mathrm{Co}}}}^{2 + } + {{{\mathrm{HSO}}}}_5^{^ - } \to {{{\mathrm{Co}}}}^{3 + } + {{{\mathrm{SO}}}}_4^{ \bullet \!- } + {{{\mathrm{OH}}}}^ -$$

(21)

$${{{\mathrm{Fe}}}}^{2 + } + {{{\mathrm{HSO}}}}_5^ - \to {{{\mathrm{Fe}}}}^{3 + } + {{{\mathrm{SO}}}}_4^{ \bullet \!- } + {{{\mathrm{OH}}}}^ -$$

(22)

The 3D-Excitation Emission Matrix Fluorescence Spectrophotometer (3D-EEMs) was utilized to define the degradation pathway of TC by FeCo-LDH-1.5/PMS (Fig. 5a–f). The characteristic peak of humic acids-like matter was observed at the range of Ex/Em (325–375 nm)/(475–550 nm), indicating the partial decomposition of TC. In addition, a noticeable increase in the peak was observed by increasing the time range from 5 to 10 min, which may be accredited to the complete decomposition of TC. Subsequently, a different fluorescence peak appeared at Ex/Em = (300–350 nm)/(375–450 nm), inferring that humic acid-like matter decomposed into CO₂ and H₂O. The efficiency of FeCo-LDH-1.5/PMS was not affected by changing the pH level. Thus, it could be demonstrating high degradation efficiencies upon being used in various aquatic environments such as medical wastewater, tap water, and rivers. The leaching of Co from FeCo-LDH-1.5/PMS after three cycles led to decreasing the concentration of Co from 11.61 to 8.85% and thereby the degradation efficiency decreased from 92% to 84%.

Fig. 5: Emission fluorescence characteristics of LDHs.

a–f The 3D-EEMs of the TC degradation by FeCo-LDH-1.5, adapted with permission from ref. ¹³⁹, Copyright, 2022, Elsevier.

In another study, Li et al.¹⁴⁰ prepared a series of CoFeLa-LDH using different molar ratios between Co, Fe, and La for the degradation of TC. The XRD patterns of CoFe-LDHs showed that the augmentation in the La concentration decreased the intensity of the characteristic peaks of CoFe-LDH. Although the peak of La-O was not detected by the XRD at low concentrations of La, it was identified by the FTIR spectrum. From SEM and Brunauer–Emmett–Teller (BET) analyses of CoFeLa-LDHs, it was noticed that CoFeLa-LDH2 with Co/Fe/La ratios (20:8:2) had a better crystallinity with an average particle size of 497 nm and a higher surface area (122.1 m² g⁻¹) compared to the other ratios. More importantly, it was observed that CoFeLa-LDH2 without PMS had a low degradation efficiency of just 10.5% against TC, and there was no such a significant difference when PMS was used on its own, as the efficiency was only 5.5%. Moreover, CoFeLa-LDH2/PMS showed a significant degradation efficiency of 90.1%. Moreover, the highest degradation efficiency was obtained at 0.05 g/L CoFeLa-LDH2, whereas increasing the CoFeLa-LDH2 dosage to 0.1 g/L led to quenching the reactive radical species, and decreasing the efficiency. In addition, the maximum TC degradation was detected at pH 5 as the amount of ${{{\mathrm{SO}}}}_4^{ \bullet - }$ and $^ \bullet \!{{{\mathrm{OH}}}}$ diminished by their reaction with H⁺ at lower pH levels as elucidated by Eqs. 23 and 24. While at high pH levels, the concentration of $^ \bullet \!{{{\mathrm{OH}}}}$ was higher than ${{{\mathrm{SO}}}}_4^{ \bullet - }$, which resulted in minimizing the degradation performance of CoFeLa-LDH2/PMS since $^ \bullet \!{{{\mathrm{OH}}}}$ is less reactive than ${{{\mathrm{SO}}}}_4^{ \bullet - }$. Nonetheless, NiFeLa-LDH/PMS demonstrated excellent catalytic efficiency toward TC over a wide pH range (3–10) because the controlling reactive radicals were ${{{\mathrm{SO}}}}_4^{ \bullet - }$ and $^ \bullet \!{{{\mathrm{OH}}}}$, which were generated at pH < 7 and pH > 9, respectively. Moreover, a high degradation efficacy was observed at pH < 7. Such an observation was accredited to the generation of ${{{\mathrm{HSO}}}}_5^ -$ in acidic and neutral media that could be attracted to the positively charged NiFeLa-LDH and the enhancement of free radicals’ generation. This result was in line with Chena et al.⁵⁰ who elucidated the constant degradation performance of MgMn-LDO towards TC over a wide pH range from 4 to 10⁵⁰.

$${{{\mathrm{H}}}}^ + + {{{\mathrm{SO}}}}_4^{ \bullet - } + {{{\mathrm{e}}}}^- \to {{{\mathrm{HSO}}}}_4$$

(23)

$${{{\mathrm{H}}}}^ + + \,^ \bullet \!{{{\mathrm{OH}}}} + {{{\mathrm{e}}}}^- \to {{{\mathrm{H}}}}_2{{{\mathrm{O}}}}$$

(24)

Quenching experiments and EPR indicated that both ${{{\mathrm{SO}}}}_4^{ \bullet - }$ and $^ \bullet \!{{{\mathrm{OH}}}}$ participated in the TC degradation, but ${{{\mathrm{SO}}}}_4^{ \bullet - }$ was the predominant species. In addition, the contribution of the non-radical species to the catalytic degradation process was observed during the reaction. Moreover, the adsorption mechanism was explained by XPS where the decline in the binding energy of Co (II) and the increment of Co (III) indicated the conversion of Co (II) to Co (III). Likewise, the binding energy of both Fe and La changed, which evinced that Co (II), La (III), and Fe (II) contributed to the generation of ${{{\mathrm{SO}}}}_4^{ \bullet - }$. Notably, the influence of the presence of interfering anions like ${{{\mathrm{NO}}}}_3^ -$, Cl^–, and ${{{\mathrm{HCO}}}}_3^ -$ was investigated, revealing a significant decrease in the degradation efficacy of TC in the presence of both Cl^– and ${{{\mathrm{HCO}}}}_3^ -$. Such behavior is most likely due to the scavenging of the produced $^ \bullet \!{{{\mathrm{OH}}}}$ and ${{{\mathrm{SO}}}}_4^{ \bullet - }$ by these co-existing anion species. These results agreed with Zhang et al.¹⁴¹ who reported a notable decline in the degradation efficiency of carbamazepine (CBZ) utilizing Ni/Co-LDH/PMS from 91 to 21% in the presence of Cl^– or ${{{\mathrm{HCO}}}}_3^ -$. Nevertheless, it was concluded that ${{{\mathrm{NO}}}}_3^ -$ had a negligible effect on the degradation efficiency of both CBZ and TC. Also, Chen et al.¹⁴² confirmed the negative effect of a high concentration of Cl⁻ on the degradation efficiency of sulfamethoxazole (SMX) by MgMnLDH-Cu-4/PMS. Such an effect could be attributed to the generation of ${\rm{Cl}}^{\bullet}$ from the interaction between ${{{\mathrm{SO}}}}_4^{ \bullet - }$ and Cl⁻, which then combined with Cl⁻ to produce ${{{\mathrm{Cl}}}}_2^{ \bullet - }$ that has a lower degradation ability than ${{{\mathrm{SO}}}}_4^{ \bullet - }$. In contrast, Zeng et al. study¹⁴³ elucidated the positive impact of F⁻ and Cl⁻ on the SMX degradation since F⁻ enhanced the release of more $^ \bullet \!{{{\mathrm{OH}}}}$ at acidic pH, resulting in an amelioration in the degradation efficiency of SMX. However, there was a significant inhibition effect of F⁻ on the degradation process at pH 9. The co-existing Cl⁻ reacted with $^ \bullet \!{{{\mathrm{OH}}}}$ to produce ${\rm{Cl}}^{\bullet}$ which had the same degradation ability as $^ \bullet \!{{{\mathrm{OH}}}}$ as follows

$$\,^ \bullet \!{{{\mathrm{OH}}}} + {{{\mathrm{Cl}}}}^ - \to {{{\mathrm{ClOH}}}}^{ \bullet \!- }({{{\mathrm{k}}}} = 4.3 \times 109\,{{{\mathrm{M}}}}^{ - 1}{{{\mathrm{s}}}}^{ - 1})$$

(25)

$${{{\mathrm{ClOH}}}}^{ \bullet \!- } + {{{\mathrm{H}}}}^ + \to {{{\mathrm{Cl}}}}^ \bullet + {{{\mathrm{H}}}}_2{{{\mathrm{O}}}}({{{\mathrm{k}}}} = 2.1 \times 1010\,{{{\mathrm{M}}}}^{ - 1}{{{\mathrm{s}}}}^{ - 1})$$

(26)

In another investigation, Yang et al.⁶¹ fabricated a series of Co_mFe_n-LDHs with different molar ratios between Co and Fe through the combination of in situ exfoliation and co-precipitation methods. Such an innovative technique resulted in the formation of a thinner structure of Co_mFe_n-LDHs with high stability, prevented the leaching of Co and Fe, and increased the active sites. The XRD patterns showed an enhancement in the crystallinity of Co_mFe_n-LDHs with raising the Co²⁺ content. On the contrary, the increase in the Fe³⁺ ratio diminished the peak intensity of Co_mFe_n-LDHs, which may be attributed to the ability of Fe³⁺ of inhibiting the crystallization in acidic conditions. Furthermore, the low interlayer spacing (7.63) of Co1Fe1-LDHs compared to other fabricated LDHs reflected its thinner structure. Noteworthy, the increase in the Co²⁺ content improved the degradation efficiency of ciprofloxacin (CIP) up to 86.9%. Furthermore, it was found that the optimum pH of the CIP degradation by Co₁Fe₁-LDHs/PMS was in the range of 5–9. This finding could be explained by the leaching of Co²⁺ and Fe³⁺ at pH < 3, whereas at pH 11, there was an electrostatic repulsion force between the dominant species of PMS (${{{\mathrm{SO}}}}_5^{2 - }$) and the negatively charged surface of Co₁Fe₁-LDHs. Conversely, Zeng et al. study¹⁴³ exhibited an enhancement in the SMX degradation at a pH range of 3–5, which may be attributed to the strong attraction between the positively charged surface of CoAlLDH@CoFe-PBA and the anionic species of PMS (${{{\mathrm{HSO}}}}_5^ -$). However, at pH ranging from 5 to 6.7, there was repulsion between SMX⁻, ${{{\mathrm{HSO}}}}_5^ -$, and the negatively charged surface of the catalyst. Also, the possibility of the reactive species’ inhibition (${{{\mathrm{SO}}}}_4^{ \bullet - }$ and $^ \bullet \!{{{\mathrm{OH}}}}$) at high pH levels retards the degradation process of SMX. However, the improved degradation efficiency was recorded at higher pH, which indicated that there were other dominant reactive species. Consequently, it was essential to determine the controlling reactive species at different pH levels. It was concluded from the quenching experiments that ¹O₂ and $^ \bullet \!{{{\mathrm{OH}}}}$ were responsible for the degradation process in an acidic medium, while ${{{\mathrm{SO}}}}_4^{ \bullet - }$ along with ¹O₂ were playing the major role in SMX degradation in an alkaline medium.

In one more attempt, Guo et al.¹⁴⁴ fabricated Calotropis gigantean fiber (CGF) and polyvinylidene fluoride (PVDF) membraned-decorated CoCu-LDH to degrade SMX via the PMS activation. In the case of CoCu-LDH@CGF composite, the tubular structure of pure CGF did not change but the surface texture became rougher, while the pure PVDF showed a highly porous surface compared to CoCu-LDH@PVDF composite. The change in the porosity between LDH@PVDF composite and pure PVDF was also confirmed by the permeability test where the permeability of PVDF decreased from 170.61 to 62.39 L⁻¹ h⁻¹ m⁻² bar⁻¹ and also the porosity decreased from 37.42 to 21.41% after the combination with CoCu-LDH. It was monitored that PVDF/PMS system exhibited a degradation efficiency that was almost equal to adsorption efficiency, indicating that PVDF had not the ability to activate PMS. The CoCu-LDH@PVDF composite membrane’s stability was tested by reusing the membrane for ten cycles. Interestingly, it was found that the diminution in the SMX degradation efficiency was insignificant, and the changes in XRD pattern, the porosity, and the permeability of the used CoCu-LDH@PVDF after the 10^th cycle were negligible. Notably, the Co²⁺ leaching percent from CoCu-LDH@CGF was larger than CoCu-LDH@PVDF, which may be attributed to the high concentration of the doped CoCu-LDH in CoCu-LDH@CGF (100 mg/L) compared to CoCu-LDH@PVDF (17.5 mg/L). Furthermore, CoCu-LDH was completely covered by a PVDF membrane, preventing Co²⁺ leaching. The degradation mechanism of SMX using CoCu-LDH@PVDF is summarized in Fig. 6a. In addition, the XPS spectra of CoCu-LDH before and after (Fig. 6b-f) the activation process showed a slight change in peak intensity and peak areas of C, O, Co, and Cu confirming the oxidation/reduction reactions on the surface of CoCu-LDH.

Fig. 6: The proposed degradation mechanism of SMX based on XPS spectra.

a The PMS activation b XPS survey, c C1s, d O1s e Cu2p and f Co2p, adapted with permission from ref. ¹⁴⁴, Copyright, 2022, Elsevier.

In another study, Zeng et al.¹⁴⁵ examined the catalytic activity of CoFe-LDH for the degradation of CBZ using the incorporation of Prussian blue analogues (PBA). The SEM and TEM images (Fig. 7a, b) of PBA showed its uniform cubic morphology. Furthermore, the SEM image Fig. 7c revealed the sheet-like morphology of PBA/CoFe-LDH, agreeing with what observed in the TEM image Fig. 7d. Furthermore, the energy dispersive spectroscopy (EDS) (Fig. 7e) indicated the presence of Co, Fe, N, and C in PBA/CoFe-LDH. The formation of the PBA/CoFe-LDH nanosheet may be attributed to electrostatic forces between the tightly bound regions of the LDH layers, which inhibited PBA growth.

Fig. 7: Morphological and elemental characterization of LDHs.

a SEM of PBA, b TEM of PBA, c SEM of PBA/CoFe-LDH, d TEM of PBA/CoFe-LDH and e EDS of PBA/CoFe-LDH, adapted with permission from ref. ¹⁴⁵, Copyright, 2022, RSC.

Firstly, PMS was used for the degradation of CBZ and there was no obvious effect on the concentration of CBZ. Similarly, PBA/CoFe-LDH was used in the absence of PMS, resulting in a quite low degradation efficiency of CBZ (<2%), but the efficiency increased to 31.8% using PBA/CoFe-LDH/PMS. Nevertheless, the degradation of CBZ by PBA/CoFe-LDH/PMS was slow, which could be accredited to the necessity of entering ${{{\mathrm{HSO}}}}_5^ -$ into the interlayers of PBA/CoFe-LDH to be activated. Whereas in most catalytic activated systems, ${{{\mathrm{HSO}}}}_5^ -$ could react with active sites on the surface of the catalyst, leading to a rapid generation of the active species. The high stability of PBA/CoFe-LDH was confirmed by its recycling for more than six cycles with no obvious change in the degradation efficiency. In addition, the XRD patterns of PBA/CoFe-LDH before and after the catalytic reaction showed a subtle difference. Such a high stability of PBA/CoFe-LDH is most likely due to the slight leaching percent of both Co (8.1 μg/L) and Fe ions (5.4 μg/L). Notably, different intermediates were produced throughout the degradation of CBZ, so the toxicity of CBZ and all intermediates were tested on different aquatic organisms including fish and daphnia. Some intermediates showed acute toxicity compared to CBZ towards fish, while others were quite toxic to daphnia.

In another investigation, He et al. studied the degradation mechanism of secnidazole (SNZ) by Co₃Mn–LDH/PMS. It was reported that SNZ was hardly decomposed by the Co₃Mn–LDH catalyzed PMS process in the presence of FFA, indicating that ¹O₂ played a significant role in the SNZ degradation process. The reaction was slightly repressed by TBA, while MeOH demonstrated a main inhibitory effect on the degradation process, denoting that both ${{{\mathrm{SO}}}}_4^{ \bullet - }$ and $^ \bullet \!{{{\mathrm{OH}}}}$ were involving in the degradation reaction. The possible generation of ${{{\mathrm{O}}}}_2^{ \bullet - }$ was investigated by using p-BQ, and it was shown that ${{{\mathrm{O}}}}_2^{ \bullet - }$ had an essential role in the degradation process. Furthermore, when potassium dichromate (K₂Cr₂O₇) was utilized as an e^– quencher, the degradation dwindled by 78.8%, showing that the e^– generated by Co₃Mn–LDH was critical in attacking the peroxy bond of PMS for the production of the aforementioned ROS. Consequently, it was revealed that the degradation process was driven by both of the radical and the non-radical pathways. Besides, XPS spectra demonstrated that 8.3% of Co²⁺ and 8.4% of Mn³⁺ converted to Co³⁺ and Mn⁴⁺, respectively after the reaction, demonstrating their contribution in the production of ${{{\mathrm{SO}}}}_4^{ \bullet - }$, $^ \bullet \!{{{\mathrm{OH}}}}$, ${{{\mathrm{O}}}}_2^{ \bullet - }$ and ¹O₂. However, it has to be noticed that high-valence Co³⁺ and Mn⁴⁺ are difficult to be reverted to a low-valence state, resulting in lower degradation efficiency in the recycling experiments. Consequently, He et al.¹³⁷ endeavored to investigate the reusability of Co₃Mn-LDH by the fabrication of 3D Co₃Mn-LDH/rGA. Unexpectedly, the degradation efficiently of metronidazole (MTZ) by 3D Co₃Mn-LDH/rGA/PMS was found to be above 67.0% after the 10^th cycle. The formation of 3D Co₃Mn-LDH/rGA was demonstrated by the abundant 3D pore network of rGA, in which abundant pores having an average size of about 2.0 nm were observed. This network could have provided broad pathways for the transfer of electrons. Additionally, it can be seen from the HRTEM image that the nanosized Co₃Mn-LDH was collected on the rGA sheets. The pristine Co₃Mn-LDH showed a weak photocurrent response (0.445 mA cm⁻²) under visible light irradiation. However, the photocurrent response increased obviously to 0.632 and 1.02 mA cm⁻² after loading Co₃Mn-LDH on 2D rGO and 3D rGA, respectively, indicating the improved separation efficiency of photo-induced carriers¹⁴⁶. In addition, the photocurrent response of Co₃Mn-LDH/rGA with 3D graphene structure was figured out to be higher than that of Co3Mn-LDH/rGO, which may be ascribed to the abundance of e⁻ transfer channels in 3D graphene aerogels.

In another investigation, Ma et al.¹⁴⁷ prepared a composite of Mg-Fe LDH@biochar for the degradation of doxycycline (DOX). The distinctive peaks of metal-oxygen linkages (MO and O-MO, where M = Mg and Fe) manifested clearly between 1000 cm⁻¹ and 450 cm⁻¹¹⁴⁸. These metal–oxygen linkages of the LDH and oxygen-containing functional groups on the biochar surface may play an essential role in DOX degradation. Additionally, a highly graphitized carbon structure with a number of lamellar could increase charge transfer and enhance DOX degradation that reached 88.76% in 120 min at pH 7 through the non-radical pathway. Moreover, Guo et al.¹⁴⁹ investigated the degradation of Lomefloxacin (LOM) utilizing Co-Cu-LDH composite, and the efficiency was 96.19% in 30 min at pH 6.67. Such a result was attributed to the presence of many active sites on the surface of Co-Cu-LDH. Furthermore, CBZ was totally degraded (100%) in just 20 min at pH 5.8 by CoMgFe-LDO catalyst⁴⁸. Such an excellent efficiency was assigned to the synergistic effect between Co and Fe active sites in the PMS solution. Also, it has to be noticed that ${{{\mathrm{SO}}}}_4^{ \bullet - }$, ¹O₂, and $^ \bullet \!{{{\mathrm{OH}}}}$ were responsible for LOM degradation, while in the case of DOX removal, the non-radical pathway was predominant.

To sum up, this section discussed the recent research articles investigating the degradation of pharmaceuticals via the activation of PMS by LDH-based catalysts. The commonly adopted preparation approaches for preparing LDH are hydrothermal, and co-precipitation methods, followed by modification methods like transition metals immobilization, calcination, and in situ etching by sodium sulfide or ZIF-67.

Surprisingly, it was found that the degradation efficacy of the most common residual contaminants; TC, DOX, IBA, SMX, SCP, CBZ, LOM, and CNZ was in the range of 86–100%. To fulfill such propitious results, several modification techniques have been applied including the formation of carbonaceous substances/LDH composites, which provide more electrons in the catalyst/PMS system. Furthermore, the preparation of layered triple hydroxides (LTH) such as MgCuFe, CoFeNi, and CoFeLa widely provides more electrons, enhances the generated ROS, and improve the oxidation/reduction cycle. Also, the incorporation of transition metals (Co, Cu, etc.) into LDHs enhanced the catalytic activity of LDHs toward pharmaceuticals’ degradation. In addition, LDO can be considered an easy, simple, costless, and effective modification method for LDHs as it was mentioned in the previous section. Notably, the calcined LTH exhibited advanced degradation efficiency toward the pharmaceutical residues and good reusability. On the other hand, the fabrication of LDH membranes showed auspicious enhancements in the reusability, durability, and catalytic activity of LDHs. Such a conclusion is based on the fact that the degradation efficiency of the pharmaceutical contaminants exceeded 92% along with a subtle decrease in the efficiency in the range of 0.5–3.1% after 10 cycles of regeneration.

Regarding the degradation mechanism of the pharmaceutical contaminants via the activation of PMS by LDHs-based catalysts, it was observed that the radical pathway is the most dominant mechanism. However, in rare cases, it was found that the degradation was driven by both radical and non-radical pathways, since the applied characterization tools and quenching test clarified the contribution of ${{{\mathrm{SO}}}}_4^{ \bullet - }$, $^ \bullet \!{{{\mathrm{OH}}}}$, ${{{\mathrm{O}}}}_2^{ \bullet - }$, and ¹O₂ in the degradation process. The recent studies regarding the degradation of pharmaceuticals through the PMS activation process by LDH composites are listed in Table 2.

Degradation of organic dyes

Unlike the traditional Fenton oxidation process, PMS activation has been considered a powerful and cost-effective approach for the degradation of persistent complex structures such as organic dyes owing to the strong oxidizing power of ${{{\mathrm{SO}}}}_4^{ \bullet - }$, its long lifetime, and excellent activity of the ROS over a wide pH range^150,151. Recently, transition metal-based LDHs, particularly, cobalt-based LDHs showed a thermodynamically favorable behavior in the PMS activation. Such a behavior was based on the suitable redox potential position of Co (II)\Co (III) (1.8 V) between ${{{\mathrm{HSO}}}}_5^ -$\${{{\mathrm{SO}}}}_5^{ \bullet - }$ (1.1 V) and ${{{\mathrm{HSO}}}}_5^ -$\${{{\mathrm{SO}}}}_4^{ \bullet - }$ (2.6–3.1 V), resulting in a remarkable degradation efficiency of various dyes^152,153. For instance, a NiCo-LDH derived from the in situ etching of Ni-MOF template was tested for the PMS activation⁶³. TEM images confirmed the successful synthesis of NiCo-LDH sheets from MOF microspheres. An optimal concentration of PMS (10⁻⁵ M) allowed NiCo-LDH to degrade reactive red-120 dye (RR-120) with an astounding efficiency of 89% within 10 minutes. Notably, the radical quencher experiment showed that both ${{{\mathrm{SO}}}}_4^{ \bullet - }$ and $^ \bullet \!{{{\mathrm{OH}}}}$ free radicals were mainly responsible for the effective degradation of RR-120. Moreover, NiCo-LDH accomplished an excellent degradation efficiency of 97% under neutral conditions compared with acidic and alkaline conditions. This behavior was attributed to the protons scavenging for both ${{{\mathrm{SO}}}}_4^{ \bullet - }$ and $^ \bullet \!{{{\mathrm{OH}}}}$ under the acidic conditions¹⁵⁴. In addition, the insufficient contact between NiCo-LDH surface and PMS due to the large number of OH⁻ formed surface complexes with LDH under alkaline conditions³⁶. Similarly, Zhao et al. demonstrated a rapid and a complete degradation of acid orange G (AOG) within 240 seconds under natural conditions (pH = 6.87) using Co-Mn LDH³⁶. Such a result was primarily attributed to the generation of ${{{\mathrm{SO}}}}_4^{ \bullet - }$ and $^ \bullet \!{{{\mathrm{OH}}}}$ free radicals, which had a strong oxidizing power towards the AOG decomposition. This assumption was further verified through the radical scavenger experiment that showed a significant inhibition of the AOG degradation below 50% after the addition of 170 mM of ethanol (EtOH) and TBA, confirming that ${{{\mathrm{SO}}}}_4^{ \bullet - }$ and $^ \bullet \!{{{\mathrm{OH}}}}$ were the main active species. More importantly, the XPS analysis exhibited an obvious shift of the Co2p_1/2 and Co2p_3/2 peaks from 796.8 eV and 781.1 eV to 796.6 eV and 780.5 eV, respectively, after the catalytic process. This finding was assigned to the coexistence of Co²⁺ and Co³⁺ due to the thermodynamically suitable standard reduction potential of Co³⁺\Co²⁺ (1.81 V) relative to that of ${{{\mathrm{HSO}}}}_5^ -$\${{{\mathrm{SO}}}}_5^{ \bullet - }$ (1.1 V), resulting in a feasible reduction of Co²⁺ into Co³⁺ by ${{{\mathrm{HSO}}}}_5^{ \bullet - }$. In addition, the atomic ratio of Mn³⁺/Mn⁴⁺ revealed a significant decline from 71.1: 28.9 to 55.2: 44.8 with a slight shift of the Mn 2p_3/2 peaks from 641.9 eV and 643.7 eV to 641.6 eV and 643.1 eV, respectively, confirming the generation of Mn⁴⁺ through the catalytic reaction¹⁵⁵. Thus, based on the XPS investigation and the radical scavenger experiments, the plausible mechanism of PMS activation was postulated as shown in the following equations: Firstly, the metal cations (Co²⁺ and Mn³⁺) converted ${{{\mathrm{HSO}}}}_5^ -$ into ${{{\mathrm{SO}}}}_4^{ \bullet - }$. Subsequently, ${{{\mathrm{SO}}}}_4^{ \bullet - }$ oxidized H₂O molecules and produced $^ \bullet \!{{{\mathrm{OH}}}}$ radicals. Afterwards, the generated ROS (${{{\mathrm{SO}}}}_4^{ \bullet - }$ and $^ \bullet \!{{{\mathrm{OH}}}}$) mineralized AOG into CO₂ and H₂O. In addition, ${{{\mathrm{SO}}}}_5^{ \bullet - }$ contributed to the AOG degradation reaction during the recovery of high valence state metals (Mn⁴⁺ and Co³⁺).

$${{{\mathrm{Co}}}}^{2 + } + {{{\mathrm{HSO}}}}_5^ - \to {{{\mathrm{Co}}}}^{3 + } + {{{\mathrm{SO}}}}_4^{ \bullet - } + \,^ - \!{{{\mathrm{OH}}}}$$

(27)

$${{{\mathrm{Co}}}}^{3 + } + {{{\mathrm{HSO}}}}_5^ - \to {{{\mathrm{Co}}}}^{2 + } + {{{\mathrm{SO}}}}_5^{ \bullet - } + {{{\mathrm{H}}}}^ +$$

(28)

$${{{\mathrm{Mn}}}}^{3 + } + {{{\mathrm{HSO}}}}_5^ - \to {{{\mathrm{Mn}}}}^{4 + } + {{{\mathrm{SO}}}}_4^{ \bullet - } + \,^ - \!{{{\mathrm{OH}}}}$$

(29)

$${{{\mathrm{Mn}}}}^{4 + } + {{{\mathrm{HSO}}}}_5^ - \to {{{\mathrm{Mn}}}}^{3 + } + {{{\mathrm{SO}}}}_5^{ \bullet - } + {{{\mathrm{H}}}}^ +$$

(30)

$${{{\mathrm{SO}}}}_4^{ \bullet \!- } + {{{\mathrm{H}}}}_2{{{\mathrm{O}}}} \to {{{\mathrm{HSO}}}}_4^ - + \,^ \bullet \!{{{\mathrm{OH}}}}$$

(31)

$${{{\mathrm{SO}}}}_4^{ \bullet - } + \,^ - \!{{{\mathrm{OH}}}} \to {{{\mathrm{SO}}}}_4^{2 - } + \,^ \bullet \!{{{\mathrm{OH}}}}$$

(32)

$${{{\mathrm{Co}}}}^{2 + } + {{{\mathrm{ e}}}}^ - \to {{{\mathrm{Co}}}}^{3 + }\Delta {{{\mathrm{E}}}} = 1.81{{{\mathrm{V}}}}$$

(33)

$${{{\mathrm{Mn}}}}^{4 + } + {{{\mathrm{e}}}}^ - \to {{{\mathrm{Mn}}}}^{3 + }\Delta {{{\mathrm{E}}}} = 0.95{{{\mathrm{V}}}}$$

(34)

$${{{\mathrm{Co}}}}^{2 + } + {{{\mathrm{Mn}}}}^{4 + } \to {{{\mathrm{Co}}}}^{3 + } + {{{\mathrm{Mn}}}}^{3 + }\Delta {{{\mathrm{E}}}} = 0.86{{{\mathrm{V}}}}$$

(35)

$${{{\mathrm{SO}}}}_4^{ \bullet - } + \,^ \bullet \!{{{\mathrm{OH}}}} + {{{\mathrm{dyes}}}} \to {{{\mathrm{several}}}}\,{{{\mathrm{steps}}}} \to {{{\mathrm{CO}}}}_2 + {{{\mathrm{H}}}}_2{{{\mathrm{O}}}} + {{{\mathrm{SO}}}}_4^{2 - }$$

(36)

According to the above reactions, the higher valence species of Co³⁺ and Mn⁴⁺ could be regenerated into Co²⁺ and Mn³⁺ through the reduction process by ${{{\mathrm{HSO}}}}_5^ -$, propagating the redox chain reactions¹⁵⁶. Also, MnFe-LDH accomplished a fascinating degradation efficiency of acid orange 7 (AO7) of 97.56% in 30 min under optimum conditions of pH = 6.1, catalyst dosage = 0.2 g\L, and PMS dosage = 0.2 g\L¹⁵⁷. Such an outstanding result was accompanied with the electronic synergy between the redox couples of Fe³⁺\Fe²⁺ (0.77 V) and Mn³⁺\Mn²⁺ (1.51 V) through the efficient charge transfer from Fe²⁺ to Mn³⁺ to generate Fe³⁺ and Mn²⁺. Consequently, ${{{\mathrm{SO}}}}_4^{ \bullet - }$ and $^ \bullet \!{{{\mathrm{OH}}}}$ radicals were generated and employed in the efficient degradation of AO7. MnFe-LDH exhibited a low degradation efficiency of 41.13% in an acidic medium due to the observable leaching of the metal active sites (i.e., Mn = 4.49 mg/L, Fe = 0.11 mg/L). Thus, the PMS activation process was suppressed as indicated by Eqs. 37 and 38.

$${{{\mathrm{H}}}}^ + + {{{\mathrm{SO}}}}_4^{ \bullet - } + {{{\mathrm{e}}}}^ - \to {{{\mathrm{HSO}}}}_4^ - + {{{\mathrm{SO}}}}_4^{2 - }$$

(37)

$${{{\mathrm{H}}}}^ + + \,^ \bullet \!{{{\mathrm{OH}}}} + {{{\mathrm{e}}}}^ - \to {{{\mathrm{H}}}}_2{{{\mathrm{O}}}}$$

(38)

In another attempt, a magnetically separable Fe₃O₄@CoFe-LDH was prepared by the co-precipitation method for the degradation of AO7¹⁵⁸. The magnetic core served as an electron donor to boost the transferred charges to the metal active sites, activating the Co(II)/Co(III) redox chain, and accelerating the generation of ${{{\mathrm{SO}}}}_4^{ \bullet - }$ radicals for better catalytic degradation of AO7. Therefore, the observed result was an excellent degradation performance of 95.1% after 15 minutes. Furthermore, Fe₃O₄@CoFe-LDH demonstrated potential heterogeneous catalytic applicability in real wastewater, inferring its outstanding stability with a degradation efficiency of 88.7% after 5 cycles of regeneration.

Within the same aspect, Li et al.¹⁵⁹ designed an innovative and scalable strategy for the efficient catalytic degradation of methylene blue (MB). Their work was based on an in situ facile hydrothermal growth of CoFeLDH onto PVDF foam assisted by the addition of NH₄F forming 3D porous foam. Notably, SEM images depicted the formation of needle-like CoFeLDH nanoarrays onto PVDF after the addition of NH₄F due to the formation of F-metal coordinate bonds, providing myriad numbers of active sites to accelerate the PMS activation. Accordingly, CoFeLDH(F)/PVDF nanoarrays achieved catalytic degradation efficiency against methylene blue (MB) two times higher than CoFeLDH/PVDF nanosheets. This spectacular result could be attributed to the high porosity and super hydrophilicity that facilitated the mass transfer and the interaction of MB with the LDH surface. Additionally, the F-Metal coordinate interaction accelerated the valence state changes of the metal active sites. Consequently, the efficient generation of ${{{\mathrm{SO}}}}_4^{ \bullet\! - }$ free radicals was promoted for a rapid de-colorization of wastewater. Interestingly, CoFeLDH(F)/PVDF nanoarrays revealed enhanced ESR signals corresponding to ${{{\mathrm{SO}}}}_4^{ \bullet\! - }$ and $^ \bullet \!{{{\mathrm{OH}}}}$ radicals comparable to CoFeLDH/PVDF nanosheets, indicating their effective role in generating free radicals. According to the EPR measurements and the radical scavenging experiments, ${{{\mathrm{SO}}}}_4^{ \bullet \!- }$ and ¹O₂ were concluded to be the main degrading species as confirmed by the reduced degradation efficiencies to 45.7% and 53.5% after the addition of FFA and methanol, respectively. The PMS catalytic activation mechanism was suggested in the following equations. Firstly, sulfate radicals were produced via the redox reactions of Fe (II) and Co (II), as elucidated in Eqs. 40 and 41. Secondly, a sequential generation of ROS including ¹O₂, $^ \bullet \!{{{\mathrm{OH}}}}$, and ${{{\mathrm{O}}}}_2^{ \bullet\! - }$ occurred (Eqs. 44–47) during the reduction of high valence state cations Fe (III) and Co (III).

$${{{\mathrm{HSO}}}}_5^ - + {{{\mathrm{SO}}}}_5^{2 - } \to {{{\mathrm{HSO}}}}_5^ - + {{{\mathrm{SO}}}}_4^{2 - } + 2\,^1{{{\mathrm{O}}}}_2$$

(39)

$$\equiv {{{\mathrm{Fe}}}}\left( {{{{\mathrm{II}}}}} \right) + {{{\mathrm{HSO}}}}_5^ - \to \equiv {{{\mathrm{Fe}}}}\left( {{{{\mathrm{III}}}}} \right) + {{{\mathrm{SO}}}}_4^{ \bullet \!- } + {{{\mathrm{OH}}}}^ -$$

(40)

$$\equiv {{{\mathrm{Co}}}}\left( {{{{\mathrm{II}}}}} \right) + {{{\mathrm{HSO}}}}_5^ - \to \equiv {{{\mathrm{Co}}}}\left( {{{{\mathrm{III}}}}} \right) + {{{\mathrm{SO}}}}_4^{ \bullet \!- } + {{{\mathrm{OH}}}}^ -$$

(41)

$$\equiv {{{\mathrm{Fe}}}}\left( {{{{\mathrm{III}}}}} \right) + {{{\mathrm{HSO}}}}_5^ - \to \equiv {{{\mathrm{Fe}}}}\left( {{{{\mathrm{II}}}}} \right) + {{{\mathrm{SO}}}}_5 ^{\bullet\! -} + {{{\mathrm{H}}}}^ +$$

(42)

$$\equiv {{{\mathrm{Co}}}}\left( {{{{\mathrm{III}}}}} \right) + {{{\mathrm{HSO}}}}_5^ - \to \equiv {{{\mathrm{Co}}}}\left( {{{{\mathrm{II}}}}} \right) + {{{\mathrm{SO}}}}_5 ^{\bullet \!-} + {{{\mathrm{H}}}}^ +$$

(43)

$$2{{{\mathrm{SO}}}}_5 ^{\bullet \!-} \to 2{{{\mathrm{SO}}}}_4^{2 - } + \,^1{{{\mathrm{O}}}}_2$$

(44)

$$\equiv {{{\mathrm{Fe}}}}\left( {{{{\mathrm{III}}}}} \right) + {{{\mathrm{HSO}}}}_5^ - \to \equiv {{{\mathrm{Fe}}}}\left( {{{{\mathrm{II}}}}} \right) + {{{\mathrm{SO}}}}_4^{2 - } + \,^ \bullet \!{{{\mathrm{OH}}}}$$

(45)

$$\equiv {{{\mathrm{Co}}}}\left( {{{{\mathrm{III}}}}} \right) + {{{\mathrm{HSO}}}}_5^ - \to \equiv {{{\mathrm{Co}}}}\left( {{{{\mathrm{II}}}}} \right) + {{{\mathrm{SO}}}}_4^{2 - } + \,^ \bullet \!{{{\mathrm{OH}}}}$$

(46)

$${{{\mathrm{O}}}}_2 + {{{\mathrm{e}}}}^ - \to {{{\mathrm{O}}}}_2^{ \bullet \!- }$$

(47)

$$\equiv {{{\mathrm{Co}}}}\left( {{{{\mathrm{III}}}}} \right) + \equiv {{{\mathrm{Fe}}}}\left( {{{{\mathrm{II}}}}} \right) \to \equiv {{{\mathrm{Co}}}}\left( {{{{\mathrm{II}}}}} \right) + {{{\mathrm{Fe}}}}\left( {{{{\mathrm{III}}}}} \right)$$

(48)

$${{{\mathrm{ROS}}}} + {{{\mathrm{Dyes}}}} \to {{{\mathrm{Intermediate}}}}\,{{{\mathrm{product}}}} \to {{{\mathrm{small}}}}\,{{{\mathrm{products}}}}$$

(49)

Surprisingly, under optimal conditions, CoFeLDH(F)/PVDF nanoarrays demonstrated a fantastic catalytic degradation efficiency of 99.56% in a dynamically packed column reaction system after the 10^th cycle in 5 minutes. Furthermore, CoFeLDH(F)/PVDF achieved good catalytic degradation efficiencies against several organic pollutants such as Rhodamine B (RhB) (100%), TC (98.59%), BPA (86.65%), and atrazine (ATZ) (45.84%), verifying its remarkable catalytic efficacy and applicability in real wastewater treatment.

In this respect, Cheng et al.¹⁵³ also used FeCo-LDH to degrade RhB via PMS activation. FeCo-LDH was fabricated in a metal molar ratio of 1:2 for Fe: Co. Different quenching experiments were established by using two radical scavengers with different concentrations: TBA for $^ \bullet \!{{{\mathrm{OH}}}}$ and ethanol for $^ \bullet \!{{{\mathrm{OH}}}}$, ${{{\mathrm{SO}}}}_4^{ \bullet - }$, and ${{{\mathrm{SO}}}}_5^{ \bullet - }$. It was concluded that the degradation process mainly depends on sulfate radicals, whereas increasing the concentration of TBA had no significant effect on the RhB degradation efficiency. Such a result was further confirmed by observing a substantial drop in the degradation efficiency upon increasing the concentration of ethanol. As shown in Fig. 8a, ${{{\mathrm{SO}}}}_4^{ \bullet - }$ radicals were generated from the PMS activation by the oxidation of Fe (II) to Fe (III) and the reduction of Co (III) to Co (II). While ${{{\mathrm{SO}}}}_5^{ \bullet - }$ was produced by the activation of PMS in presence of Co (III), and $^ \bullet \!{{{\mathrm{OH}}}}$ was generated from the interaction between ${{{\mathrm{SO}}}}_4^{ \bullet - }$ and water molecules. Finally, the produced free radicals attacked the RhB molecules and degraded them into CO₂ and H₂O as clarified in Fig. 8b. The feature that made PMS/FeCo-LDH highly efficient in the catalytic degradation process was the distinctive structure of LDH, which can keep the same structure due to the reversible oxidation and reduction of Co (II) and Fe (III).

Fig. 8: The PMS activation by LDHs-based catalysts and the structure characterization of LDHs.

a The degradation mechanism of RhB, b the PMS activation by CoFe-LDH adapted with permission from ref. ¹⁵³, Copyright, 2022, Elsevier. c The schematic diagram of MgCoAl-LDH nanosheets and nanoscrolls, d TEM of MgCoAl-LDH nanosheets, e TEM of MgCoAl-LDH nanoscrolls, and f XRD of MgCoAl-LDH nanosheets and nanoscrolls adapted with permission from ref. ¹⁶⁰, Copyright, 2022, Elsevier.

In another study, Sun et al.¹⁶⁰ inspected the catalytic efficiency of both MgCoAl-LDH nanosheets and MgCoAl-LDH nanoscrolls towards the degradation of MB via the PMS activation (Fig. 8c). TEM images (Fig. 8d, e) revealed the hexagonal structure of MgCoAl-LDH nanosheets and the curl structure of MgCoAl-LDH nanoscrolls. While XRD patterns showed that there was no difference between the crystallite phases of LDH nanoscrolls and nanosheets in terms of the position and the intensity as illustrated in Fig. 8f. It was reported that MgCoAl-LDH nanoscrolls achieved a complete degradation of MB within 40 min compared to MgCoAl-LDH nanosheets that achieved only 72% under the same reaction conditions. The superior catalytic performance of MgCoAl-LDH nanoscrolls was clearly attributed to their high surface area (63.49 m² g⁻¹) compared to that of MgCoAl-LDH nanosheets (15.43 m² g⁻¹). It is noteworthy that the radical quenching experiments revealed that ¹O₂ was the main active species, showing a remarkable decline in k_obs value by 90.2% after the addition of 1 mM L-histidine that was used as a trapping agent for ¹O₂. Additionally, the XPS spectrum survey confirmed the predominant role of Co³⁺/Co²⁺ redox cycle in the PMS activation. Such a role was described by the diminished intensity of the characteristic peak of Co³⁺ due to the reduction by PMS accompanied by a remarkable increase in the intensity of the peak corresponding to Co²⁺. On the contrary, the peaks corresponding to Mg1s and Al2p at 1303.8 eV and 74.3 eV, respectively, did not show any change after the catalytic reaction, indicating their insignificant role in the PMS activation. More importantly, the influence of co-existing anions such as ${{{\mathrm{NO}}}}_3^ -$, ${{{\mathrm{SO}}}}_4^{2 - }$, Cl⁻, ${{{\mathrm{HCO}}}}_3^ -$, ${{{\mathrm{CO}}}}_3^{2 - }$, and dihydrogen phosphate (H₂PO₄⁻) on the catalytic degradation of MB was inspected. Even at high concentrations of these interfering ions, the presence of ${{{\mathrm{NO}}}}_3^ -$ and ${{{\mathrm{SO}}}}_4^{2 - }$ had a minor effect on the degradation efficiency of MB. While the degradation efficacy significantly reduced by increasing the concentration of Cl⁻ and ${{{\mathrm{HCO}}}}_3^ -$ due to their reaction with ${{{\mathrm{SO}}}}_4^{ \bullet - }$, forming ${{{\mathrm{Cl}}}}_2^{ \bullet - }$ and ${{{\mathrm{HCO}}}}_3^ \bullet$, respectively. These formed species had lower oxidizing power, which in turn suppressed the degradation efficiency of MB. Furthermore, it is worth mentioning that the concentration of ${{{\mathrm{CO}}}}_3^{2 - }$ obviously controlled the degradation efficiency of MB. The dramatic decrease in the degradation efficiency at a low concentration of ${{{\mathrm{CO}}}}_3^{2 - }$ (0.5 mM) was due to the change in the pH medium, which resulted in the conversion of ${{{\mathrm{CO}}}}_3^{ \bullet - }$ to ${{{\mathrm{HCO}}}}_3^ \bullet$ that had a low oxidation potential. However, at a high concentration of ${{{\mathrm{CO}}}}_3^{2 - }$ (10 mM), the degradation efficiency increased as a result of generating a large amount of $^ \bullet \!{{{\mathrm{OH}}}}$ that had excellent oxidizing power. Such an assumption was further verified by EPR measurements, which showed a noticeable increase in the $^ \bullet \!{{{\mathrm{OH}}}}$ signals with higher concentrations of ${{{\mathrm{CO}}}}_3^{ \bullet - }$ comparable to the ¹O₂ EPR signals that showed no change in the signals’ intensities with different concentrations. Additionally, EPR measurements manifested that the concentration of ¹O₂ increased gradually by raising the concentration of H₂PO₄⁻, which reduced the dissociation energies of O-O bonds in PMS. Therefore, the PMS activation was accelerated, leading to the enhancement of the degradation efficiency.

Shortly, this section summarizes the recent assessments of various anionic and cationic dyes’ degradation through the activation of PMS by LDH-based catalysts (Table 2). The most common methods for producing the used LDH-based catalysts were hydrothermal and co-precipitation. Notably, the investigated PMS/LDH-based catalysts exhibited an enhanced degradation efficiency toward RR-120, AO7, Congo red (CR), AOG, MB, and RhB dyes, with almost 100% in some cases. LDHs were modified using magnetic nanoparticles (Fe₃O₄), particularly to improve their regeneration. Additionally, the efficiency was further enhanced owing to the contribution of Fe ions in the PMS activation and the increased production of ROS. In some cases, the degradation efficacy declined by just 1% after 10 cycles of regenration. A conclusion that is in line with what was concluded in the previous sections of LDO and LTH, which revealed an advanced degradation performance toward multiple dyes along with a good reusability. Noteworthy, the variable morphology of the produced catalytic materials was quite effective on the degradation efficiency of dyes. For example, CoFeLDH(F)/PVDF nanoarrays showed a degradation percent of MB two times higher than CoFeLDH/PVDF nanosheets. In addition, the superior catalytic performance of MgCoAl-LDH nanoscrolls compared to nanosheets was based on the observed surface area of MgCoAl-LDH nanoscrolls, which was four times higher than that of MgCoAl-LDH nanosheets.

The investigation of the co-existing ${{{\mathrm{CO}}}}_3^{2 - }$ impact in the reaction system revealed a dramatic decrease in the degradation efficiency of dyes at low concentrations of ${{{\mathrm{CO}}}}_3^{2 - }$. Such a decrease was accredited to the changing pH levels, resulting in the conversion of ${{{\mathrm{CO}}}}_3^{ \bullet - }$ to ${{{\mathrm{HCO}}}}_3^ \bullet$ that had a low oxidation potential. In contrast, the degradation efficiency improved at high concentrations of ${{{\mathrm{CO}}}}_3^{2 - }$ as a result of generating a large amount of $^ \bullet \!{{{\mathrm{OH}}}}$, which had an excellent oxidizing power.

Most of the reported studies clarified that the degradation of dyes using the activation of PMS by LDH-based catalysts proceeded via the radial pathway mechanism as follows;

• Firstly, the transition metal cations converted ${{{\mathrm{HSO}}}}_5^ -$ into ${{{\mathrm{SO}}}}_4^{ \bullet - }$

• Secondly, ${{{\mathrm{SO}}}}_4^{ \bullet - }$ oxidized H₂O molecules, releasing $^ \bullet \!{{{\mathrm{OH}}}}$ radicals.

• Finally, the generated ROS (${{{\mathrm{SO}}}}_4^{ \bullet - }$ and $^ \bullet \!{{{\mathrm{OH}}}}$) mineralized dyes into CO₂ and H₂O.

In conclusion the recent progress of PMS-based AOPs that represent a revolution in the field of catalysis, focusing on LDH-based catalysts owing to their advanced catalytic performance, excellent stability, and reusability are summarized in this review. The remarkable catalytic activity of LDHs towards PMS activation resulted in an excellent degradation efficiency of various organic contaminants, which could be reaching 100% owing to the exceptional properties of LDH-based catalysts. In addition, the efficient reusability of LDH-based catalysts has been confirmed as the documented degradation efficiency was more than 80% after many cycles of regeneration in the majority of research studies. Noteworthy, it was concluded that the combination of LDHs with some materials like polyvinylidene fluoride, Prussian blue analogues, biochar, and Calotropis gigantean fiber could improve the chemical stability of LDHs and directly ameliorate its reusability. Most of the reported studies investigated the leaching amounts of the metal-containing LDHs after the PMS activation, and reported that the leaching amounts were less than the allowed limits. The common preparation routes were thoroughly discussed, clarifying the merits and demerits of each one, and elucidating some techniques to enhance the produced LDHs. Such a discussion was based on the crucial influence of the synthesis approach of LDHs on their characteristics and catalytic efficiency. The co-precipitation approach was determined to be suitable for the production of LDHs because of its simple processing, fast rate, and high yield. The ratio of M²⁺/M³⁺ is another important factor that affects the LDHs’ characteristics such as surface area. Moreover, the calcined LDHs were found to possess a higher surface area along with an enhanced catalytic activity and reusability compared to the non-calcined ones. In addition, LDOs are characterized with other advanced properties compared to LDHs, including well dispersion and easy separation, rendering them highly efficient catalysts for PMS activation and even more efficient than LDHs.

All in all, the degradation of organic contaminants through PMS activation by LDH-based catalysts offers many advantages to overcome the common limitations of PMS-AOPs including:

• 1. LDHs can attract the negatively-charged PMS to its surface via the electrostatic interactions owing to its large pH_PZC, promoting the generation of ROS and improving the degradation efficiency of organic contaminants.

• 2. The remarkable stability of LDHs can solve one of the serious issues that hinder PMS-AOPs to be applied on an industrial scale, which is the low stability of the majority of catalysts. In general, the acidic medium in the PMS process could increase the metals’ leaching and collapse the structure of the catalyst. However, the large pH_PZC of LDHs could offer a pH buffer that prevents the formation of an acidic medium.

Nonetheless, LDHs have more aspects to go through and still need to be further investigated, so the main recommendations are:

• 1. Developing the green synthesis routes of LDHs using benign materials and consuming low energy. Although there are many studies reported the fabrication of LDHs in water as a benign solvent, the procedures were based on using toxic materials such as urea, ethylene glycol, dimethyl sulfoxide, etc.^161,162. Consequently, several investigations should be conducted to find out alternative and eco-friendly materials.

• 2. Improving the crystallinity of LDHs that constitutes a critical challenge to the researchers in characterizing LDHs. Recently-synthesized materials such as LDHs and MOFs have no standard and researchers depend on previous studies that have some differences. Notably, it was found that the increase in the pH of the metal precursors significantly enhances the crystallinity of the obtained LDHs at a highly basic medium (pH 10–13.2)¹⁶³. Therefore, further assessments of the preparation controlling factors are highly recommended.

• 3. Although several research studies confirmed the low amount of metals’ leaching, the accumulation of these metals in water bodies definitely affects human health and the aquatic environment. Meanwhile, some studies revealed a high metal leaching concentration at a highly acidic medium (pH ≤ 3)^61,101. Thereby, producing LDHs with controlled metal leaching at highly acidic medium still needs to be further investigated.

• 4. Determining the optimal ratio of M²⁺/M³⁺ as it is a missing factor in the majority of research articles. It was reported that the increase in the Cu concentration in NiCu-LDH, results in Jahn–Teller distortion¹⁶⁴. Hence, adjusting the M²⁺/M³⁺ ratio is an essential factor that should be considered during the fabrication of LDHs.

• 5. Targeting the utilization of real wastewater samples as there are few studies addressing this issue.

• 6. Considering the economic aspect by determining the cost of LDHs’ synthesis to design a comprehensive feasibility study.