Development in desalination and wastewater treatment: state of the art challenges, role of solar energy, and recommendations

The scarcity of freshwater is a huge challenge for humanity as it is the essence of life and a precondition for human existence and other living beings. Human activities such as agriculture, sanitation, electricity generation, recreation, industry and commerce, drinking, and many more entail a huge amount of freshwater (Bhoj et al. 2021). Principally, freshwater occupies a minor portion (∼ 3%) of all the 333 million cubic miles of world water reserves (i.e., 71% of the global area) (Bureau of Reclamation 2020), and the remaining lies as ocean-based or salty water (Shiklomanov 1993). Freshwater can come from the ground, ice sheets, ice caps, glaciers, icebergs, bogs, ponds, lakes, rivers and streams, and underground (as groundwater). It is characterized by low concentrations of dissolved salts and other total dissolved solids (TDS) and explicitly excludes saline seawater and brackish water. It also includes mineral-rich waters such as chalybeate springs (Fuller 2004). The ratio of salted water to freshwater found on the earth is 50:1 (Shiklomanov 1993), which further decreases to 1/150th (i.e., 22,300 cubic miles) of 1% of total water; most people depend on it for daily use. Unfortunately, the distribution of freshwater is also uneven across the globe, where most of it is found in the form of ice/snow, groundwater, and moisture (0.001% (USGS 2018), within the atmosphere) in plenty at some places and sparse at others. In essence, only 0.007% of the total water on earth is available to sustain 7.8 billion people (National Geographic 2013). Hence, newer ways for water desalination and wastewater purification must be explored to quench the thirst for an alarming increase in the world population.

Thus, the alarming rise in freshwater requirements compounded with chronic shortage motivated researchers across the globe to develop newer water harvesting and purification techniques (Tu et al. 2018) with the aim of producing freshwater in an economically viable and user-friendly manner (Lundqvist et al. 2019). Though small, any step to mitigate water shortage counts big when the water crisis deepens with time. Various technological advances have been made in water treatment and purification, among which few are in membrane separation (Zhou et al. 2020), seawater desalination (Wazwaz & Khan 2020), wastewater treatment (Wazwaz et al. 2019a, 2019b), and atmospheric water harvesting (Salehi et al. 2020). These technologies have made significant progress over the years; however, there still exists a domain for developing a decentralized, low-priced, and long-lasting water purification technology that has improved the efficiency and accessibility of freshwater for resource-limited provinces. At the same time, the technique should also have minimal influence on the environment and resource ingestion. In this context, seawater desalination and wastewater treatment, equipped with modern technologies, can challenge water scarcity. Recently, seawater desalination projects have been explored widely to alleviate water shortage. Notably, the traditional desalination plant is often accompanied by fossil fuel burning (especially in MED desalination) and hence progressively indicates a large amount of carbon dioxide release, affecting the air quality. On the other hand, the use of solar energy (as photovoltaic (PV) cells) in desalination is anticipated to not only reduce the fossil fuel dependency significantly but also minimize the environmental damage by reducing the greenhouse gas (GHG) emissions. For instance, the large-scale employment of PV cell-powered low-temperature multiple-effect desalination (LT-MED) replaces fossil energy and minimizes the GHG release from 1.61 × 10⁵ to 3.86 × 10⁶ t CO₂-eq/year with a payback period of 0.33 years (Ai et al. 2023). Thus, solar-based desalination uses clean and renewable energy from the sun and also supports sustainable energy transition worldwide. Furthermore, the total cost of desalination is a frequent barrier to its widespread adoption, owing to an energy-intensive reverse osmosis (RO) step. Solar PV desalination provides a cost-effective solution by utilizing abundant and freely available solar energy. The overall cost of solar desalination is becoming more competitive as the cost of solar PV systems continues to fall.

Thus, this review explicitly explored the recent developments in seawater desalination and wastewater treatment and equipped the readers with the necessary knowledge to start working on the growing water crisis. It also sheds light on the recent developments related to desalination and wastewater treatment that would provide a timely impulse to mitigate the water crisis, especially, the use of solar energy to make desalination economically attractive.

The desalination process takes away the salt and mineral components from salty water and produces fresh drinking water. The various salts that are removed during the desalination process include sodium sulfate (NaSO₄), sodium chloride (NaCl), magnesium chloride (MgCl₂), lithium chloride (LiCl), and magnesium sulfate (MgSO₄). Desalination of seawater is achieved either by thermal means or by membrane desalination technology and requires high energy to operate, producing salty residue as a byproduct (Alkaisi et al. 2017). Thermal desalination is considered an ancient method of obtaining freshwater through the condensation of vapor from the boiling of seawater (Winter et al. 2002). It includes multistage flash distillation (MSF), low-temperature multi-effect distillation (LT-MED), vapor compression evaporation (VC), cogeneration, and solar-based water distillation (Thimmaraju et al. 2018). Desalination techniques of seawater are broadly classified into thermal and membrane treatments. Each of the categories further divides into sub-routes. Among all the routes, evaporation and membrane separation are the most employed worldwide, with almost half of all plants operating are RO based, followed by LT-MED, MSF, and membrane distillation (MD) (Ali et al. 2011). In the United States alone, the desalination equipment market size is projected to increase at a compound annual growth rate (CAGR) of 9.0% from 2020 to 2027, with RO taking the major market share (Global Water Desalination Equipment Market Report 2020-2027 n.d.). Hence, the world needs seawater desalination for freshwater generation, and the technical breakthroughs (such as minimizing process energy consumption and environmental pollution) will make seawater desalination more efficient.

The commercial membrane used to purify water is a cellulose acetate membrane, (Kaiser et al. 2017), which remained favorable till the 1980s, and after that, robust thin-film composite membranes are used for the RO-based seawater desalination (Ortiz-Medina et al. 2018). The thin-film composite membrane is found stable over a wide pH range compared to cellulose-based membranes. It showed higher water permeability compared to other membranes because of its extremely thin (∼100 nm) polyamide (PA)-selective layers. Water-containing salt transport across the thin-film composite membrane is regulated by a solution-diffusion mechanism (Geise et al. 2010). Further, a dual process of NF is also utilized to desalinate the seawater (Altaee & Adel 2011) (see Figure 2(b)). The seawater fed to the dual-NF membrane system is connected in series. Permeate from the first NF membrane is fed to the second one, which finally produces potable water. The energy consumption of the dual-NF process is found to be directly proportional to the salinity of permeate.

Notably, despite having several advantages, membrane-based desalination often faces fouling issues, which reduces the process efficiency. The membrane fouling and its mitigation are discussed later in sections 2.1.3.1 and 2.1.3.2. Further, membrane desalination often requires high pressures (or higher energy) to facilitate water permeability through the membrane. It is worthwhile to note that sustainable membrane desalination requires balancing energy consumption and water production. Using nanotechnology (or nanocomposite membranes) can resolve these issues to a greater extent by allowing improved water permeability through the membrane and lowering the process cost.

Recently, the nanotechnology-based membrane, having enhanced permeability of water and maintaining a high salt rejection, has been employed for desalination (Seo et al. 2018). A vast category of nanomaterials (about 1–100 nm in size, as bimetallic nanoparticles, magnetic nanoparticles (MNPs), bifunctional nanoparticles, nanofibers, and nanosheets) are explored as membrane nanocomposite for desalination (Wu et al. 2011; Yang et al. 2012; Uppal et al. 2013; Ren et al. 2014; Tijing et al. 2014; Ding et al. 2017; Tayel et al. 2019; Xu-Peng et al. 2019; Wang et al. 2019; Azadi et al. 2021; Bhoj et al. 2021). The membranes comprising of thin-film of cerium oxide (CeO₂) nanoparticles successfully achieved high salt rejections (>90%) and surmounted the attack of Escherichia coli compared to pure PA membrane during desalination (Lakhotia et al. 2018). Upon slight modification, the CeO₂ membrane increased the rejection of hydrophobic contaminants and reduced the fouling due to a quick transformation Ce⁺³ ↔ Ce⁺⁴ for enhancing its antioxidant properties and acting as a biocide against bacterial strains. Zirconium (IV)-carboxylate metal-organic framework (MOF) with UiO-66 nanoparticles is incorporated in the PA-selective layer and forms a thin-film nanocomposite membrane utilized for desalination (Ma et al. 2017). The addition of UiO-66 (0.1 wt.%) provided a maximum water flux of 40 and 25% over the unmodified pure PA thin-film composite control under PRO and forward osmosis (FO) modes at the feeding of 1M NaCl. UiO-66 acted as water channels to facilitate the permeation of water while blocking hydrated cations (Na⁺, K⁺, Ca²⁺, and Mg²⁺). The bimetallic metal-organic framework (BMOF) prepared using zinc (Zn) and cobalt (Co) metals and a porous carbon indicated an enhancement in the salt removal during desalination (Ding et al. 2017). The nickel and cobalt nanoparticles could enable brackish and seawater desalination under the magnetic field's influence. The MNPs of Fe₃O₄ embedded in gelatin successfully isolated the salt (or sodium chloride) from the water (Tayel et al. 2019). Another Fe₂O₃ nanoparticles, when combined with a polyvinyl chloride (PVC) membrane, improved the membrane's antifouling properties, solute elimination, and water flux (Alghamdia et al. 2019). The PVC composite membranes with 6% MNPs indicated enhancement in the water flux (55%) and salt rejection (by 40%) when compared with pure PVC membranes. Chitosan nano-biopolymer (CS-NPs) embedded in membrane matrix utilized in the FO desalination application and achieved a high osmotic water flux and low reverse salt transport (Ghaemi & Zahra 2019). The CS-NPs improved the water hydrophilicity of the membrane surface, resulting in more adhering of water molecules over the membrane surface and hence facilitating its transfer through the membrane. Thus, the nanocomposite membranes appeared promising for improving seawater desalination. However, more research and development for nanocomposite membranes is still needed to optimize their stability, performance, and long-term durability. This is because the dispersion of nanomaterials within the polymeric matrix remains challenging. Often, the agglomeration of nanomaterial within the membrane leads to uneven pore size, which affects the permeability and ultimately disturbs the selectivity. Further, issues such as nanomaterial detachment from the membrane, fouling of the membrane, and process cost will always be crucial for realizing the application of nanotechnology (or nanocomposite membranes) for practical desalination systems.

The water and ions are transported through CNT membranes between 6 and 11 A°. The high flow rate is mainly attributed to the atomic smoothness and molecular ordering of the CNTs where water molecules pass on a one-dimensional single-file procession (Rashid & Stephen 2017). The possible water transport through a CNT/polymer composite membrane is illustrated in Figure 3(b).

Despite benefits, applying CNTs for desalination faces challenges because of the complexities in fabricating sub-nanometer tubes. Based on the fabrication methods, the CNTs membranes are classified as (i) freestanding CNT membranes and (ii) mixed (or nanocomposite) CNT membranes. The freestanding CNT membranes are further divided into vertically aligned CNT (VA-CNT) and buckypaper membranes. The VA-CNT membrane embedded with silicon nitride is reported to increase the water flux by threefold greater than those predicted using the Hagen–Poiseuille equation (Holt et al. 2004). Similar VN-CNT prepared using chemical vapor deposition (CVD) indicated high water permeability and antifouling properties necessary for desalination (Holt et al. 2004). The salt rejection of the VN-CNT membrane is found comparable with the UF membrane. During the desalination, the bulky paper CNT membranes utilized in direct contact membrane distillation (DCMD) removed 99% salt and provided about 12 kg/m²h water flux (Dumée et al. 2010a). The similar buckypaper CNT membrane with polymer support enabled 95% salt rejection but affected the water permeability in DCMD: Direct contact membrane distillation during desalination (Dumée et al. 2010b). Further, the super-hydrophobic ceramic-based CNT composite membrane is used for desalination (Dong et al. 2018) and exhibited a stable high flux of 99.9% rejection of Na⁺. A thin-film membrane of multi-walled carbon nanotubes (MWCNT) and aromatic PA is utilized for desalination (Inukai et al. 2015). The MWCNT (≈ 15.5 wt.%) improved the membrane efficiency in terms of flow rate and antifouling process because of its features, i.e., length, dispersibility, diameter, chemical functionalities, and straightness. Chlorine-resistant RO membrane fabricated using a multi-walled CNT-PA is used for desalination (Ortiz-Medina et al. 2018). The module efficiency after chlorine exposure (4,800 ppm·h) remained unaffected at a level of 99.9% and significantly reduced to 82% in the absence of MWCNT. Further, the surface roughness of the modules altered significantly by adding MWCNT (12.5 wt.%), and it improved degradation resistance against chlorine exposure, increased the water flux and maintained salt rejection performance. Incorporating carboxylated MWCNTs into a PA rejection layer improved the membrane hydrophilicity, surface area, roughness, and porosity, eventually increasing the FO water flux (Rashed et al. 2020). However, the salt rejection dropped with an increase in CNT content due to the agglomeration of CNT and their negative impact on the interfacial polymerization (IP) process of the PA rejection layer. Thus, despite showing the potential for desalination, CNT-based membranes face the constraint of tedious fabrication methods, which often results in the different diameters of tubes and chirality. This eventually affects the membrane performance during desalination. Further, the purity of membranes (i.e., metal catalyst residue and amorphous carbon) can restrict water transport, enhance fouling, and minimize salt rejection. The tube diameter and functionalization of CNTs can affect the water selectivity drastically during desalination. Thus, it is challenging to obtain high salt rejection rates without sacrificing the water flux for CNT-based membranes.

It is interesting to note that the nanoplatelets graphene-coated polyethylene membrane is utilized for desalination in a pilot-scale DCMD unit. The membrane exhibited the highest flux of 16.7 L/h/m (LMH) with 99.5% salt rejection. The desalination process required 152 kWh/m³ specific thermal energy at 85 °C inlet feed temperature, 75 L/h feed rate, 48 L/h permeate rate, and salt initial concentration of 57,500 ppm. The membrane, however, was affected because of fouling after 77 h of operation, which declined the flux to 69%. In another study, the ultrathin GO membranes were utilized for pilot-scale desalination of dye house effluent. The GO membranes after a long 45 days' pilot-scale trial enabled 77% rejection of salt from the mixed salts present in the dye house effluent and produced the water complying with the Chinese industry standard.

In addition to the CNTs and graphene, the MOFs and COFs in porous crystalline form can be good alternatives in membrane desalination because of their structural diversity and flexibility and easy integration into polymer matrices. An aluminum MOF-303 membrane utilized for desalination showed good divalent ions rejection (i.e., 93.5% for MgCl₂ and 96% for Na₂SO₄) because of its unified size and electrostatic repulsion. The MOF membrane is recommended for water softening. In another study, the 2D nanoporous MOF membranes showed excellent water flux and high salt rejection. The membrane remained high by controlling its thickness from one layer to five layers. The double-layered MOF membrane enabled 100% salt (or NaCl) rejection with high water permeability (45 L/cm²/MPa/day). A hydrophobic MOF loaded on a cellulosic membrane enabled 72% water permeability and 180% enhancement in the water flux compared to an unmodified cellulosic membrane during the desalination. Further, the MOF introduction improved the antifouling properties of the membrane compared to the unmodified one. The MOF nano-flakes horizontally aligned over PA membranes showed enhanced desalination performance with very high boron and NDMA: N-nitrosodimethylamine rejections (>90%) compared to the state-of-the-art RO membranes. In a novel approach, the MOFs loaded over light-responsive poly(spiropyran acrylate) (PSP) produced around 140 L/kg day freshwaters during adsorptive desalination. The sunlight-regenerable MOF-PSP adsorbent system consumed very low energy (0.11 Wh/L) for desalinating 2,333 PPM brackish water.

Like MOFs, covalent organic frameworks (COFs) are also utilized for seawater desalination because of their uniform pore size enabling molecular/ion separations. In a study, the COF membranes containing TaPa-SO₃H nanosheets linked by TpTTPA nanoribbons exhibited 99.91% salt rejection and provided an ultrafast water flux (267 kg/m² h). The COF membrane performed 4–10 times better than conventional membranes, with excellent stability and tolerance to salinity (up to 7.5%). A multilayered imine-linked COFs (i.e., TpPa-1 COFs) membranes indicated excellent ion rejection but on the cost of water presence. TpPa-1 COF membranes provided 100% magnesium chloride ion (MgCl₂) rejection along with water permeation one to two folds higher than commercial nano-sized membranes. In an integrated approach to addressing energy and water crises, the benzoxazole-linked COF sponge supported on a porous polymer scaffold (polydimethylsiloxane) is utilized for desalination and uranium recovery. The framework showed an enhanced evaporation rate (1.39 kg/m² h) and an excellent uranium recovery capacity (5.14 ± 0.15 mg/g). Thus, both MOFs and COFs gained considerable attention in research in recent years for membrane desalination. However, it is worthwhile to note that the synthesis of MOFs and COFs is tedious and their stability under harsh conditions of desalination (such as high salinity, high pressure and temperature) is still a question in research. Further, the scale-up of MOFs and COFs-based desalination requires interdisciplinary research including materials science, chemistry, and engineering aspects.

Membrane desalination has evolved significantly over the last three decades as an excellent alternative to adsorption, ion exchange, and sand filtration (Inukai et al. 2015). It uses a simple process of ultra/micro/nanofiltration or RO, which retains the fine solid particles/salts/ions from the feed solutions (or seawater) during filtration. A semi-permeable barrier (or membrane) is applied to separate impurities/salts from the feed water. The membrane is classified according to membrane pore sizes, viz. MF, UF, NF, and RO. The specific pore size of the membrane allows molecules of a known size (size smaller than membrane) to pass through it (via diffusion), while other molecules/ions/salts/particles larger than membrane size get retained. Usually, particle size is the sole criterion for the selection of a membrane (Muro et al. 2013). The diffusion rate of molecules/ions/salts/particles through the membrane requires driving forces such as concentration gradient, pressure gradient, and permeability under operating conditions. The membrane separation process is used for water purification, seawater desalination, and wastewater treatment. The permeability of the solvent and separation selectivity are the major factors that characterize these membranes. Several membrane separation processes used for water purification/seawater desalination are described in Table 1.

The major challenge to seawater desalination is membrane fouling, which reduces the water flux over the course of operation and hence increases the desalination operation's energy consumption. Fouling also affects the membrane's lifespan by irreversibly damaging it. The following section discusses membrane fouling as a key issue in seawater desalination and then presents the mitigation of membrane fouling by using an effective coating. Further, the economic problems concerning membrane desalination are also highlighted.

Membrane separation for water treatment has many advantages over conventional techniques; however, membrane fouling is an ongoing challenge preventing its widespread application. It is crucial to minimize membrane fouling to maintain the operation and permanency of desalination membranes. The composition and properties of seawater and brackish water used as feed in desalination plants have a direct impact on membrane efficiency (El-Manharawy & Azza 2001). Several fouling agents, such as proteins (Muro et al. 2013), and natural organic matter (NOM) (Selatile et al. 2018), are a continuous nuisance to membrane desalination. The addition of calcium ion (Ca²⁺) controlled the membrane fouling mainly from algae-rich seawater by binding with Ca²⁺ ions with negative functional groups of extracellular organic matter; a bridge is formed that results in an easily removable porous layer (Ma et al. 2020). In a study, Zhang et al. (2013) observed that the fouling of hollow-fiber UF membrane units treating algal-rich water reduced significantly because of the supplementation of K⁺, Ca²⁺, and Al³⁺. Supplementation of chemicals is costly and not environmentally and economically accepted, particularly for seawater treatment where some residuals are retained and create quality problems. In the case of UF, the precoated CNTs or nanofiber-layered CNFs: Carbon nanofibres are used for fouling reduction during water purification (Cheng et al. 2019). The presence of coating layers improved the rejection rate of NOM. CNTs performed better than CNFs in alleviating reversible and irreversible membrane fouling. The pre-adsorption using CNTs and the contribution of size exclusion to the foulants rejection of membrane was the primary mechanism where the organics linked with the underlying membrane are largely reduced. The sea or brackish water containing suspended solids, salts, and organic compounds/microorganisms, if not controlled properly, may cause a serious impact on the fouling phenomena, which eventually result in irreversible damage to desalination membranes and decrease in their lifetime. The approach to reduce membrane fouling is to perform the pretreatment of feed water (or seawater) to remove colloids, suspended solids/organic matter, and other contaminants. The possible pretreatment methods include sedimentation, flocculation, coagulation, and filtration (Khouni et al. 2020). Further, the accumulated particles/salts can be removed from the membrane surfaces by regular backwashing, improving the water permeability and minimizing energy consumption (Curcio et al. 2015). Moreover, it is possible to prevent mineral scaling and lower the risk of fouling by treating seawater with anti-scalants or dispersants. The anti-scalants provide aid in stopping the precipitation of mineral salts, while dispersants can break up clumps of particles in the water to lessen fouling from colloidal substances. Monitoring membrane performance such as permeability, pressure drop, and salt rejection can detect fouling early, which can be eliminated with less effort.

The presence of coating layers significantly delayed the transition mode of fouling microorganisms from pore size blocking to the filtration of the formed cake. For instance, the rejection rate of humic acids (HA) from water by virgin membrane unit is found only 46%, which significantly increased up to 82–90%, and 63–69% for CNTs/CNFs coated membranes. The HA diffused back to the feed due to tortuous water channels (Contreras et al. 2009) and mainly entrapped onto the CNTs/CNFs layer, filing the interstitial pore space of the coating layer, forming heterogeneous fouling cake with a low porosity (Taheri et al. 2013). Adsorption of HA alters the surface of CNTs/CNFs creating the electrostatic repulsion between particles, which affects the resistance of membrane fouling (Jermann et al. 2008). In general, predepositing with CNTs and CNFs might increase membrane flux to some extent. Removal of organic matter from water by powdered activated carbon (PAC)/UF unit is reported in the literature (Tomaszewska & Sylwia 2002; Saja et al. 2020). The HA is removed up to 90%, and the complete removal of phenol is obtained at a PAC dosage of 100 mg/L. Only 40% of HA is rejected in the UF process without PAC addition, and the phenol is passed through a membrane unit. In another report, the coating of hydrophilic hyper-branched poly(amido amine) (PAMAM) over RO membrane increased the water permeability by 20–25% (Nikolaeva et al. 2015) . Further, the coating of polyethylene glycol (PEG)-like polymer over sheet membranes increases the hydrophilicity of the membrane surface, which eventually contributes to water permeation through the membrane (Zou et al. 2011). The application of a MF unit coupled with PAC for water treatment is investigated by Kim et al. (Kim et al. 2007). The turbidity of the treated water is recorded below 0.1 nephelometric turbidity unit (NTU), and the total organic carbon (TOC) and UV254 removal rates are found more than 90%. Thus, the surface modifications of desalination membranes are done very often with the aim of minimizing membrane fouling and facilitate water permeation flux. Major developments related to pressure-driven membrane separation for wastewater are compiled in Table 2.

Table 2

Potential application of pressure-driven membrane processes in water treatment

Process .	Composition .	Wastewater treated .	Operating conditions .	Results .	Reference .
UF	Polysulfone	Vegetable oil factory	Temperature = 30 °C, pH = 9	COD (91%), TOC (87%), TSS (100%,), [] (85%), and [Cl−] (40%)	Mohammadi & Ashkan (2004)
UF	Bentonite membrane deposited on ceramic perlite support.	Soluble dyes	Bentonite = 0.25–1.50 wt.%	Rejection value Direct Red 80 (97.0%) and Rhodamine B 80.1%	Saja et al. (2020)
MF-RO	Hollow-fiber membranes with spiral wound with tape-out wrap membranes	Urban wastewater	Bioreactor system: hydraulic retention time (HRT) = 47 h, feed water inflow = 16,500 m3/day RO system: feed water flow rate = 2 m3/h, flux = 323 L/h/m2	Micropollutants pesticides and pharmaceuticals removed to discharge limit	Rodriguez-Mozaz et al. (2015)
MF	Tubular ceramic membrane using pozzolan	Pretreatment of seawater for desalination	Membrane average pore size = 0.36 μm water permeability = 1,444.7 L/h·m2 bar	Rejection of turbidity (98.25%) and retention of chemical oxygen demand (70.77%)	Achiou et al. (2017)
MF	Polyvinylidene fluoride (PVDF)	Oily wastewater	Fee flowrate = 3 mL/min, pressure = 2 bar	95% removal of organic contaminants	Wang et al. (2009)
NF	Poly(vinylidene fluoride), PVDF/Brij-58 blend	Contaminated water	Dye initial concentration = 15 ppm, pressure = 6 bar	Dye rejection value (90%).	Nikooe & Ehsan (2017)
NF	Graphene oxide (GO) with TiO2 Np	Industrial wastewater	Feed water flux = 22.43 L/h/m2, pressure = 4 bar, TiO2 loading on GO = 0.2 wt%	Na2SO4 rejection of 98.8% with excellent antifouling capacity for BSA and dyes	Wang et al. (2017)
NF	Polyethersulfone (PES) membrane with goethite-tannic acid Nps	Soluble dyes removal	Feed water flux = 23.43–60.65 kg/m2, goethite-tannic acid on PES = 0–1 wt%	Dye rejection for Direct Red 16 92.61%.	Saniei et al. (2020)
RO	Graphene oxide (GO) Nps with polyamide membrane	Seawater desalination	Feed water CaSO4 concentration = 20 mM, pressure = 25 bar, flowrate = 1 L/h, temperature = 25 °C	Inhibition of bacterial growth by 81.7% and mineral scaling	Ashfaq et al. (2020)
RO	Polyamide membrane with zwitter ion	Desalination, contaminated water	Feed water permeability = 1.65 L/h/m2/bar, pressure = 15 bar, temperature = 25 °C salt concentration = 15 ppm	Flux recovery ratio (FRR) for BSA (98.9%), SA (99.0%), and SDS (95.8%)	Li et al. (2020)

Process .	Composition .	Wastewater treated .	Operating conditions .	Results .	Reference .
UF	Polysulfone	Vegetable oil factory	Temperature = 30 °C, pH = 9	COD (91%), TOC (87%), TSS (100%,), [] (85%), and [Cl−] (40%)	Mohammadi & Ashkan (2004)
UF	Bentonite membrane deposited on ceramic perlite support.	Soluble dyes	Bentonite = 0.25–1.50 wt.%	Rejection value Direct Red 80 (97.0%) and Rhodamine B 80.1%	Saja et al. (2020)
MF-RO	Hollow-fiber membranes with spiral wound with tape-out wrap membranes	Urban wastewater	Bioreactor system: hydraulic retention time (HRT) = 47 h, feed water inflow = 16,500 m3/day RO system: feed water flow rate = 2 m3/h, flux = 323 L/h/m2	Micropollutants pesticides and pharmaceuticals removed to discharge limit	Rodriguez-Mozaz et al. (2015)
MF	Tubular ceramic membrane using pozzolan	Pretreatment of seawater for desalination	Membrane average pore size = 0.36 μm water permeability = 1,444.7 L/h·m2 bar	Rejection of turbidity (98.25%) and retention of chemical oxygen demand (70.77%)	Achiou et al. (2017)
MF	Polyvinylidene fluoride (PVDF)	Oily wastewater	Fee flowrate = 3 mL/min, pressure = 2 bar	95% removal of organic contaminants	Wang et al. (2009)
NF	Poly(vinylidene fluoride), PVDF/Brij-58 blend	Contaminated water	Dye initial concentration = 15 ppm, pressure = 6 bar	Dye rejection value (90%).	Nikooe & Ehsan (2017)
NF	Graphene oxide (GO) with TiO2 Np	Industrial wastewater	Feed water flux = 22.43 L/h/m2, pressure = 4 bar, TiO2 loading on GO = 0.2 wt%	Na2SO4 rejection of 98.8% with excellent antifouling capacity for BSA and dyes	Wang et al. (2017)
NF	Polyethersulfone (PES) membrane with goethite-tannic acid Nps	Soluble dyes removal	Feed water flux = 23.43–60.65 kg/m2, goethite-tannic acid on PES = 0–1 wt%	Dye rejection for Direct Red 16 92.61%.	Saniei et al. (2020)
RO	Graphene oxide (GO) Nps with polyamide membrane	Seawater desalination	Feed water CaSO4 concentration = 20 mM, pressure = 25 bar, flowrate = 1 L/h, temperature = 25 °C	Inhibition of bacterial growth by 81.7% and mineral scaling	Ashfaq et al. (2020)
RO	Polyamide membrane with zwitter ion	Desalination, contaminated water	Feed water permeability = 1.65 L/h/m2/bar, pressure = 15 bar, temperature = 25 °C salt concentration = 15 ppm	Flux recovery ratio (FRR) for BSA (98.9%), SA (99.0%), and SDS (95.8%)	Li et al. (2020)

Note: COD, chemical oxygen demand; TSS, total suspended solids; TOC, total organic carbon; BSA, bovine serum albumin; SDS, sodium dodecyl sulfate; SA, sodium alginate.

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Solar-powered membrane desalination is a sustainable and environmentally friendly approach for freshwater generation from seawater. RO and NF techniques can be integrated with solar energy to remove salt and other impurities from seawater or brackish water. The PV system, coupled with typical membrane-based desalination, is used to drive pumps, motors, and other electrical parts of a system, thus significantly minimizing the fossil energy load. Being renewable in nature, solar energy is the prime choice for sustainable freshwater production from seawater and for mitigating the environmental impact caused by fossil energy sources. In this context, a coupled system of solar PV and DCMD is investigated (Rabie et al. 2023). The system produced both fresh water and electricity simultaneously. Further, the Janus membrane (staple carbon fabric/polyurethane composite membrane) enables about 67% water evaporation under 1 sun illumination (Wu et al. 2023). The hydrophilicity and porosity of the Janus membrane are found to be crucial for water transportation and then water. Another study found that a 3D graphene/CNTs/polypyrrole foam (GCPF) is able to effectively desalinate high-salinity brine using solar energy without losing its mechanical properties (He et al. 2023). The 3D GCPF showed a stable water evaporation rate (4 kg/m² h) for 30 h, reducing the salinity by up to 25%; this is the best-reported result among solar desalination systems and demonstrates the system's remarkable salt tolerance and durability. Several membrane desalination technologies driven by solar power are listed in Table 3, along with their specific energy requirements and water processing capabilities. Table 3 shows that solar energy is being extensively utilized in various large-scale desalination plants worldwide to produce freshwater from sea/brackish water. ED and MED appeared promising since they indicated higher freshwater production capacity (72–200 m³/day) (Ishimaru 1994; Milow & Eduardo 1997) although MED required higher specific energy to drive the desalination plant. The MED run using solar ponds still showed good freshwater production (200 m³/day) with minimum specific energy consumption (SEC).

Amongst the several thermal routes, the evaporative approach of seawater desalination remains attractive. Seawater purification using a solar still (solar thermal desalination) is a viable option. Solar energy is utilized in the solar still to heat and then evaporate the water from the seawater or wastewater steam. The evaporated water is then allowed to condense to produce pure/fresh water. A solar still can be an effective and environmentally friendly unit for desalinating seawater. However, it includes low efficiency in producing clean water compared to other water treatment techniques. Further, the solar still strongly depends on the weather conditions, particularly the intensity of the sunlight. Also, the rate of water evaporation remains low making the solar still technique less favorable for dealing with the large quantity of saline water. Moreover, the component of solar still can be often prone to scaling and fouling issues, which eventually minimizes the water purification efficiency. The high space requirement poses difficulties for setting up the solar still in densely populated or constrained areas.

The LT-MED for seawater desalination has become striking in recent years because of solar energy utilization. While LT-MED is able to fully utilize waste heat from thermal power plants and reduce the fossil energy combustion in seawater desalination projects, it still results in relatively high energy consumption and GHG emissions from the desalinated water. PV solar energy seemed an excellent alternative for traditional thermal power plants in running MED desalination (Ai et al. 2023).

In another study, GeSe-based photo electrodes demonstrated excellent photo-thermal (PT) and PV performance (Cai et al. 2023). The coupled effect of PT and PV is found promising for seawater desalination. The salinity of seawater is decreased from 27 parts per thousand (ppt) to 0.7 ppt upon desalination. A wood-based solar evaporation generator unit made from the loading of polyoxovanadate on a wood surface is reported for seawater desalination (Zhang et al. 2023). The unit indicated 98% light absorption, lowered heat loss, enabled quick water transport, and provided 2.23 kg/m² h water evaporation rate. A novel 3D nitrogen-doped carbon containing multilayered chitosan MXene nanocomposites utilized in a desalination unit showed increased hydrophobicity, lower thermal conductivity, and excellent light absorbance, suitable for seawater desalination (Jin et al. 2023). The water evaporation is determined as 1.428 kg/m² h under 1 kW/m² irradiation. The composite material-based unit demonstrated a good capability for seawater desalination with good cyclic stability and is proposed as a potential unit for practical seawater evaporation. The carbon black/chitosan decorated microcapsules used in the seawater desalination unit showed 95% solar energy absorption, excellent salt resistance, and antibacterial ability (Chen et al. 2023). The desalination unit showed 2.58 kg/m² h water evaporation (which is enhanced by 1.03 kg/m² h than conventional solar evaporator). It also indicated 90% efficiency for successive and stable seawater evaporation. A nanostructured hierarchical carbon-aerogel composite showed promising interfacial evaporation and is utilized successfully for seawater desalination (Gan et al. 2023). In another study, the PT aerogel prepared from natural cellulose fiber possesses structural stability in seawater, indicating a high desalination performance of continuous water evaporation for up to 76 days (Nguyen et al. 2023). The study, for the first time, demonstrated the utilization of natural cellulose fiber for solar thermal desalination. Further, the utilization of polyaniline (PANi) embedded liquefied-chitin polyurethane foam in the solar thermal desalination units showed 2.18 kg/m² h water evaporation (Wan et al. 2023). A carbonized Zr-based MOF (i.e., carbon-ZrO₂/PDA/polyurethane foam (carbon-ZrO2/PDA/PU)) used in the desalination unit demonstrated 1.626 kg/m² h water evaporation with no salt deposition (Chao et al. 2023). The evaporation rate remained constant with prolonged light exposure, even in 10% brine.

The literature shows that the different categories of materials utilized in solar thermal desalination offered higher energy efficiency (by better light absorption), improved water evaporation rates, and good stability for prolonged desalination. Despite the benefits, solar thermal desalination is susceptible to seasonal and daily variations in solar energy. The weather conditions (mainly cloudy weather) and operation during the night can be a constraint for solar thermal desalination units unless an uninterrupted supply of energy (to run the desalination unit) is provided. Another issue is the availability of land since the solar thermal desalination units on a large scale require significant land area, which is often challenging in densely populated or land-constrained areas.

The biological wastewater treatment method uses bacteria to destroy organic matter present within the wastewater stream under controlled temperature, pressure, pH, and substrate concentrations (Coha et al. 2021). The wastewater generated from agricultural runoff, industry, and households can be treated through a biological route that utilizes aerobic, micro-aerophilic, and anaerobic bacteria for the biodegradation of organic matter in the wastewater. Particularly, the aerobic bacteria used oxygen (as an electron acceptor) for the oxidation of organics (electron donor) of wastewater and transformed it into carbon dioxide and water (Wen et al. 2020). The aerobic treatments of wastewater include activated sludge, rotating biological contactors, biofilm systems, moving bed biofilm, sequential batch reactor, oxidation and ditch. (El-Kamah et al. 2011).

The characteristics of pollutants detected in different types of wastewater are quite different; hence, it is hard to define a specific membrane for its rejection. Nonetheless, it is generally observed that the pollutants range in industrial wastewater as compared to sewage are quite high (Feng et al. 2021; Ganiyu et al. 2021; Li et al. 2021; Saba et al. 2021; Tan et al. 2021; Zheng et al. 2021). The values of chemical oxygen demand (COD) detected in industrial wastewater lie in the range of thousands of micrograms in a liter, while in sewage or surface water, it is seen as a few hundred. This makes industrial wastewater treatment more complex with extending retention time of up to 10–100 h, quite long compared to sewage or surface wastewater treatment, i.e., 10–20 h (Ahmad et al. 2021b; Bhat & Parag 2021; Ravikumar et al. 2021; Sandoval & Ricardo 2021; Wang et al. 2021c). The biological treatment removes the majority of the pollutants directly or using some chemical aids. The membrane process uses specific treatment with a specific membrane designed for a particular type of wastewater sewage or surface. It may include aerobic followed by membrane filtrations, anaerobic treatment process followed by aerobic then followed by membrane treatment (Ahmad et al. 2021a; Liu et al. 2021b; Qi et al. 2021; Prete et al. 2021; Vieira et al. 2021). In some studies, the anaerobic process is followed by membrane treatment, constituting 85 out of more than 180 cases and around 51% of the total treatment capacity (Coha et al. 2021; Giwa et al. 2021; Revilla et al. 2021; Taoufik et al. 2021; Xie et al. 2021). It was observed that around 90% of COD, BOD, nitrogen, and total phosphate were removed during the biological process (Chauhan et al. 2021; Du et al. 2021; Jain et al. 2021; Ng et al. 2021; Prete et al. 2021; Rodríguez-Narváez et al. 2021; Thakur et al. 2021; Vieira et al. 2021; Xie et al. 2021).

Another aspect of biological treatment incorporates the removal of nitrogen from the polluted water, which often remains complicated. Different types of bacteria capable of removing nitrogen from the wastewater include ammonia-oxidizing bacteria (i.e., nitrosomance and nitrobacter), anammox bacteria (i.e., nitritation anammox), and denitrification. Further, the presence of organics (electron donors) in the wastewater inhibits the nitrifiers due to the consumption of available oxygen by heterotrophic bacteria. The growth of heterotrophic bacteria is much higher than nitrifiers creating unfavorable conditions for ammonium oxidation (Wen et al. 2020). The growth of anammox bacteria is quite slow and sensitive to the organics, and oxygen, and the deterioration of efficiency deteriorates at an unbalanced ammonia/nitrite ratio. Pre-denitrification/post-denitrification needs electron donors (organics) for nitrogen removal by denitrifiers (Rathnaweera et al. 2018). Those treatment processes and units have advantages and disadvantages. They consume energy and produce a huge quantity of biosolids, which need further treatment for the reduction of organic and pathogenic organisms. Recently down flow hanging sponge (DHS) system was employed for wastewater treatment achieving a high-quality effluent with minimum energy requirements. The technology was efficiently used for the treatment of pre-settled sewage (Mahmoud et al. 2010), anaerobic effluent (Tawfik & Rifaat 2011), and agricultural drainage water (ADW) (Fleifle et al. 2014). Industrial wastewater treatment using DHS was attempted by El-Kamah et al. (2011), and the module was very effective for the removal of COD from the onion industry (El-Kamah et al. 2011), juice wastewater (El-Kamah et al. 2010), food industry wastewater (Ali et al. 2017), textile wastewater (Tawfik et al. 2014), and landfill leachate (Ismail & Tawfik 2016).

Anaerobic treatment of wastewater is the most acceptable technology due to energy production in the form of methane (CH₄) and/or hydrogen (H₂). Anaerobes are classified into acidogenesis for conversion of organics into volatile fatty acids (VFAs) and hydrogen (H₂) gas. The subsequent step is the degradation of VFAs and H₂ into methane (CH₄) by methanogenesis. Numerous anaerobic technologies, i.e., up-flow anaerobic sludge blanket (UASB) reactor, anaerobic baffled reactor (ABR), completely stirred tank reactor (CSTR), up-flow anaerobic staged reactor (UASR), and anaerobic sequential batch reactor (an-SBR) were employed for treatment of domestic and industrial wastewater (Elreedy et al. 2016). In the petrochemical industry, wastewater is used for 3-biofuels (H₂, Et-OH, and CH₄) production via an anaerobic packed bed baffled reactor supplied with anaerobes. Anaerobic packed reactor with polyurethane foam for simultaneous treatment of saline wastewater and energy production exists (Ali et al. 2019). The ammonia removal from urea industry wastewater was achieved by an ABR supplied with anammox bacteria (Ismail et al. 2019). Gelatinous wastewater-rich protein was used as a low-cost substrate for H₂ and CH₄ production using up-flow staged reactor and circular ABR (Mostafa et al. 2017; Meky et al. 2020). Farghaly and Tawfik (Farghaly & Tawfik 2017) successfully employed a multistage anaerobic digester for H₂ generation from paper mill industry wastewater.

Water contamination by heavy metals has seriously increased due to the intentional discharge of industrial wastewater into the environment (Osama et al. 2020). Natural treatment systems such as duckweed ponds (Osama et al. 2020) and constructed wetlands (Meky et al. 2019) are mainly employed to uptake and remove heavy metals from polluted water. Three ponds connected in series and supplied with algae and duckweed (Lemna minor) were operated at an HRT of 7 days (Sekomo et al. 2012) and achieved 94–98% for Cr removal. Zn removal was approximately 70% at a low loading rate and dropped to < 40% at a high loading rate. Pb, Cd, and Cu removal efficiencies are 36, 33, and 21%, respectively. The Cd, Pb, Cu, Zn, Ni, Cr, and As, were removed by values of 64.2, 48.7, 70.0, 93.9, 63.6, 63.8, and –236.2% in a horizontal subsurface flow-constructed wetland unit planted with Phragmites australis (Šíma et al. 2019). Wetland plants, i.e., P. australis and Phalaris arundinacea utilize the oxygen loss from their roots to oxygenate rhizosphere sediments where Fe²⁺ is easily oxidized and subsequently precipitated as ferric oxyhydroxides (Holcova 2006; Meky et al. 2019) in the belowground zones. The precipitates play a role in the scavenging of other metal ions by co-precipitation and/or adsorption in the plaque soil matrix. The dissolved metals are taken up by living plants. However, the highest metal concentrations are allocated in the belowground biomass and the lowest residual values are in the aboveground zones of wetland plants (Vymazal et al. 2009). The hydrophyte species of cattail, water hyacinth, duckweed, and water cabbage removed 96.2, 72.2, 60.4, and 93.3%, for Cd, 83.6, 82.3, 90.0 and 81.7% for Cu and 95.9, 78.0, 91.3, and 97.1% for Pb, respectively (Ayaz et al. 2020).

Chemical impurities like heavy metals, pesticides, dyes, and solvents enter the water bodies through agriculture/industrial effluents, wastewater plants, and households due to lax enforcement of rules. These impurities have chronic effects on the environment that can only be detected after a long period of their discharge (Schindler & John 2006). One of the effective mechanics of degrading the harmful effect of these compounds is through biodegradation, as discussed in Section 2.3.1. Despite the effectiveness of biological treatment methods, many organic compounds produced by the chemical industry are resistant to biological treatments. This requires some form of chemical oxidation such as wet oxidation, wet air oxidation (with and without catalyst), supercritical water oxidation, and advanced oxidation process (AOP). These processes degrade organic pollutants by forming hydroxyl radicals (García et al. 2006) and are recognized as highly efficient for recalcitrant wastewater (Oller et al. 2011).

AOP was typically used to treat industrial wastewater having high-load pollutants in MBRs. The membrane reactors and the MBRs can work in combination with different types of oxidation processes simultaneously, i.e., hybrid reactors (Du et al. 2021). Many different combinations such as wet air oxidation process, Fenton process, hydrogen peroxide, and ozonation are nowadays is in used by different countries for wastewater treatment (Chen et al. 2021; Hube & Bing 2021; Shi et al. 2021). In combination with the pretreatment, AOP shows a promising future compared to the individual process in terms of economy and area requirements. Such a combination may reduce total energy consumption, thereby overall cost (Jiang et al. 2021; Saba et al. 2021; Zheng et al. 2021). Effluent toxicity is also an essential factor after treatment, diminishing the effect of such systems. The balancing effect of combination leads to a more promising future (Keskin et al. 2021). Various reactive oxygen species (ROS) and hydroxyl radicals have oxidative capacity and high reactivity toward the organic contaminant in wastewater during AOP treatment. AOP process can be divided into different classes in terms of ROS generated during treatment, such as photocatalytic oxidation, ozone oxidation, and electrochemical oxidation (Singh et al. 2021). The main emphasis while using the AOP process is the high generation of ROS and low generation of harmful organics byproducts. As emerging technology clubbed with different physical processes, AOP has various advantages of less or no generation of secondary pollutants and a high capacity to mineralize the organics. But at the same time, it also deals with the physical limitations during its application (Cai et al. 2021; Gutiérrez et al. 2021; Hansen et al. 2021; Srivastava et al. 2021). A few factors such as a wide range of applications and non-conducive reaction conditions lead to high treatment costs. To achieve high pollutant degradation efficiency, Fenton oxidation technology requires an environment pH of less than 3.

In ozone oxidation technology the oxidation capability is better enhanced by the alkaline environment (Gaurav et al. 2021; Hafeez et al. 2021; Ma et al. 2021). The adjustment of the pH value of organic wastewater will raise treatment costs in order to get better treatment results. The development of efficient and cost-effective catalysts is important to photocatalytic oxidation and catalytic wet air oxidation technology (Gopinath et al. 2021; Januário et al. 2021; Khraisheh et al. 2021). It saves technical expenditures gives good treatment outcomes, and efficient and economic catalysts. Furthermore, the high temperature and pressure requirements of the catalytic wet air oxidation technology raise treatment costs and make the reaction vessel very demanding. A combination of multiple AOPs has been created to improve organic degrading efficiency in order to solve the high operating expenses and restricted efficiency of a single technology (Sandoval & Ricardo 2021).

The MF process has matured, thanks to the increased membrane production and improved automated systems (Hegab et al. 2017). To get the best results, a hybrid MF process is recommended for water treatment as a stand-alone unit, which often remains inefficient in handling the load from a commercial wastewater system or seawater desalination. The capital and operating cost of the MF module is still a challenge, particularly for low-income countries where the skills of the people for operating such technology are quite poor. Despite all the challenges, more and more countries are adopting MF for water treatment; for instance, the Sultanate of Oman (an arid country) with more than 402 sewage treatment plants utilizes tertiary wastewater technology and membrane bioreactors (MBRs) and UF to alleviate the acute water shortage problem (Jaffar et al. 2017).

The total cost associated with membrane biological treatment setup includes the initial investment for design and dedicated construction like drainage pipe systems and many other related engineering as well as non-engineering aspects (Das et al. 2021; Kishor et al. 2021). It has been seen that cost lies in the range of 800–2,200 USD (m³ day⁻¹) on an average. The total investment can be high considering the pollution standards of a particular area/city with a lower effluent quality concentration in industrial wastewater (Mekonnen & Arjen 2016; Tu et al. 2018). In many cases, it was seen that integrated and longer processes can be adopted to tackle this situation. MBRs are more economical and advantageous if constructed underground in terms of footprint reduction, and these will lead to increasing the overall treatment capacity (Asif & Zhenghua 2021; Wang et al. 2021a). The membrane process can be good in order to reduce the long-term cost over benefits for treatment projects (Pan et al. 2021; Ibrahim et al. 2021; Wu et al. 2021). Additionally, considering the long-term use of such a treatment process can also reduce the production cost of components of membrane and its elements, thereby lowering the cost of membranes (Cai et al. 2021; Gutiérrez et al. 2021; Hansen et al. 2021; Khan et al. 2021). Further, solar energy employment could be crucial to minimize the overall cost of desalination plants. Solar energy can be shared with the power load to run the important components of membrane/thermal desalination plants such as pumps, motors, pressure exchangers, and evaporators. Solar PV panels can eliminate the requirement for conventional electricity.

Further, utilizing solar energy for membrane desalination could be promising to make it economically attractive since it covers the significant cost of (in terms of power used) running the high-pressure pumps during RO. For instance, a typical RO-driven seawater desalination plant located in Egypt reportedly consumes around 6.5–9 kWh/m³ pump electricity, which can be easily covered with a small-scale solar PV cell system (typically 300 W solar panels with 40 panels that can produce 30–45 kWh electricity). Further, the RO unit located in Greece for water desalination reportedly consumes 5 kWh/m³ pump electricity which can be compensated using solar PV panels. Moreover, solar collectors and PV panels can also be integrated with the MD technology to make it economically recognized. The study has demonstrated the potential of a solar-powered air gap membrane distillation (AGMD) approach for desalinating seawater. The AGMD achieved up to 85 °C temperature with 1,002 W/m² °C solar radiation during peak summer. By maintaining the inlet temperature of the AGMD module to 73 °C, the water purification increased up to 1.62 kg/m² h). The requirement of temperature is fulfilled using two PV panels (1.6 m² area/panel and 300 W output power integrated with three batteries (12 V, 200 Ah)).

Energy reduction is an important aspect of the wastewater/desalination process and the construction of integrated industrial parks can achieve a lower SEC enhancing individual treatment capacity. Reducing aeration time and membrane fouling can improve wastewater treatment efficiency. A centralized wastewater treatment setup beside industrial and desalination zones can reduce facility footprint and investment costs. It is anticipated that the integration of these facilities (because of the water production and use) will bring overall energy minimization to the process. Wastewater treatment produces energy in the form of methane and hydrogen that can be used as energy for desalination to contribute positively to the overall energy minimization of the process avoiding fossil energy used for desalination.

The advances in desalination technologies keep propelling the demand for desalination plants with different types of nanoparticles (CeO₂, UiO-66, CS-NPs) and carboxylated multi-walled CNTs into a PA provides high salt rejection (as high as 90%) with excellent selectivity and permeability. Further, the use of nanoporous reduced rGO membranes provided 27.7–62.6% desalination rates and up to 96.8% salt rejection, along with protection against membrane fouling. Thermal desalination may be promising since it reduces the reliance on fossil energy and significantly minimizes GHG emissions (up to 94%). Utilizing solar energy for membrane desalination could cover the running cost of high-pressure pumps. A 300 W solar panel with 40 panels can produce 30–45 kWh which can easily run three RO pumps typically needing 6.5–9 kWh/m³ pump electricity. Solar collectors and PV panels can also be integrated with the MD technology, using AGMD water purification up to 1.62 kg/m² h can be achieved.

M. S. K. conceptualized the whole article, prepared the initial draft, and edited and revised the article; A. F. rendered support in membrane desalination writing and editing, K. B. A. conceptualized the whole article, worked on solar desalination, and revised and finalized the manuscript; M. K. A. M. and M. D. worked on membrane desalination challenges, and revised and edited the article; A. T. worked on nanotechnology and carbon-based desalination, and revised and edited the article; U. M. worked on wastewater treatments, and revised and edited the article.

The authors extend their appreciation to the Deanship of Scientific Research at King Khalid University for funding this work through Large groups (project under grant number RGP. 2/512/44).

All relevant data are included in the paper or its Supplementary Information.

The authors declare there is no conflict.