A review on advances and mechanism for the phycoremediation of cadmium contaminated wastewater

1. Introduction

Globally, the need for clean water has been cited among the top unprecedented vivacious concerns (Goswami et al., 2021). At present, above 20% of the entire world population is devoid of clean and innocuous drinking water (Goswami et al., 2018), and according to a report by the U.N. World Water Development Agency, this issue will be even grimmer in the upcoming future. Further, the condition is prophesied to be even more deteriorating because of the upsurge in the anthropogenic actions and direct disposal of environmental contaminants into the freshwater streamlines (Gupt et al., 2021a, Gupt et al., 2021b). Today's modern era is based on technology and industrialization to produce various commodities for human needs (Devi et al., 2021). For manufacturing these goods, industries require various raw materials, chemicals, and machinery to process finished products (Borah et al., 2021). The processing of these materials leads to the further release of complex wastewater possessing various chemical pollutants or toxic industrial waste into the environment (Gupta et al., 2020). The chemical pollutants comprise heavy metals, flammable and putrescible substances, lethal wastes, explosives, and refinery products (Kushwaha et al., 2016). These chemical pollutants impose deleterious health issues on all flora and fauna and cause environmental quality deterioration (Lata et al., 2021).

Cadmium (Cd2+) is one of the most notorious toxic environmental contaminants whose extensive practice for industrial purposes has initiated various issues related to the environment and human health globally (Bind et al., 2019). Based on toxicity, European Water Framework Directive has positioned Cd into the list of toxic chemicals (Oosterhuis et al., 2000). The guidelines for Cd in drinking water is 3 μg/L (WHO, 2017). The maximum limit for Cd set by USEPA and China is 5 μg/L. Earlier, various researchers studied the Cd occurrence, behavior, bioavailability, and its remediation in soil and groundwater (Bind et al., 2018).

Regardless of various monitoring steps being implemented to restrict the release of Cd2+ in the environment, pollution levels are still alarming (Sachan et al., 2021). Various conventional methods viz., chemical precipitation, chemical oxidation or reduction, coagulation, evaporation, ion-exchange, reverse osmosis, electrochemical, solvent extraction, membrane filtration, cementation, and adsorption by activated carbon have been applied for the remediation of HMs (Goswami et al., 2017). However, these methods are expensive, inefficient, require a large amount of chemicals, leads to the generation of secondary pollutants, and cause obliteration of soil fertility (Kushwaha et al., 2021b).

Research has been enduring to develop a highly efficient and cost-effective method for remediation of Cd2+ contaminated wastewater (Sathe et al., 2021). Microbes have the ability to degrade various organic and inorganic pollutants but are unable to degrade metals (Singh et al., 2019). Still, they have the ability to alter metal mobility in the environment via altering their physical and chemical features (Dutta et al., 2019). Due to the various anthropogenic activities such as industrial waste discharge, the mining process resulted in the deposition of metals in the surrounding (Kushwaha et al., 2015). Microbes gather the ability to thrive in such critical surroundings due to specific resistance mechanisms (Kushwaha et al., 2021a). It has been reported that under the metal stress condition, microbial structure, community, biomass, activity, and diversity got severely affected (Goswami et al., 2020). Microbes which are capable of enduring toxic environment, have more ability to persist/survive (Sathe et al., 2020). According to metal speciation, the survival strategy of microbes is due to biochemical, structural, physiological, and genetic mutability (Kushwaha et al., 2017). To overcome the stress of HM, microbes adapt a variable resistance mechanism (i.e., accumulation, sorption, mineralization, and transformation) (Kushwaha et al., 2018).

Biosorption is a well-known technique utilizing the live or dead microbes for the HMs or other pollutants removal (Rizvi et al., 2020). The use of biosorbents is cost-effective, and often cell wall of live microorganisms can be reconstructed after absorption (Yadav et al., 2021a). It is energy independent process; HM ions are generally adsorbed on the biosorbents active sites or on the cell surface (Kushwaha et al., 2020). Bacteria, fungi, algae, plant residues, fruits, animals’ skin, biopolymers, coal fly ash, and active sludge can be used as an adsorbent (Gautam et al., 2021). Commercially available biosorbent should have the following properties: (1) proper shape, size, and physical properties, (2) high adsorption capacity, (3) removal of biosorbents from solution should be cost-effective, (4) high thermal stability and mechanical strength, (5) readily available, reusable, and regeneration property, and (6) no usage of chemicals (Kumar et al., 2018).

Amongst all microbes, algae are natural renewable biomass and show varied affinity towards HMs, and thus, are a promising candidate and can be employed as biosorbent (Kumar et al., 2016). The use of micro- and macroalgae for remediation of pollutants is termed as “phycoremediation” (Jais et al., 2017). Microalgae are known to have high photosynthetic activity and grow under extremely stressed conditions such as high salinity, temperature, osmotic pressure, reduced nutrient condition, and HMs (Kumar et al., 2020). Therefore, Cd2+ removal by microalgae is an eco-friendly approach and a simple and robust process with high growth rate, and generation of value-added products (Kumar et al., 2014). Presently, writing to this review, the keywords search analysis in Web of Science on the keywords “Microalgae” and “Cd” resulted in 144 (accessed on November 01, 2020) for all the documents in the past five years (2016–2020) with at least five minimum number of occurrences of the associated keywords (Fig. 1). This review focus on Cd, its industrial usage, availability, and toxicity, along with the existing remediation techniques. Further, discussing the microalgal mechanism for Cd2+ removal and implementation of microalgae to treat industrial effluent wastewater.

2. Cadmium

Cd2+ was discovered as a zinc carbonate impurity belonging to IIB group element. It is a rare earth element with 0.15 mg/kg abundance in Earth's crust and 1.1 × 10−4 mg/L in the sea (Hayes, 2016). Cd2+ is one of the ubiquitously dispersed noxious and mobile pollutants in soil. It is most commonly present in the +2 oxidation state. Generally, Cd is soft; therefore, alloying with Zn, it gets improved. Annually, ∼13,000 tons of Cd2+ are produced worldwide (Orisakwe, 2012) specifically for Ni–Cd batteries, as a chemical stabilizer, pigments, alloying, and coating. It is less soluble in sulphuric and hydrochloric acids and readily in nitric acid. Cd2+ salts with strong acids are readily soluble in water. Fig. 2 represents the occurrence of Cd2+ concentrations in soils, sediments, and water, globally.

Consequently, a high amount of Cd2+ is continuously being released, leading to increased concentration within the environment. Cd2+ has a similar ionic radius, chemical nature, and identical charge as calcium. Therefore, it replaces calcium and enters the human body, and gets accumulated in various organs at a high level (Kubier et al., 2019). Irrigation of agricultural land with Cd-contaminated water leads to bioaccumulation into the crops. It has been observed that in China, Jamaica, and Korea, the elevated level of Cd in soil and groundwater leads to enhanced bioaccumulation of Cd in rice (Liu et al., 2017). Hence, in the human body, Cd2+ enters mainly via terrestrial route, i.e., consuming vegetables. In 1984, UNEP proposed a regulation for twelve hazardous pollutants globally, amongst which Cd2+ was placed first. Along with, WHO and IARC have also listed Cd in category one, i.e., a carcinogen (Liao et al., 2015). Cd2+ is known to have no function in plants, animals, and humans; however, some literature reports essentiality in animals. When Cd2+ is present in high concentration, severe effects have been observed in digestive and respiratory tracts and predominately accumulates in kidneys with a 10–20 years’ biological half-life. Long-term exposure to Cd2+ leads to irreversible renal effects.

2.1. Soil water and groundwater

Cd2+ in soil, water and groundwater usually occur up to 5 μg/L and 1 μg/L, respectively (Smolders and Mertens, 2013). Naseem et al. (2014) found 10 μg/L Cd2+ content in groundwater instigated from Jurassic sulfide-bearing sedimentary rocks in Pakistan. Duijnisveld et al. (2018) reported 0.11 μg/L (loess aquifers) to 2.7 μg/L (sandy aquifers) in Germany. Tedd et al. (2017) studied the Irish groundwater and found 0.2 μg/L Cd2+ concentration that further increased to 0.5 μg/L in groundwater originated for non-calcareous sediments. Table 1 shows various types of Cd2+ pollution in soil and groundwater.

Table 1. Types of cadmium pollution in soil and groundwater.

Source	Pollution type	Maximum Cd level	Reference
Mining
Pb–Zn mining/refinery	Wastewater	Ground water: 77 μg/L	Paulson (1997)
Cu mining	Wastewater	Soil: 499 mg/kg	Bech et al. (1997)
Au–Cu mining	Wastewater	Soil:121.5 mg/kg	Avkopashvili et al. (2017)
Au–Ag–Pb–Zn mining	Wastewater	Ground water:19 μg/L	Rosner (1998)
Industry
Metal industry	Atmospheric deposition	Groundwater: 74 μg/L	Dwivedi and Vankar (2014)
Textile industry	Wastewater	Soil: 83.6 mg/kg Groundwater: 40 μg/L	Deepali and Gangwar (2010)
Waste management
Landfill	Leachate	Soil: 378 mg/kg	Chen and Liu (2006)
Landfill	Leachate	Groundwater: 51 μg/L	Abd El-Salam and Abu-Zuid (2015)
Brownfield	Wastewater	Groundwater: 474 μg/L	Li et al. (2017)
Electronic waste recycling	Wastewater	Soil: 47.7 mg/kg Groundwater: 280 μg/L	Panwar and Ahmed (2018)
Household wastes	Wastewater	Groundwater: 580 μg/L	Ololade et al. (2009)
Sewage and waste disposal	Wastewater	Groundwater: 90 μg/L	Affum et al. (2015)
Disposal facilities	Wastewater	Soil: 32 mg/kg	Beyer and Stafford (1993)
Agriculture
P fertilizer production	Atmospheric deposition	Soil: 9.3 mg/kg Groundwater: 3 μg/L	Mirlean and Roisenberg (2006)
P fertilizer application	Infiltration	Groundwater: 60 μg/L	Vetrimurugan et al. (2017)
Sewage sludge application	Irrigation	Soil: 90 mg/kg	Moral et al. (2005)
Urban areas
Road traffic	Infiltration	Groundwater: 2.34 μg/L	Wessolek and Kocher (2002)
Sewerage	Leakage	Groundwater: 5 μg/L	Eiswirth and Hotzl (1997)

Groundwater in sandstone, unconsolidated sand, and gravel aquifer systems (USA) were detected to be > 1 μg/L Cd2+ concentration (Ayotte et al., 2011). Cd2+ content in groundwater from glacial aquifer system, USA was in the range 0.018–1.0 μg/L; still, Cd2+ content in 84% of samples was beneath the detection limit (Groschen, 2009). In the USA, Cd2+ concentrations were up to 6000 μg/L in groundwater near waste sites (ATSDR, 2012). Cd2+ level was up to 2700 μg/L in municipal solid waste in European Union (EU, 2007). Thus, it can be concluded from studies that Cd2+ concentrations are regarded as having a natural background comprising both natural and anthropogenic activities.

2.2. Natural sources

2.2.1. Air

Naturally, Cd2+ concentration occurs due to the genesis of rocks, airborne soil particles arising from forest fires, desert, volcanoes, biogenic materials, meteoric dust, and hydrothermal vents (ATSDR, 2012). Sources of Cd2+ due to anthropogenic activities are fossil fuel combustion, cement production, steel, iron, manufacturing of phosphate containing fertilizers, production of non-ferrous metal, incineration of sewage and municipal sludge, and road dust (ATSDR, 2012). In California, wildfire has enhanced the Cd2+ content in stormwater by two folds (Burke et al., 2013). Cd2+ concentration is also increased in the soil by using biomass ash (containing 30 mg/kg Cd2+ concentration) as fertilizers.

Initially, the bioavailability of Cd2+ is low due to an increase in pH by ash and thus, have high adsorption; however, with time, bioavailability is increased due to the decrease in pH and rainfall (Li et al., 2016). In Germany (2014), the total Cd deposition was 12.8 t, out of which 61% deposition was mainly due to the natural emission (Ilyin et al., 2016). Fig. 3 shows the anthropogenic and natural sources of Cd2+ and the biogeochemical cycle showing Cd2+ transportation in the atmosphere.

2.2.2. Rocks and soils

Generally, in sedimentary rocks Cd2+ concentration (0.01–2.6 mg/kg) is greater than metamorphic (0.11–1.0 mg/kg) and igneous rocks (0.07–0.25 mg/kg) (Smolders and Mertens, 2013). Cd2+ has a similar ionic radius to various metals (such as Fe, Ca, Pb, Zn, and Co) and can replace them in various minerals (phosphate and carbonate-containing rocks) (Smolders and Mertens, 2013). For example, in sphalerite (ZnS), Cd2+ can substitute Zn; in pyrite (FeS2), Cd2+ replaces Fe (Tabelin et al., 2018). Moreover, Cd2+ is also adsorbed by Fe (II) hydrous oxide, which plays a significant role in Cd2+ contribution in the aqueous phase with change in redox conditions (Hindersmann and Mansfeldt, 2014). In phosphorites and phosphate minerals (used for fertilizer production), Cd2+ is present as an impurity which is due to the replacement of Ca (Kubier et al., 2019). However, depending on the geogenic process, Cd2+ concentration varies. In Nauru (Pacific island), the highest Cd2+ concentration (240 mg/kg P2O5) was observed (Mar and Okazaki, 2012). Black shales also contain high Cd2+ content due to enhanced biogenic enrichment and marine primary production (Liu et al., 2017). Therefore, black shale weathering can lead to enhanced Cd2+ concentration in the environment.

Worldwide, Cd2+ concentration in uncontaminated soil is 0.36 mg/kg, though concentration can vary depending on the region, country, and soil type. For example, average Cd2+ concentration is 0.27, 0.01, 0.18, 0.3, and 0.2 mg/kg in USA, Australia, Brazil, Japan, Europe, respectively (Taylor et al., 2016). Cd2+ concentration also varies depending on the soil taxonomy, Histols (0.62 mg/kg) and Aridisols (0.3 mg/kg) have the maximum Cd2+ concentration, whereas, Spodosols, Alfisols, and Ultisols contain lowest Cd2+ concentration with 0.2, 0.11, and 0.05 mg/kg, respectively (Holmgren et al., 1993).

Cd2+ content is also affected by the soil depth, i.e., Cd2+ concentration decrease with an increase in depth (Kubier et al., 2019). Cd2+ content was observed to be 0.06, 1.8, 1.88, and 2.0 mg/kg in sandy subsoils, soils at the swamp, soils above fluviatile, and sediments in the intertidal zone, respectively (LABO, 2017). Soil texture also influences the Cd2+ levels; with increasing clay and peat content in soils, Cd2+ content increases (LABO, 2017). In soil, Cd2+ concentration decreases with the increasing distance between urban areas and industries. Joimel et al. (2016) observed gradient Cd2+ concentration with regard to terrestrial use, in the range from 0.13 to 0.13 mg/kg in industrial, mining, military areas, garden, farming, grassland, vineyard, and forest.

2.2.3. Anthropogenic cadmium sources

Anthropogenic activities such as fossil fuel combustion, sewage sludge, mining, metal and fertilizer manufacturing industries, landfills result in the release of Cd2+ in soil and groundwater (Goswami et al., 2019). Most commonly, the use of phosphate fertilizers releases a high amount of Cd2+ in the environment. Studies show that P fertilizers change the soil chemistry and get transferred to the food chain. The widespread release of Cd2+ in the environment is due to the reuse of wastewater, atmospheric emission, agricultural practices act as a diffuse source (ATSDR, 2012). Globally, the dumping of Ni–Cd batteries in landfills is the chief source of Cd2+ (Khan et al., 2017). Additionally, Cd2+ comprising products such as coatings, pigments, alloys, and stabilizers in polyvinyl chloride also contribute to Cd2+ pollution (Goswami et al., 2019).

2.3. Hydro-chemical behaviour

Cd2+ occurs as a divalent cation in an aqueous solution. Cd2+ mobility is affected by solution pH due to hydrolysis of metal, ion-pair formation, organic matter solubility, oxy-hydroxides surface charge, and organic matter. At pH < 6.5 and under aerobic condition Cd2+ remains in solution. However, with an increase in pH, metal adsorption or precipitation increases on mineral surfaces (Caporale and Violante, 2016). Under anoxic conditions and in the presence of sulfidic, Cd2+ forms a redox-sensitive aqueous complex (CdHS+) and precipitates sphalerite (ZnS), galena (PbS), and chalcopyrite (CuFeS2) that are stable containing trace amount of Cd2+ (Tabelin et al., 2018).

2.3.1. Solubility and complexation

The mobile fractions of Cd2+ include water-soluble Cd2+, organo-metallic complex, and unadsorbed Cd2+ (Loganathan et al., 2012). The adsorbed Cd2+ fraction consists of weakly and normal bound Cd2+ on mineral surfaces, and it is mainly responsible for variation in Cd2+ level in natural water bodies. Cd2+ bound with soil matrix or associated as surface complexes are considered as a stable fraction. Cd2+ complexes with anions (CdCl+ or CdSO40) and dissolved organic matter (DOM), which are water-soluble (Loganathan et al., 2012). Due to which CdCl+ exists in the aqueous phase while other HMs concentration decreases. Moreover, complexation with organic and inorganics causes Cd2+ dissolution from oxy-hydroxides, sulphides, or phosphates (Najafi and Jalali, 2015). The groundwater composition affects the nature of Cd2+; the total soluble Cd2+ (55–90%) exist as Ca2+ ions, while rest Cd2+ occurs in association with organic and inorganic such as CdCl+, CdCl3−, CdCl20, Cd (SO4)22−, CdSO40, CdHCO3−, CdCO30, Cd (CO3)22−, CdOH+, Cd (OH)20, Cd (OH)3-, Cd2OH3+, CdNO3+ (Baun and Christensen, 2004). Cd2+ occurs as Cd (HS)20, Cd (HS)3-, Cd (HS)42−, and CdHS+ under sulfidic and aerobic conditions. In addition, Cd2+ forms the most stable complexes with carbonate, bisulfide, sulphate, and chloride anions.

The change from reducing to oxidizing conditions releases Cd2+ present in sulphide minerals or complexed with organic matter. The shift in the redox environment causes mineralization of organic matter, and sulphide gets dissolved (Tabelin et al., 2018). The change in groundwater conditions, i.e., reducing to the oxidizing environment, results in the release of Cd2+. For example, under oxidizing conditions, Cd2+ and HMs are released from volatile acid sulphide (greigite, amorphous FeS, or mackinawite) and pyrite (Zhang et al., 2014). Cd2+ and other HM concentrations in floodplain soils are affected by dissolved Fe, S, and Mn concentrations. The change in the redox-sensitive elements is affected by pH, dissolved organic carbon (DOC), and redox potential (Shaheen et al., 2014). The formation of the Cd2+ complex is too altered via ionic strength and other competing cations, viz. Ca or other HMs (Beisecker et al., 2012).

In naturally occurring organic matter, phenolic and carboxylic groups occur abundantly; but thiols and amines (less abundant) present in organic matter form stable complexes with Cd2+ (Kanamarlapudi et al., 2018). In disparity to the other HMs, which form complex with acidic hydrophilic and hydrophilic DOM, Cd2+ forms complexes with the neutral hydrophilic DOM fraction (Kozyatnyk et al., 2016). Generally, Cd2+ binds with low molecular weight (LMW) DOM in contrast to most HMs bound with high molecular weight DOM. The interaction of Cd2+ with LMW DOM is weaker and is easily replaced by other HMs (Kozyatnyk et al., 2016).

2.3.2. Sorption

In the aquatic environment, solubility and mobility of Cd2+ are influenced by pH, DOC, dissolved inorganic carbon (DIC), and occurrence of oxy-hydroxides (Mn, Fe, and Al) and clay (Li et al., 2016). In groundwater, co-precipitation, and sorption often governs the Cd2+ concentration (Carrillo-Gonzalez et al., 2006). The correlation between Kd value of Cd2+ with various factors was found to be in the following order: organic matter < clay < oxy-hydroxides < cation exchange capacity < pH. Loganathan et al. (2012) reported that with the increase in ionic strength, the sorption of Cd2+ decreases due to the competition with other cations. It reduces Cd2+ activity, complexation of an ion with lower sorption affinity, reduced pH, and variation in electrostatic potential. Cd2+ sorption can be either exothermic or endothermic, depending upon the temperature (Karak et al., 2015). Cd2+ is adsorbed onto the negative charge carrying surfaces unless it exceeds the point of zero charge. Thus, the pH of groundwater influences and controls the binding site availability with the aquifer.

Generally, sorption of Cd2+ on mineral surfaces is a two-stage sorption model. Initially, at highly energetic binding sites, Cd2+ is adsorbed, and in the second stage, Cd2+ diffusion into the mineral surface is slow and time-dependent (Loganathan et al., 2012). According to Smolders and Mertens (2013), sorption of Cd2+ with surface functional groups is as follows: $S - O H + C d^{2 +} (a q) \leftrightarrow S - O C d^{+} + H^{+}$

In solution, fixation of Cd2+ at a high Cd level occurs via co-precipitation, while at low Cd2+ content, chemisorption occurs (Ahmed et al., 2008). In solution, Cd2+ precipitates along with calcite. Primarily, Cd2+ is adsorbed quickly onto the calcite surface, and then Cd2+ diffusion into the crystal lattice slows down, resulting in the generation of CdCO3. This process is affected by solution pH and competition with Mg2+. $C d^{2 +} + C a C O_{3} = C d C O_{3 (s)} + C a^{2 +}$

Cd2+ sorption on Mn, Fe, and Al-oxides occurs by ion exchange with the surface OH functional group (Loganathan et al., 2012). In P fertilizers, Cd2+ substitution or sorption occurs, e.g., Cd2+ lead phosphate hydroxide ((CdPb9 (PO4)6 (OH)2), fluorapatite (Ca10 (PO4)6 F2), gypsum (CaSO4), and tricalcium phosphate (Ca3 (PO4)2) (Azzi et al., 2017). But, Cd2+ mobility depends on the P concentration, i.e., at low P concentration, Cd2+ mobility is enhanced due to increased ionic strength and reduced pH (Grant, 2018).

2.3.3. Competition

Cd2+ forms soluble complexes with chloride with varying charges such as CdCl20, CdCl+, CdCl42−, CdCl3−, which significantly decreases the Cd2+ sorption (Núñez-Delgado et al., 2017). Cd2+ sorption was significantly enhanced in the presence of ligands such as phosphate and sulphate due to cation exchange capacity and removal of competing Ca. Studies suggested that Cd2+ is generally bound with an easily soluble fraction of sediments and soils, whereas other HMs (Cu and Pb) bound strongly with sulfidic, organic, and residual fractions. The order of affinity for HMs with organic matter is as follows: Cu2+ > Cd2+ > Fe2+ > Pb2+ > Ni2+ > Co2+ > Mn2+ > Zn2+ (Bolan et al., 2003). The presence of Cl− and NO3− reduces the Cd2+ and forms soluble complexes with chloride with varying charges such as CdCl20, CdCl+, CdCl42−, CdCl3−), which significantly decreases the Cd2+ sorption (Núñez-Delgado et al., 2017).

2.4. Cadmium toxicity

Exposure to Cd2+ leads to Cu, Zn, and Mg imbalance in the body and urine (Genchi et al., 2020). Enhanced Cd2+ level in the body causes anosmia, renal failure, neurological disorder problems in the male reproductive system, osteal disorders, and cancer (de Queiroz and Waissmann, 2006). People working in the direct contact with Cd2+ such as smelters, miners, and battery reclamation have a high degree of developing prostate cancer (Tchounwou et al., 2012). A study showed that individuals working in jewelry-making companies had elevated Cd2+ content above the permissible limit and had chest pain, anxiety, and breathing problems. Exposure to Cd2+ has similar nephrotoxic effects on both men and women, but women and kids are more susceptible to Cd2+ exposure. Due to the lesser iron storage and up-regulation of Fe channels in women and kids, it leads to high Cd2+ uptake since both Cd2+ and Fe have analogous uptake mechanisms (Chaumont et al., 2011). It is also believed that Cd2+ affects the neurogenesis in adults via the JNK and p38 AP kinase pathway that results in cell apoptosis (Wang et al., 2017a, Wang et al., 2017b).

Workers working in the Cd-rich environment suffer from urolithiasis (kidney stones) and with the progression of glomerular and renal failure (Guo et al., 2018). Few cases have reported the renal failure leads to renal tubular damage-causing uremia or death. The initial renal tubular impairment can be detected by the presence of proteins N-acetyl-β-D-glucosaminidase (NAG), retinol-binding protein (RBP), and α-1-microglobulin in the lysosomes of proximal tubular epithelial cells. The increase in NAG level is often correlated with an increase in Cd2+ level in urine. Additionally, an increase in RBP and β2-microglobulin was observed in Cd2+ level in urine (Chaumont et al., 2011). Fig. 4 shows the mechanisms of Cd2+ toxicity at the cellular level.

In the human body, the liver is one of the vulnerable organs. The exposure of the liver to Cd2+ causes swelling in hepatocytes and damage to lysosomes (membrane-bound), thereby induces hepatotoxicity. Further, Cd2+ exposure also causes impairment to mitochondrial cristae, infiltration in inflammatory cells, and hypertrophy of Kupffer cells (Adikwu et al., 2013). Studies showed that when normal and immortalized human liver cells were treated with Cd2+, loss of mitochondrial potential was observed along with caspase-9 and caspase-3 activation leading to cell apoptosis (Lasfer et al., 2008). Cd2+ toxicity on bone was first reported in the Jinzu river basin area, Japan, in the 1940s, when people consumed rice grown on Cd2+ polluted soil. People started developing bone fractures, malformation of bones, and bone pain, known as an itai-itai disease (Inaba et al. 2005). Cd2+ damages the renal Ca, PO4 transport, metabolism of vitamin D, and parathyroid hormone (Ogawa et al., 2004). Sughis et al. (2011) found that Cd2+ causes bone resorption even in young children, indicating the osteotoxic effect with calciuria. Researchers also observed that Cd2+ does not affect at low concentration but has a toxic impact on bones.

In the lungs, Cd2+ enters via house dust, cigarette smoke, or exposure from construction sites. In human lungs, 40–60% of Cd2+ is adsorbed via tobacco smoking (Ganguly et al., 2018). The study revealed that 55-year old smokers contain 30 mg of Cd2+, while in non-smokers, 15 mg Cd2+ was observed. The blood of smokers has 4–5 times more Cd2+ than non-smokers (Kazantzis, 2004). When exposed to Cd2+ > 5 mg/m3, it causes damage to lung epithelial cells, and long-term Cd2+ exposure results in destruction to the nasal epithelium and olfactory function (Faroon et al.,2013). Cd2+ causes inflammation in the lungs influencing the synthesis and release of various signaling molecules such as cytokines, adhesion molecules, and leukotrienes/prostaglandins (Huang et al., 2017).

Initial reports showed the Cd2+ effect on the nervous system, including vertigo, headache, and sleep disturbance. At the cellular level, Cd2+ causes oxidative stress and DNA damage. The exposure of human fibroblasts to Cd2+ leads to G2 arrest (Cao et al., 2007). The cell cycle arrest exhibits up-regulation cyclin kinase inhibitors (p27 and p21), hindering cell progression via cyclin-dependent kinase inhibition leading to DNA damage (Aimola et al., 2012).

Cd2+ also affects the function of various proteins at the cellular level thus inhibiting DNA synthesis leading to DNA damage. Cd2+ replaces Zn in the zinc-finger domain exhibiting errors in DNA repair mechanism leading to DNA lesions, increasing cancer risk (Cao et al., 2007). The exposure of cells to Cd2+ and Ni results in mutagenic non-allelic recombination. Madejczyk et al. (2015) exposed rat cells to Cd2+ and observed generation of OH- radicals causing temporal changes in DNA. Jianhua et al. (2006) observed mutations in hypoxanthine-guanine phosphoribosyl transferase (biomarker for identifying DNA damages) on exposure to Cd2+.

In other organ systems, Cd2+ causes toxicity, such as atherosclerosis, resulting in cardiovascular diseases and enhanced progression (Messner and Bernhard, 2010). Cd2+ also enhances the generation of reactive oxygen species (ROS) and prevents anti-oxidative activity by binding with metallothionines (Unsal et al., 2020). Cd2+ can also cause atherosclerosis due to kidney damage, increased blood pressure (Caciari et al., 2013), estrogenic activity mediated by Cd2+ (Kluxen et al., 2012), and epigenetic changes (Wang et al., 2012).

3. Cadmium removal

From the last few decades, it is becoming clear that conventional methods to treat HM polluted water are too expensive. Conventional methods for treating HM contaminated wastewater comprising chemical precipitation, ion exchange, adsorption, membrane technologies, electrochemical treatment, coagulation-flocculation, evaporation recovery, etc. (Agarwal et al., 2020).

3.1. Precipitation

The most frequent process used for the removal of HM sulphides, hydroxides, or carbonates is precipitation. It is a simple and cost-effective method. The precipitation of Cd2+ can be performed by using lime, barium acetate, di-isobutyl di-thiophosphate, and Mg. Though precipitation is a cost-effective method, its effectiveness at low Cd2+ concentration is a challenge (Purkayastha et al., 2014).

3.2. Coagulation/flocculation

HM removal from wastewater is also carried out by coagulation and flocculation, trailed by sedimentation and filtration. Various coagulants such as ferric chloride, ferrous sulphate, and aluminium have been employed, resulting in the effective HM removal. Further, calcium hydroxide and aluminium sulphate are utilized for HM removal and both coagulants significantly decreased the Cd2+ concentration. Xu (2010) investigated Cd2+ removal by manganese coagulants and observed effective removal. The main drawback is the high amount of sludge generation and chemical usage.

3.3. Solvent extraction

This method is used to separate or remove HMs from highly concentrated solutions. This method is simple, rapid, and low-cost. Research has been carried out to separate Cd2+ and other metals using numerous extractants such as Aliquat 336, Cyanex 301, Cyphos IL, Cyanex 302, and Cyanex 272. Eyupoglu and Polat (2015) reported selective Cd2+ separation from acidic iodide solution by liquid-liquid solvent extraction system. Mahandra et al. (2017) carried out liquid-liquid extraction for Cd2+ separation from aqueous solution using Cyphos IL 104 phosphonium. Likewise, Cd2+ was removed from sulphate medium by Aliquat 336 and Cyphos IL 101 (Swain et al., 2016). However, during this process, a large amount of toxic solvent is generated, which can be a source of secondary pollutants. Additionally, at low Cd2+ concentration, this process is not applicable.

3.4. Ion-exchange

A chemical reaction occurs between an electrolyte present in a solution and an insoluble electrolyte. The main advantage of ion exchange for HM removal is its high efficiency, fast kinetics, and high treatment capacity. This method can be used to select specific metals or all metal ions from the contaminated aqueous system. Various resins such as dolomite, Amberlite IR120 and 718, and Dowex 50 W can be used for HM removal from solutions during the process. Koivula et al. (2000) studied the Cd2+ removal using S-950 chelation ion exchange resin and obtained 83.9% of Cd2+ removal. Fernandez et al. (2005) observed 99.5% removal efficiency when Amberlite IRC-718 was used. Kocaoba and Akcin (2005) studied the Cd2+ and Cr removal by Amberlite IR 120 (a strong cationic resin). Pehlivan and Altun (2006) investigated the Cd2+ removal on synthetic resin Dowex 50 W and obtained 97% removal. Kocaoba (2007) studied two resins Amberlite IR 120 and Dolamite, and observed 97.4% recovery for Amberlite IR 120. A good Cd2+ removal and recovery efficiency were observed in all studies, making this method suitable for Cd2+ removal. However, the primary disadvantage is the operational cost.

3.5. Electrochemical treatment

Electrochemical treatment techniques are widely applied for the HM treatment of contaminated wastewater due to its several advantages: rapid separation of metals, easy process control, requirement of few chemicals, good reduction yields, and low sludge generation. Nevertheless, its use is restricted due to high initial capital investment and the need to supply electricity, which is expensive. Thus, it is clear from the techniques above that it is essential to consider various factors, including the initial metal concentration, wastewater composition, capital investment, operational charge, plant flexibility and consistency, and environmental impact for handling HM contaminated wastewater.

Electrochemical systems include plating metal ions on a cathode surface and recovering metals in elemental form. HM can also be removed from wastewater using electrocoagulation. This process is simple, can be operated easily, less time is involved during operation, less sludge generation, and chemical usage is low (Chen, 2004). Electro-coagulation (EC) is one such electrochemical method that includes the generation of coagulants in-situ by dissolving either aluminium or iron ions from aluminium or iron electrodes electrically (Chen, 2004). The metal ion generation takes place at the anode, and hydrogen gas is released from the cathode. The hydrogen gas helps to separate the flocculated particles from the solution by making them easily float (Chen, 2004). Kobya (2010) treated electroplating water containing Cd-cyanide and Ni-cyanide in an electrochemical reactor consisting of an iron electrode and reported Cd2+ was efficiently and effectively removed from electroplating wastewater.

Electro-flotation (EF) is another electrochemical method in which solid/liquid separation occurs, and pollutants float to the liquid surface due to the generation of hydrogen and oxygen gases as tiny bubbles by water electrolysis. Electro-flotation has an extensive range of HM application removal from industrial wastewater (Koelmel et al., 2016).

Electro-deposition is another method that is regarded as a clean technology for recovering metals from the aqueous system with no secondary sludge generation (Issabayeva et al., 2006). Here the removal and recovery due to the applied potential or current. The HMs in different oxidation states become reduced at the cathode and are plated onto it (Koelmel et al., 2016).

3.6. Flotation

Flotation is a well-known process that has found widespread utilization in wastewater treatment. Ion flotation, dissolved air flotation (DAF), and precipitation flotation are the chief flotation methods utilized to remove metals from wastewater. The flotation method has been employed to separate HMs from a liquid phase using bubble attachment originating in mineral processing. In a typical DAF process, micro-bubbles of air are used to attach to the suspended particles in the water, allowing it to form agglomerates which further aid in forming less dense flocs that can be easily removed. The flotation method offers many advantages over several other traditional methods, viz. high metal selectivity, high removal efficiency, high overflow rates, low detention periods, low operating cost and production of more concentrated sludge (Rubio et al., 2002), but its use is limited due to high initial capital cost as well as high maintenance and operation costs.

3.7. Sulphate reducing bacteria-based treatment systems

Bioreactors can be categorized based on their retaining capacity of the microbes that achieve the anticipated chemical adaptations (Lens et al., 2002). Fig. 5 enlists the active and passive treatment schemes along with diverse reactor types applied for treating HM contaminated wastewater.

3.7.1. Active bioreactor systems

The sulfidogenic bioreactors imply a very diverse methodology for remediating the HM contaminated wastewater. A wide range of factors governs the anaerobic sulphate reduction procedure, viz. sulphate concentration, HM and species concentration, pH, temperature, carbon source/substrate/electron donor, etc.

The potential benefits of active biological treatment structures are a choosy recovery of HMs from the aqueous system and readily controllable act and significantly dropped sulphate concentrations. Henceforth, sulfidogenic bioreactor application for Cd2+ elimination is extensively functional on a large scale in relation to the passive treatment arrangements. On the downside, the construction, operation, and maintenance expenses of these arrangements are considerable (Kiran et al., 2017).

Various sulfidogenic reactors were engaged for biological sulphate reduction and metal removal and are described widely in the literature. Further, these sulfidogenic reactors are up-scaled from a lab to a field. The merits and limitations of such sulfidogenic bioreactors are represented in Table 2. Some of the salient characteristics of these bioreactors are described ahead.

Table 2. Main features of active and passive systems for biological treatment of heavy metal containing wastewater (Kaksonen and Puhakka, 2007).

Parameter	Active system	Passive system
Operation and maintenance expenditure	High	Low
Labour requirement	More	Less
Treatment area	Small	Large
Metal recovery	Possible	Difficult
Process control	Good	Poor
Degree of predictability	Good	Poor
Cost	High	Low

3.7.2. Passive biological treatment systems

It includes anoxic ponds, wetland systems, substrate injection into the subsurface, infiltration beds, etc. Water bodies viz. anoxic ponds are affixed with organic substrates. Mining wastewater remediation via sulphate reducing bacteria (SRB) engage an open pit as a large-scale basin (Riekkola-Vanhanen and Mustikkamaki, 1997). Liquid manure was utilized as a source of SRB, and press juice from silage was added as the potential electron donor. Riekkola-Vanhanen and Mustikkamaki (1997) described that the increase in the water pH by the occurrence of sulphate-reducing activity resulted in the decline of sulphate, Zn, Fe, and Mn concentrations and redox potential. The limitations of anoxic ponds are (a) prerequisite of a large land area, (b) requirement of appropriate sludge removal and treatment, and (c) inappropriate for cold weather environments (Varon and Mara, 2004).

For many decades, wetlands have assisted in the economic designs to augment the treated acid mine drainage (AMD) characteristics (Huntsman et al., 1978). Similarly, these designs have previously been conceded at the laboratory scale (Gazea et al., 1996). They are active for the HM elimination, radionuclides, and sulphate from mining wastewater on a large scale (Noller et al., 1994). Wetland structures are categorized as anaerobic and aerobic; the anaerobic system utilizes SRB for treating wastewater. The mechanisms through which HM bearing wastewater are treated in wetland systems comprise a wide range of systems, viz. adsorption, filtration, sedimentation, uptake by plant biomass, and precipitation of metals by geochemical processes (Stottmeister et al., 2003). Wetland treatment arrangements encounter specific problems, viz. it is unproductive in dry and semiarid climatic conditions and Cd2+ dissolution upon an acquaintance of metal sulphides to air in the drought seasons.

3.7.3. Membrane filtration

Membrane filtration (MF) utilizes various types of membranes and is found to be effective for Cd2+ removal due to easy operation, less space consumption, and increased efficiency. The crucial element is a semi-permeable membrane, and the most significant parameter is retention. Various factors such as the composition of the solution, temperature, pH, type of solute, membrane material, membrane composition and pore size, and hydrodynamics affect the retention. Cd2+ removal from an aqueous medium can be achieved by reverse osmosis (RO), nano-filtration (NF), ultrafiltration (UF), and electrodialysis.

RO and NF processes use membranes of different pore sizes as a separation method. RO is mainly adopted for treating wastewater as it suggests a prodigious deal in cost reduction and conserves natural resources. The advantages of NF are easy operation, less consumption of energy, high pollutant removal efficiency, and reliability. Qdais and Moussa (2004) investigated the removal of Cd and Cu by RO and NF from wastewater. They observed 98.5% and 82–97% Cd2+ removal efficiency using RO and NF, respectively, with an initial concentration of 25–250 ppm. Since, RO uses high pressure than NF; thus, NF is economically more suitable.

UF is carried out at low trans-membrane pressure to remove colloidal and dissolved material present in water. Subsequently, the pore size of the membrane used in UF is larger than Cd2+ ions; therefore, the HM ions can be easily pass through the membrane. Thus, improved micellar ultrafiltration (MEUF) and polymer boosted ultrafiltration (PEUF) was projected (Fu and Wang, 2011). The removal efficiency of HM by MEUF is governed by metal-ion concentration, solution ionic strength and pH, and membrane operational parameters. Still, these membrane separation techniques have some drawbacks regarding stability under acidic and alkali conditions, fouling produced by organic and inorganic substances, and decreased durability.

3.7.4. Adsorption

Conventional technologies generate a large amount of sludge that is difficult to dispose of and act as a secondary pollutant. Therefore, it is important to grow an economical technique for the removal of Cd2+ from wastewater. Adsorption has gained much attention due to its cost-effectiveness, and it is available both locally and naturally.

3.7.4.1. Activated carbon

Activated carbon (AC) has been extensively utilized for the removal of various pollutants by adsorption. The adsorption capacity of AC is influenced by adsorbent nature, adsorbate nature and concentration, and solution pH. Ku and Peters (1987) reported 90% Cd removal from industrial wastewater containing cyanide ion (a chelating agent) by Darco S51 AC. Babic et al. (2002) observed that at pH < 3.5, adsorption was negligible by AC cloth in aqueous solution; however, adsorption increased suddenly between pH 4–9 with a maximum at > 9.5 pH.

3.7.4.2. Nano adsorbents

Nano adsorbents consist of two categories: carbon nanotubes (CNT) and nano-sized metal oxides (NMO). The benefits of using CNT are its unique structure, semiconductor, electronic, mechanical, optoelectronic, physical and chemical properties. CNTs are of two types: single and multi-walled CNTs. NMO comprises manganese oxide, cerium oxides, ferric oxide, and titanium oxide. The disadvantage of using NMO is that the material gets agglomerated because of the van der Waals forces, which leads to a decrease in the adsorption capacity.

Benjamin and Leckie (1982) reported the effect of SO4, S2O3, and Cl on the Cd2+ adsorption on oxide surfaces and observed 80% removal efficiency. Liu et al. (2009) synthesized magnetic chitosan nanocomposites for HM removal. A coordinate bond is formed between chitosan present on magnetic nanoparticles surface and HM. This process is reversible, and HM can be recovered by using weak acidic conditions with ultrasound radiation. Vukovic et al. (2010) functionalized MWCNT utilizing O - (7-azabenzotriazol-1-yl), N, N, N, N-tetramethyluranium hexafluorophosphate and found that even at a higher concentration (50 mg/L), it is not cytotoxic. Krawczyk and Jeszka-Skowron (2016) determined Cd2+ in water sample by adsorbing on MWCNTs. The detection limit was observed to be 0.001 μg/L.

Perez-Aguilar et al. (2011) studied batch adsorption of Cd2+ by three different CNTs (oxidized form of nitrogen-doped multiwall carbon nanotubes, multiwall carbon nanotubes, and single-wall carbon nanotubes). They observed that Cd2+ diffusion varied depending upon the CNTs morphology. Behbahani et al. (2013) functionalized MWCNTs with di-phenylcarbazide and studied for Cd2+ adsorption in water. Results showed that MWCNT has an 86 mg/g adsorption capacity for Cd2+.

3.7.4.3. Biosorption and natural adsorbents

The benefits of using biosorbents are high adsorption capacity, no usage of chemicals, less sludge generation, easily available, reusable, regeneration property and cost-effectiveness (Wang et al., 2017a, Wang et al., 2017b). Several studies were carried out for the sustainable removal of Cd2+ by biosorbents (Table 3, Table 4, Table 5). Coal fly ash (industrial by-product) results from coal combustion in thermal power plants (Kumar et al., 2019). Results showed that removal efficiency decreased with the decrease in pH. However, the presence of Ca and Cl showed no effect on Cd2+ removal efficiency. Papandreou et al. (2007) reported Cd2+ removal by coal fly ash pellets (CFP) in an aqueous solution. The adsorption capacity of CFP was observed to be 18.98 mg/g. Joseph et al. (2020) prepared FAU-type zeolites from coal fly ash to remove Cd2+, Co2+, Cu2+, Pb2+, and Zn2+ in an aqueous medium. The adsorption capacity of FAU-type zeolites was observed to be in following order: Pb2+ > Cu2+ > Cd2+ > Zn2+ > Co2+. Ziyadanogullari et al. (2004) investigated Cd2+ removal by soil comprising of magnesite in an aqueous medium. They found that soil was heated to 700 °C showed nine times more Cd2+ adsorption than untreated soil (25 °C).