Heavy Metal Pollution and Solutions for Its Control: General Aspects with a Focus on Cobalt Removal and Recovery from Aqueous Systems

Abstract

Heavy metal pollution is a worldwide and stringent concern following many decades of industrialization and intensive mining without (in some cases) consideration for environmental protection. This review aims to identify the existing and emerging techniques for heavy metals (HM) removal/recycling from water and wastewater, with an emphasis on cobalt. Unlike many other heavy metals, cobalt has not been considered a detrimental element for the environment and human beings until recently. Thus, several methods and applicable techniques were evaluated to identify the best treatment approaches applicable to cobalt-polluted water and wastewater. The most feasible depollution methods adapted to the source, environment, and economic conditions were investigated and concluded. The operations and processes presented in this paper are conventional and innovative as well, including precipitation, membrane separation, with emphasis on ultrafiltration (UF) and nanofiltration (NF), but also reverse osmosis/forward osmosis (RO/FO), sorption/chemisorption processes, flotation/mechanical separation operations combined with coagulation/flocculation, photocatalysis, and electrochemical processes. For each one, depending on the frequency of use, physicochemical mechanisms and optimal operational conditions were identified to carry out successful cobalt removal and recovery from aqueous environments.

1. Introduction

The natural emissions of heavy metals (HMs) occur under numerous but certain environmental conditions. Volcanic eruptions, sea-salt sprays, forest fires, and rock weathering are generating particles containing HMs, among other pollutants [1]. HMs can be found in various forms, from hydroxides, oxides, sulfides, sulfates, phosphates, and silicates, to other organic compounds. However, HMs can result especially from anthropogenic sources/industrial processes and end up in all environmental components (soil, water, air, and their interfaces). HMs can also be found in domestic wastewater (WW), landfill leachates, agricultural runoffs, and also in acid rain [2].

Nearly 200 years of industrial revolution have required, and still require, the processing of more and more metals or metal derivatives. This led to significant accumulations of HM compounds in the environment. All these compounds are characterized by toxicity. First, this HM pollution is caused by industrial wastes resulting from the processing of concentrated or diluted ores. However, most of this pollution/toxicity results from chemical manufacturing processes, known as primary sources of hazardous metals, especially finishing metals, mining operations, smelting, battery manufacturing, electroplating, etc. [3,4,5,6]. In a comprehensive work, Ali et al., 2021 summarized natural and anthropogenic processes generating HMs, as depicted in Figure 1 [1].

Figure 1. Sources of heavy metals [1].

Figure 1. Sources of heavy metals [1].

The enhanced use of metals and their derived chemicals in a continuously higher number of manufacturing processes has resulted in significant amounts of effluents containing high levels of toxic HM, which cause environmental issues, especially due to their persistence and degradation inability [7]. Because of HMs’ non-degradability, the resulting pollution can be defined as irreversible, and consequently, it must be carefully monitored and managed. Currently, the presence of HMs is strictly monitored in soil, outdoor air, surface and groundwater, or drinking water in accordance with regulatory laws, but human exposure/biological monitoring is rarely achieved [8,9]. While the data show a continuous decline in HM emissions, improvements in technologies and targeted legislation are needed to reduce the health impact of the pollution [10].

Regarding anthropogenic pollution, it is worth mentioning that a careful analysis of all industrial facilities, operating or abandoned (such as closed mines), will confirm the generation of HM pollution. For instance, lead emissions were still reported in the US many years after leaded gasoline was banned, indicating remaining waste streams [11]. A few other examples can be added: smelting activities release arsenic, copper, and zinc [12,13,14]; agrochemical use is often associated with HM contamination of soil groundwater and accumulation in vegetables [15,16,17,18]; usage and burning of fossil fuels are releasing nickel, vanadium, mercury, selenium, and tin [19,20]. These are just some of the essential industrial processes generating the contaminants that contribute significantly to HM pollution in the environment.

In a generalized evaluation, it can be said that increasing necessities and requirements for a better life on a larger population scale determined a similar ascending trend to environmental pollution. In this global context of improving environments and health, another example is represented by the shift to a low-carbon economy that triggered the necessity of energy storage capacities. Since 1991, when rechargeable lithium-ion batteries (LIBs) were commercially launched, progress in the field and their impact on society cannot be questioned [21]. In addition to the obvious advantages of LIB usage in everyday life, including the improved capability to power electric transportation, this industry generated the requirement for other metals that were not used as often. This is the case with cobalt (Co), for which the commercial demand increased exponentially. In fact, approximately 60% of global cobalt is used for battery production, while the prognosis for the future indicates a high supply risk and only sustainable sourcing could solve this increased consumption need [22]. Although solutions are being researched for novel cobalt-free cathodes, the hope is for a gradual reduction in the cobalt ratios in the near future [21].

Cobalt, a transition element, discovered by the Swedish chemist Georg Brandt in 1735, is widely spread in all environmental components, and it is one of many essential micronutrients for prokaryotes, animals, and a beneficial (or even essential) micronutrient for plants. For human beings, Co is a cofactor in Vitamin B12 (known as cyanocobalamin), which is essential for brain functioning, nervous system activity, and hemopoiesis [23,24]. At higher concentrations, cobalt toxicity—be it natural or synthetically produced—results in detrimental effects in all biological systems, as new studies report [25,26,27]. There are numerous disorders associated with increased Co exposure, which are related to various tissues and organs (thyroid gland, lungs, skin, immune system), and its carcinogenic potential is also demonstrated. Moreover, Co’s toxic potential, toxicity mechanism, and clinical response depend on exposure settings, intake routes, and Co source [23]. For instance, Co ions and nanoparticles exhibit cytotoxic potential, while metal and salts can induce genotoxic effects [28,29,30].

Since Co was previously considered to not have a significant environmental impact among HMs, no specific limits were imposed for many years. However, with the constant increase in its concentration, national restrictions worldwide have been set. There are specific limits for Co implemented in different countries; for example, an irrigation water limit of 50 µg/L (U.S. Environmental Protection Agency, US EPA), freshwater 1 µg/L Federal Environmental Quality Guidelines, Canada), and a groundwater limit of 4 µg/L (Wisconsin Department of Health Services) [31,32]. Regarding the drinking water quality, current guidelines do not provide limiting values for Co in general. There are some exceptions. The US EPA’s limiting value for Co is 100 µg/L (0.1 mg/dm³) [33]. In Romanian Regulation NTPA 001/2002, regarding the establishment of pollutant loading limits of industrial and urban wastewater when discharged into natural receivers, a 1 mg/dm³ limit value was indicated for Co [34]. Since numerous studies are now assessing Co toxicity, acute and chronic, to humans and freshwater organisms, it is expected that WHO and EU standards will provide Co guidelines values in the near future [23,35].

In this context of a constant increase of Co industrial demand and acknowledging widespread exposure to Co and Co compounds, the major objectives of this review are:

(i)

Describe and evaluate the present technologies for HM removal (Co in particular) and recovery from contaminated water and WW.

(ii)

Identify research trends and new treatment perspectives.

Although there are numerous published review articles related to Co, only a few refer to Co removal from water/WW specifically. While most of them cover a certain method or process (like adsorption, for instance, as being the most applied in these systems [36,37]), this review is trying to cover most of the applied/applicable treatment techniques, from classical precipitation to advanced photocatalytic methods. Treatment techniques are described in detail, and their implementation in Co removal is emphasized.

The literature survey was conducted using Advances Search options of the major research platforms/bibliographic databases (ScienceDirect Freedom Collection, Elsevier, Royal Society of Chemistry Journal Collection, SpringerLink Journals, Wiley Journals, MDPI—Publisher of Open Access Journals). The terms used for the study were [“heavy metals” OR “cobalt”] AND [“wastewater” OR “water”] AND [“name of method/technique”]. The search was not limited to certain years, excepting “adsorption”. Due to numerous publications related to adsorption/chemisorption processes for cobalt (HM) removal, a detailed review article was developed and published in 2018 by Islam et al. [36]. Thus, our survey was limited from 2019 to 2024 for this particular separation technique.

Articles of any type were considered, including reviews, original research articles, and communications, to allow a deep understanding of different authors’ perspectives. No specific software or AI-powered tool was used in the literature survey or manuscript development.

2. Methods in Water and WW Treatment

As emphasized before, pollution with HMs exists in different environmental compartments. Air particulate matter (PM) containing HM, resulting from industrial activities, and volatile compounds (VC), resulting from burning fossil fuels, can be found. In the soil, HMs accumulate by their transfer from water (leaching) and from air (re/sedimentation and adsorption). Thus, methods of controlling the effects of HMs should refer to each of these compartments. Nevertheless, drastic and more restrictive environmental rules on HM discharge limits are needed, consequently leading to the development and implementation of new, more efficient treatment technologies.

This review will focus on traditional, existing, and emerging technologies and methods in water and WW treatment. The water bodies are the main transport vectors for HM pollution in living systems. To control bioaccumulation, the authorities have continued to lower the permissible limits over the years for both drinking water and water for agricultural use [11,38,39,40].

It is known that HM pollutants occur in two forms: particulate (either elemental or as insoluble compounds) or dissolved. The dissolved HMs are most problematic as they cannot simply be filtered. As will be further discussed, some HM removal technologies convert the dissolved metal ions into insoluble compounds, such as salts, and then separate them from the solution by filtration. Others, such as ion exchange, will remove dissolved ions directly from the solutions, but with some limitations. So, dissolved HM species, as typical for most industrial effluents, can be removed via several methods, among which the most common are metal hydroxide precipitation, metal sulfide precipitation, ion exchange, evaporation, and membrane separation. Because of their high price, evaporation and membrane separations are only used in special cases [41,42,43,44].

HM species, in general, and Co, in particular, can be successfully removed and recovered using a few methods, including precipitation, adsorption/adhesion, filtration/elution, electrochemical treatment, adsorption, and membrane separations. However, Co (as any other HM) is rarely found to be the single WW contaminant. Just a few examples can be provided as follows: the electroplating industry generates Co²⁺-Cu²⁺ contamination, the processing of nickel ores will generate Ni²⁺-Co²⁺ WW, the post-processing of metal materials generates complex WW with several HMs in composition, and so on. Therefore, a successful effluent treatment will often require integrated solutions [45].

The following sections will refer to the generalities of each applicable technique/method and particularities of water or WW containing Co as single or co-contaminant treatment. In most cases, because of the limited number of publications found, the evaluation is narrative. A systematization was performed only in the case of adsorption, where the number of publications is higher.

2.1. Precipitation

In most cases, the HM species that pollute the water can be precipitated from solutions, depending on the case, such as hydroxides, sulfides, or carbonates. Carbonate precipitation is rarely, if ever, used exclusively as a precipitation method. However, it can be used in combination with other methods, especially with hydroxide precipitation.

HM precipitation can be accomplished with common inorganic precipitants like lime, caustic soda, soda ash, sodium bicarbonate, sodium sulfide, sodium hydrosulfide, ferrous sulfate, ferric chloride, alum, or filter alum. Any of these is easy to find and accessible, so the precipitation method should be effective both economically and technically [46,47]. The selection of the best precipitant and dosage is not always easy, and most of the time can be concluded only after experimental investigations (jar tests). Since industrial WWs are often complex mixtures of contaminants, it is important that experiments are conducted in similar environments as much as possible (filed evaluation). Two major factors determine the chemical selection: the final HM concentrations in treated water (to meet standards or recycling requirements) and the sludge purpose (treated for reuse or disposal) [46,48].

Various factors influence precipitation and sludge characteristics: pH, reaction time, aging time, mixing intensity, and temperature. Nevertheless, the most important parameter is the pH of metal removal and precipitate characteristics, as highlighted by Zhang et al., 2024. In their recently reported study, sodium carbonate was used for Co recovery from sulfate solutions with the aim of obtaining a Co carbonate as a valuable material. Depending on pH, Co precipitated as carbonate hydroxide between 9 and 11, cobalt carbonate around 7, and mixtures between 7 and 9 [49]. Thus, the pH in the operation should always be adjusted and correlated with expected treatment outcomes.

In the simplest applications, metals can be precipitated from solution as hydroxides as exemplified by chemical reactions (1) and (2):

$${\mathrm{M}}^{2+}+{2 \mathrm{O}\mathrm{H}}^{-}\to {\mathrm{M}\left(\mathrm{O}\mathrm{H}\right)}_2$$

(1)

$${\mathrm{M}}^{3+}+{3 \mathrm{O}\mathrm{H}}^{-}\to {\mathrm{M}\left(\mathrm{O}\mathrm{H}\right)}_3$$

(2)

The resulting HM hydroxides are denser than water and can be separated by mechanical filtration or, more conventionally, settled by gravity in a clarification system. To design a separation process based on HM precipitation as hydroxides, the correlation between the hydroxide solubility and solution pH must be considered. Figure 2 shows metal hydroxide solubilities, indicating the possibility of achieving compliance with permissible limits for discharge despite the weak bonding energies [50].

Figure 2. Hydroxides solubility vs. pH for some heavy metals [50].

Figure 2. Hydroxides solubility vs. pH for some heavy metals [50].

As can be seen in Figure 2, most metals exhibit amphoteric properties, meaning there is a particular pH value/or a narrow pH interval at which they are at a minimum solubility. Any change in pH value (increase or decrease) transfers the HMs back into the solution as free cations. Knowing that the typical limits for most HM concentrations in treated polluted water are in the range of 0.50 to 1.00 mg/L (0.5–1 ppm), it is obvious that most pollutants of this sort can be removed by hydroxide precipitation. Disregarding the precipitant and precipitation conditions, it must be mentioned that in real WW (in practice), theoretical solubilities and pH ranges specific to single metal cation solutions are to be considered just as indicative. Each system, depending on the complexity of its composition, will behave differently.

Figure 3’s case A shows schematically a hydroxide precipitation treatment plant for WW with HM content. In this case, the resulting sludge consists of only metal hydroxides as a consequence of pH adjustment and control. The pH changes can be accomplished by adding sodium hydroxide or hydrogen chloride aqueous solutions to the reactors, as required by each system’s specifics.

Most metals can also be precipitated from the solution as sulfides. Usually, sulfide precipitation will allow for a much lower residual concentration of ion metals in solution under much broader conditions in general. Chemical sulfide precipitation occurs by the following chemical reaction (3):

$${\mathrm{M}}^{2+}+{\mathrm{S}}^{2-}\to \mathrm{M}\mathrm{S}$$

(3)

The sulfides’ working procedure is a modification of hydroxides precipitation treatment technology for HM removal from WW, illustrated in Figure 3, case B, where sodium sulfide solutions are used as precipitating agents.

In these systems, controlling the S²⁻ concentration in wastewater and even in water after treatment is mandatory because, during the neutralization process (alkalinity treatment), the release of H₂S may occur. In any case, the possibility of H₂S discharge must be avoided because of its toxicity hazards, corrosion hazards, and potential explosion hazards (H₂S generates explosive mixtures with air).

Sulfide precipitation systems can be designed for various flow rates and various HM concentrations aiming at either the removal or the recovery of metals and metalloids. As is always recommended, a co-precipitating/coagulating agent is used together with sulfide, disregarding the concentrations and conditions. An anionic polymer is usually added to aid the flocculation process. This results in additions to both the cost of operation and the sludge volume, but it does allow for the precipitation of the target species, which would be quite difficult to achieve otherwise.

Unlike hydroxides, metal sulfides are not amphoteric, so wider pH ranges can provide a high degree of metal removal in shorter amounts of time [47]. The lower solubility of sulfide products allows for much lower metal loading in the effluent stream under a much wider range of characteristics referring to pH range, sequestering the agents’ presence and their concentration, operation temperature, and other parameters with low influence in HM sulfides precipitation. In fact, the most important parameters influencing the sulfide process are pH and the type of sulfur source used for metal stabilization and recovery [51]. Since at a pH lower than 7, there is always the risk of toxic hydrogen sulfide release, the alkalization of WW is the first step in the development of this process. Inorganic coagulants can be added in this stage to form micro-precipitates, around which the sulfides will then accumulate.

Because of the very high nucleation rates, the HM sulfides generally form rigid inorganic colloids (small and compact particles) [52]. However, compared to metal hydroxides, sulfide precipitates exhibit better-settling properties, and the formation of colloids causes problems in the separation process because of their very small dimensions [47].

Many of the above-mentioned aspects seem to lead to the assessment that the hydroxide precipitation of HMs would be superior (or preferred) to precipitation as sulfides. However, the process of depollution by sulfide precipitation, even with the danger of accidental discharge of H₂S, can be intrusive because the solubility of HM sulfides is much lower than that of HM hydroxides. Given Co as an example, its solubility as hydroxide is 2.2 × 10⁻¹, while as a sulfide, the solubility is as low as 1.0 × 10⁻⁸ (values reported for pure water) [47]. Despite being considered classical or traditional, metal sulfide precipitation has gained much interest in recent years, showing promising perspectives due to novel methods to control the supersaturation, design and optimization of new solid–liquid separation systems and new reactor types and new sulfide sources [53].

Generally speaking, both precipitation methods can be applied due to simplicity of operation and good performance. However, there are some major drawbacks to be noted: filtration issues in case of gelatinous precipitates (hydroxides especially), sequestering of metals via chelating agents (citrates, EDTA, amines, etc.), solubility of precipitate under pH changes, and improper pH for discharge of treated WW [54]. In particular, chelating compounds like citrates, EDTA, CDTA, amines, etc., will link the metal ions in a chemical complex, isolating them from free anions in solution, such as hydroxides. Another limitation, specific for sulfide precipitation, results from the remaining S⁻² in the solution. Nevertheless, it is important to mention that sulfide precipitation has a much higher affinity for HM cations than hydroxides, and the precipitation kinetics exceed the common sequestering kinetics. Therefore, precipitating metals such as sulfides, as opposed to hydroxides, will often ameliorate sequestering issues, allowing for the insoluble salts to form when chelating agents are present in processed water.

In addition to the general guidance on HM precipitation, the process is characterized by a high degree of specificity, meaning that customized solutions will be needed (or can be developed) for any studied environment. This specificity results especially from the composition of the aqueous system, especially when certain metal recovery is important, as is the case with Co. Although there are reported studies regarding only Co precipitation, these are conducted on synthetic waters as Co, in real cases, is one of the two or more co-contaminants.

Studies on the co-precipitation of HMs were also conducted. For instance, precipitation in Co-Mo aqueous systems was investigated [55]. The results depict a dynamic system where both ion metals precipitation is influenced by the co-ion presence (equilibrium speciation-controlled system). In addition, due to the Co²⁺ presence, the Mo³⁺ precipitates around the pH values of 4–5, Co²⁺ removal in this system was achieved for higher pH values (around 12), as cobalt hydroxide. Higher precipitation rates were obtained when the reaction temperature was increased to 60 °C. The results also showed that most of the molybdenum (90%) was found in the solution at this pH, allowing for preferential Co²⁺ precipitation. However, higher Mo³⁺ concentrations determined its physical entrainment in the precipitate and a slightly lower pH value for Co²⁺ precipitation.

Important influencing factors, like precipitant type, precipitant-to-metal ratio ([P]/[M]) and pH, were investigated on Co removal from synthetic electroplating wastewater. Both hydroxide and carbonate as precipitating agents were used in single- and co-contaminated systems (Co²⁺-Cu²⁺) [24]. The results showed improved removal efficiencies when carbonate was used (soda ash in particular), but consequently, higher sludge volumes were obtained. Cobalt was almost completely removed at pH around 10, but a higher [P]/[M] ratio was needed in co-contaminated waters.

The earliest Co removal studies were conducted on Ni²⁺-Co²⁺ co-contaminated systems or Zn²⁺-Co²⁺ (hydrometallurgical, electroplating, mining WW) [56,57,58,59]. Each of these studies shows high specificity and the necessity of every factor influencing evaluation. Customization for each system is required, as no general conclusions or solutions can be drawn.

Conventional precipitation methods are applicable, especially when metal ions concentration in WW is elevated. More recently, new precipitation methods, having an eco-friendly approach, have been reported. Some of these methods, employing enzymes and microorganisms, are known as bioprecipitation.

For instance, Chaudhuri et al. (2013) used alkaline phosphatase (AP) to mediate the precipitation of HMs in single systems and industrial effluents (tannery and electroplating WW). In the presence of the AP and using a proper substrate, insoluble metal-phosphates are generated. The first studies employed the use of p-Nitrophenyl phosphate (pNPP) as a substrate and AP from two different sources, Escherichia coli C90, bacterial alkaline phosphatase (BAP) [60], and calf intestinal alkaline phosphatase (CIAP) [61]. More recent studies involved the use of ascorbic acid 2-phosphate as a natural and biodegradable alternative for synthetic substrates [62]. Kinetic studies in all cases were conducted for pH between 8–11, with ion metal concentrations ranging from 250 ppm to 1000 ppm. The temperature was maintained at 37 °C (optimum for the enzyme), and the reaction time was set between 30 and 130 min.

The metal phosphate precipitation mechanism is given in Equations (4) and (5) below:

$$\mathrm{s}\mathrm{u}\mathrm{b}\mathrm{s}\mathrm{t}\mathrm{r}\mathrm{a}\mathrm{t}\mathrm{e}+\mathrm{e}\mathrm{n}\mathrm{z}\mathrm{y}\mathrm{m}\mathrm{e}\to \mathrm{b}\mathrm{y}\mathrm{p}\mathrm{r}\mathrm{o}\mathrm{d}\mathrm{u}\mathrm{c}\mathrm{t}+\mathrm{p}\mathrm{h}\mathrm{o}\mathrm{s}\mathrm{p}\mathrm{h}\mathrm{a}\mathrm{t}\mathrm{e}+\mathrm{e}\mathrm{n}\mathrm{z}\mathrm{y}\mathrm{m}\mathrm{e}$$

(4)

$${\mathrm{m}\mathrm{e}\mathrm{t}\mathrm{a}\mathrm{l}}^{\mathrm{n}+}+\mathrm{p}\mathrm{h}\mathrm{o}\mathrm{s}\mathrm{p}\mathrm{h}\mathrm{a}\mathrm{t}\mathrm{e}\to \mathrm{m}\mathrm{e}\mathrm{t}\mathrm{a}\mathrm{l}̵\mathrm{p}\mathrm{h}\mathrm{o}\mathrm{s}\mathrm{p}\mathrm{h}\mathrm{a}\mathrm{t}\mathrm{e}\downarrow \left(\mathrm{p}\mathrm{p}\right)$$

(5)

For Co²⁺ bioprecipitation, good results were obtained for both enzymes: an average of 71% in the case of BAP and more than 90% with CIAP in single-ion solutions. Almost 98% of the Co²⁺ precipitates with CIAP after 120 min reaction time was at a pH of 10 when the initial ion metal concentration in WW was 250 ppm. The initial metal ion concentration was found to be an important factor in enzyme activity, especially in the case of BAP, which was affected more.

Another interesting alternative to conventional precipitation, in this case, is sulfate reduction, which is represented by biological sulfate reduction (BSR). In this process, the sulfate in WW is reduced to sulfide in the presence of sulfate-reducing bacteria (SRB). The resulting “biogenic sulfide matrix” will facilitate ion metal precipitation as metal sulfides [63]. In a comprehensive study, Liu et al. (2019) revealed the importance of sulfide sources for Ni²⁺ and Co²⁺ precipitation. Chemical sulfide and four different bioreaction media, as sources of biogenic sulfide with specific compositions, were evaluated in batch experiments at a pH of 7.8 and variable initial metal ion concentrations [64]. For Co²⁺, the results showed improved precipitation (>98%) for biogenic sulfide, disregarding its source and initial metal ion concentration. However, the composition of the biogenic matrix seriously affected the settling characteristic of the obtained precipitate, especially in the case of residual phosphate. Another important conclusion of the study relates to the selective recovery of Co²⁺ and Ni²⁺ since the Ni²⁺ precipitation was more sensitive to different environmental conditions.

A more effective alternative to HM precipitation is represented by fluidized-bed crystallization (FBC) [65]. Compared to the conventional precipitation method, FBC ensures reduced sludge formation using millimeter-size seeding particles (silica sand, quartz, etc.). Metals could precipitate in the solution and then adhere to seeds or simply precipitate upon the seed’s surface. The resulting sludge separation is feasible due to the larger obtained particles, low humidity, and relatively high ratio between precipitate and seed per se [66,67,68]. The large surface area in FBC allows high reaction rates, so the method is suitable for treating high quantities/flows of WW in continuous operation [67].

Figure 4 provides a comprehensive representation of a fluidized bed reactor (FBR) where crystallization occurs and of the main characterizing indicators for the process [66].

Figure 4. (a) FBR schematics, (b) process performance indicators: χ, conversion (proportion of inlet metal that precipitates) and η, recovery (proportion of precipitate retained onto seeds); and (c) image of precipitate formed around silica sand seeds [66].

Figure 4. (a) FBR schematics, (b) process performance indicators: χ, conversion (proportion of inlet metal that precipitates) and η, recovery (proportion of precipitate retained onto seeds); and (c) image of precipitate formed around silica sand seeds [66].

Lewis and van Hille, 2006, reported metal sulfide precipitation in such a system, including Co²⁺ [65]. They used a glass column 1.5 m in height with a 25 mm diameter filled with sand. Aqueous media and a solution for metal ions were fed at the bottom of the column, while the sulfide inflow was divided into three to reduce supersaturation (Figure 5a). The study emphasized the importance of suprasaturation for low and high concentration levels of ion metals and the sulfide’s excess influence. Thus, supersaturation increases the quantity of fine particles; furthermore, at a high pollutant level, the attached precipitate is unstable because of the rapid growth generating fine particles, which also detriments removal efficiency. An excess of the sulfide also has a negative effect, enabling the formation of polysulphide species.

There are also drawbacks to this method: the end-product could retain impurities, so metals recyclability is questionable; supersaturation may occur, disregarding operation adjustments with a negative impact on process performance. The recyclability is, however, limited due to product structure itself, even without impurities (precipitate-coated seeds). To overcome such disadvantages, a novel crystallization method was proposed by Chen et al. (2015) [67]. This refers to homogeneous crystallization/granulation in an FBR (FBHC/FBHG) where seeding is not required, as schematically represented in Figure 5b. In such systems, mild degrees of supersaturation are maintained. High-purity homogenous crystals can be obtained as the fine precipitate particles of metal salts become the nucleation active sites. Careful consideration of operational parameters (hydraulic) will determine the end-product (precipitate) characteristics.

Initially designed for WW containing lead (Pb²⁺) remediation, the FBHC was later tested for other metals recovery, including Co²⁺. Bayon et al. (2021) conducted the remediation of semiconductor WW aimed at zero-liquid discharge policies and the recovery of valuable resources (Co²⁺, Cu²⁺, and phosphate). Co²⁺ removal was studied in a single system and in the presence of Cu²⁺. The experiments were performed for WW, which had a 7 mM metal ion concentration in environments with variable pH levels (between 4.5 and 12) at variable upflow velocities and variable phosphate concentrations. The influence of chelating agents like citrates and EDTA was also analyzed. The successful removal of Co²⁺ (~99.0%) was achieved for cobalt phosphate at a pH of approximately 8 and a $\left[{\mathrm{P}\mathrm{O}}_4^{3-}\right]/\left[{\mathrm{C}\mathrm{o}}^{2+}\right]$ molar ratio of 2.0. Proper hydrodynamic conditions allowed for uniform crystal growth in cleaner WW. Strong chelation agents were found to deeply inhibit the crystallization, while the influence of weaker chelating agents was milder. Good granulation results were also obtained in the dual systems (96.8% for cobalt at pH 7.5 and 95.9% for copper at pH 6.5) [69].

In a similar approach, Anotai et al. (2021) studied the individual and co-precipitation of Co²⁺ and Cu²⁺ from synthetic WW with sodium carbonate as a precipitation agent. Cobalt carbonate and copper carbonate hydroxide were obtained with good removal and granulation efficiencies, and the best operating conditions were identified in single and mixed systems. A chloride presence (normally occurring in WW) was, in all cases, found to be inhibitory for the process as metal chloride complexes are forming [70]. The similar influence of chloride ions and a chelation agent on Co²⁺ and Cu²⁺ recovery in FBHG was also recently reported [68].

Figure 5. (a) Seeded FBR (adapted after Lewis and van Hille, 2006 [65]) and (b) FBHC system (adapted after Chen et al., 2015 [67]).

Figure 5. (a) Seeded FBR (adapted after Lewis and van Hille, 2006 [65]) and (b) FBHC system (adapted after Chen et al., 2015 [67]).

All reported studies show the importance and accessibility of precipitation as an HM removal technique in industrial development. Numerous precipitants are available, and their use in various configurations was investigated in synthetic WW and real systems. The technical and economic feasibility of such systems is well recognized. In any case, for each specific contaminated water/WW source, depending on the composition and characteristics, the best solution in terms of operation and design should be searched, from jar tests to field evaluation, to ensure the successful achievement of targeted treatment objectives.

2.2. Electrochemical Methods (EM)

Electrochemical methods for water and WW treatment offer the advantages of chemical selectivity, reduced production of secondary wastes, compactness, and, most importantly, efficient removal of trace pollutants. In such systems, applied electrical currents promote contaminant removal by two different mechanisms, including one governed by electrokinetics—separations in bulk electrolytes—or electrosorption by intercalating in solid electrodes or becoming trapped in electric double layers [71].

Among EM, electrocoagulation has been considered an attractive technique, as it is simple and effective, and no other chemicals are required. In this technique, sacrificial anodes are on magnesium, zinc, or aluminum) to are used to generate in situ coagulates. The method can allow for HM removal in multiple ways: electrodeposition due to reduction on the cathode, forming precipitates with the OH− generated on the cathode, and adsorption by the in situ generated M(OH)_n with such capabilities [72].

Early studies indicate that a simple solution such as alternating current electrocoagulation may be suitable for Co removal from drinking water. Careful consideration of the operational parameters of more than 98% of Co(II) can be removed at pH 7.5 in approximately 35 min. Experiments were performed in batches using a 2 L reactor equipped with two monopolar aluminum plate anodes and cathodes, and the initial Co ion concentration was considered to be between 5–25 mg/L. The removal efficiency decreased when the initial Co ions concentration increased, but overall, the results were satisfactory, as the generated gelatinous charged aluminum hydroxides proved to have good adsorption ability [73].

Electrocoagulation with iron electrodes was applied to contaminated industrial WW containing Co and Cu. Favorable working conditions allowed metal ion removal and recovery as highly pure salts, especially in concentrated streams [74].

Electrolytic Co ion removal was achieved in highly concentrated synthetic WW (100–400 mg/L Co(II)) with good results using an easily available galvanized carbon steel cathode screen. Thus, Santos et al., 2016, reported more than 73% removal after 185 min using a current of 0.3 A at 30 V [75].

Electrodeposition of Co ions was successfully achieved at the laboratory scale, using a stainless-steel rod as the cathode and a graphite plate as an anode. More than 99% removal was achieved in 20 min for synthetic WW containing 10–50 mg/L cobalt ions at pH 6 [76]. The method allows for an easy and fast solution for HM recovery.

Electrooxidation on synthetic and river WW was performed using magnetite nanoparticles to remove Co and Cd, revealing almost 100% removal possibility, depending on operational parameters (applied voltage, contact time, HM initial concentration, pH, and selected organic additives) [77].

Among the existing EM, one of the most applied in water and WW treatment is electrodialysis. Semi-permeable ion-selective membranes are used in electrodialysis to extract ionic components from a solution. It is possible to conduct this procedure in batches or continuously. Concentration polarization is one of the most important technical issues related to electrodialysis. Membrane scaling because of inorganic feed solutions and membrane fouling because of organic feed solutions are further issues in specific applications [78]. A major advantage of this method, besides being economical, is the production of reusable effluent water, while HM ions can be removed even at the lowest concentrations [41]. Figure 6 shows the electrodialysis particularization for the removal of HM-positive ions [79].

Figure 6. Principle of electrodialysis separation of heavy metal-positive ions (CM—Cationic membrane, AM—Anionic membrane, D—dilute chamber, K—concentrate chamber, e1 and e2—electrode chambers) [79].

Figure 6. Principle of electrodialysis separation of heavy metal-positive ions (CM—Cationic membrane, AM—Anionic membrane, D—dilute chamber, K—concentrate chamber, e1 and e2—electrode chambers) [79].

Thus, it can be concluded that EM is available and tested. Clean and simple, these methods are good options for large-scale applications. The only issues remaining are the preparation of proper electrodes at accessible costs and the relatively high energy consumption required in the treatment process [80,81,82].

2.3. Membrane Separation

2.3.1. Ultrafiltration (UF) and Nanofiltration (NF)

Nowadays, it is more and more common to include ultra-filtration (UF) and nano-filtration (NF) technologies in a wider range of WW treatment and water reuse systems. These procedures use permeable membranes to separate HM, macromolecules, and suspended solids from solutions on the basis of the pore size and molecular weight of the separated compounds [83]. Both reduce the number of chemicals needed to treat a polluted flow and the necessary time to achieve the treatment objective. However, pre-treatment is required in almost all membrane applications, and membranes need to be replaced in time. However, depending on membrane characteristics, UF/NF can achieve more than 90/98% of removal efficiency when metal concentration ranges from 10 to 110 mg/L, at pH ranging from 5 to 9.5, applying a convenient pressure around 2–5 bar [84]. If the water or WW presents a more complex composition, then a higher pressure gauge is required and can be applied (up to 10 bar in the case of UF and 30 bar to NF).

The membranes can be classified into two major categories, depending on the material they are made of: polymeric and inorganic (ceramic/metal/or carbon made). Most of the membranes used in pressure-driven separation processes like UF and NF are made of synthetic polymers. Nowadays, given the advancement in the development and fabrication of polymeric materials, membranes can be customized at reasonable production costs. Thus, high-quality membranes with good mechanical properties and excellent separation performances are available for specific applications like HM removal from aqueous solutions.

UF membranes, with nominal pore sizes between 0.1 μm and ~50 Å, allow for the adequate removal of particles with sizes ranging from 0.01 to 0.1 µm as suspended HM particles and polymer-bound HMs [85]. UF presents some advantages, such as lower driving force and smaller space requirements due to its highly compact structure. However, UF membranes with pore sizes larger than 0.01 µm are normally considered inappropriate to retain the metal ions. To overcome this impediment, low transmembrane pressure, the use of complexing or micelle agents to increase metallic ions molecular weight or embedding adsorbents into the membrane matrix are just a few methods that can be implemented [83,86].

NF membranes with pore sizes lower than ~1 Å (0.1 nm) allow for the rejection of finer particles as HM flocs/colloids and divalent ions. Two comprehensive reviews have been recently published concerning membrane filtration methods for HM removal [87,88]. These reviews cover valuable information regarding the successful application of membrane separation techniques in HV removal; however, most of the cited studies refer to other HMs and not Co. The case studies presented below concern Co removal in such systems.

A good example is that HMs in general, and Co in particular, can be successfully removed from aqueous solutions by complexation–UF separation method is the work of Kim et al., published in 2005. The researchers used humic acid, which is an organic matter found in soil that possesses carboxylic or phenyl groups, as a natural complexation agent that is environmentally friendly and economical. The influence of Co²⁺ initial concentration, pH, and the presence of electrolytes were investigated. A membrane of regenerated cellulose was used, and the results showed that the method could offer 95% Co²⁺ removal when HM ions concentration is higher (in this case, 3 g/L) and when the solution pH is elevated (around 8). No significant effect was induced by the presence of counter ions, especially when the humic acid concentration was increased [89].

On the other hand, as previously mentioned, NF can selectively retain divalent ions, including Co^2+, through a combination of electrical interaction (charged membrane surface—metal ions) and size exclusion mechanisms [90]. Adsorption and concentration polarization also help with the purpose.

For the removal of Co²⁺, conventional and more sophisticated membranes were tested. Disregarding the membrane used or the WW characteristics, the plant setup is simple, consisting of the filtration module, feed pump, feed tank, and instrumentation (pressure gauges, pressure valves, and flow meters), as schematically presented in Figure 7.

Figure 7. Typical NF system [91].

Figure 7. Typical NF system [91].

Using the experimental setup presented above, a conventional NF polyamide membrane (AFC30) was tested for the removal of cobalt and lead from synthetic WW. The influence of membrane properties (pore radius and thickness-to-porosity ratio) and the influence of the operating condition (solution pH, permeate flux, salt concentration, and co-ions presence) were evaluated. The results showed membrane applicability in such systems with a high rejection of Co even in mixed-salt solutions and disregarding the pH values [91].

An AFC40 tubular polyamide NF membrane was also tested for Co-ion retention from aqueous solutions. Similarly, the membrane characteristics and operational parameters were analyzed, and the best conditions were concluded: more than 97% rejection rate for a solution having around 100 mg/L Co ions was obtained (cobalt nitrate solution was used for the experiments) at pH 3 for the feed solution and 20–25 bar pressure range. A feasible performance was demonstrated even for the highly concentrated Co-containing WW since the membrane can be operated at high flux and low pressure. A very important aspect of this wide study is demonstrating that a simple model, such as the Spiegler–Kedem model, can be easily applied to accurately describe the practical NF applications in this area [92].

In a different approach, considering that other parameters could also affect NF performance, Nguyen et al. (2011) investigated the treatment of WW polluted with three divalent ions (cations) characterized by similar atomic weights (cobalt, nickel, and copper), using two commercial NF membranes [93]. This important investigation revealed the influence of the solvation of solute, concentration polarization, and complex formation; thus, the rejection order was found to be Ni²⁺ > Cu²⁺ > Co²⁺.

Another study proved the suitability of NF in the industry, achieving metal recovery and process liquid recycling in the same time. It involves the copper–cobalt extraction process where NF can allow a high percentage of metals recovery as concentrated streams from pregnant leach solution (60% recovery), heap leach wash water (80% recovery) and pond WW (more than 85% recovery). The other obtained streams, having acceptable HM contents, can be recycled in the main process (concentrated acid, diluted acid, and water) [94].

Practical experiences in different fields where membranes were implemented showed that successful operation and separation depend strongly on membrane characteristics. Water systems are no exception. So, to improve the separation performance of heavy metals, newer membranes were developed using different modification techniques. For instance, nanoparticles of graphene oxide (GO), functionalized using (3-aminopropyl) triethoxysilane, were used to modify the thin polyamide layer of an NF membrane by surface deposition or embedding [95]. The study demonstrated increased water flux in the case of both modification approaches compared to the control membrane. However, the membrane obtained by GO coating had a higher negative surface charge and ensured better cobalt removal (more than 97%). The experiments were carried out on a laboratory scale with a membrane of 23 cm² and an effective surface area on synthetic water containing 100 ppm cobalt ions. The pressure feed was set to 8 bar, and the flow rate was 0.5 L/min.

With promising results at the laboratory and bench scale, it is clear that UF/NF will soon successfully be implemented at an industrial scale for HM removal as well. Some directions can be considered as priorities for improving performances and feasibility, including membrane surface functionalization, the development of emerging membrane materials, and transitioning to green synthesis routes for membrane fabrication.

2.3.2. Reverse Osmosis (RO) and Forward Osmosis (FO)

Among water and raw water treatment processes for ion removal, reverse osmosis is one of the most common. RO may be used to remove atom-sized species because of its effective size range (down to ~1 Å, similar to NF). Indeed, the RO can recover about 99% of multivalent and 95% of monovalent ions by their concentration as RO membrane reject streams. So, RO makes an excellent method for removing ionic and some nonionic impurities from water.

Using this process to concentrate HMs for recycling or in conjunction with other procedures such as electrolytic recovery or evaporation will result in a very competitive solution, even post-treatment in challenging systems such as mining WW. The main drawbacks are membrane fouling, degradation, and physical damage, which are directly correlated with an efficient pre-treatment system [96]. There are limited RO separation processes used to extract HM ions, and very few have been applied to water-containing cobalt ions [84]. This may be because other methods are easier and more convenient/feasible for such systems as RO demands high energy consumption (operation pressure of 10–100 bar). Furthermore, because of their characteristics, membranes have low water permeability [87].

Interestingly, unlike RO, Forward Osmosis (FO) has gained more attention in the last years, as it is now considered an emerging technology [97]. FO requires lower pressure (the transport across the semi-permeable membrane is actually based on and helped by the natural osmotic pressure), allows for less membrane fouling and easy membrane cleaning, and thus lowers costs for the same contaminants’ rejection performance [98]. FO uses a draw solution with higher osmotic pressure than the WW feed. The water passes through a semi-permeable membrane from the feed solution to the draw solution, and thus, the contaminants are obtained as concentrated streams, as schematically shown in Figure 8 [84].

Figure 8. Schematic representation of the forward osmosis process [84].

Figure 8. Schematic representation of the forward osmosis process [84].

Three commercial FO membranes containing 0.5–20 mg/L cobalt ions and some co-existing ions were tested on WW using NaCl solutions as draw solution. The investigation revealed the importance of a membrane active layer charge in cobalt retention and also the importance of support layer texture for solvent permeability, as expected [98]. Another interesting study focused on operational parameters’ effect on process performance in a batch and flowing operation. In a batch operation, FO proved to be less efficient, indicating that continuous processes may allow for improved water flux and improved rejection at the same time [97].

2.4. Coagulation-Flocculation

Coagulation represents the destabilization of the charge on colloids, followed by their agglomeration, while flocculation is the process of facilitating the physical bonding of small particles, creating larger aggregates [99]. When colloid densities are too low, coagulation and flocculation will increase the density and allow for (enhanced) particle sedimentation [100].

It is possible to use the coagulation–flocculation procedure to remove HMs from wastewater if HMs are found as colloidal species. The resulting bulky particles can further participate in sedimentation. Successful separation depends on coagulant type and dosage, pH, temperature, alkalinity, and mixing conditions [100,101]. An optimal approach can be assessed by practical experimentation (jar tests). This procedure is often used to improve primary settling and/or simultaneous precipitation of HMs and phosphorous or as a tertiary treatment step. It is considered cost-effective (it involves high chemical consumption, but not expensive ones) in offering enhanced HM removal efficiencies and a simple operation [102]. However, the coagulation–flocculation process has major drawbacks, such as the generation of large sludge volumes—it is not selective and cannot completely remove HMs [99]. Currently, there is a continuous search for nanoparticle-impregnated polymers with a high affinity for HM to achieve improved removal, but generally, other treatment techniques are subsequently needed to comply with admissible discharge limits for WW [101].

Because electro-coagulation can remove even the smallest colloidal particles, using an electric current may be a better option than the traditional method. As it is simply applied, fast, and eco-friendly, electro-coagulation, based on the continuous in situ production of a coagulant, has gained much attention, producing purified drinking water and low sludge production [103].

2.5. Flotation

Flotation is an interesting technique characterized by facile operation, effectiveness, and good separation yields, with a recovery of >95% for small concentrations of impurities [104,105]. Two strategies can be implemented to remove impurities/particles as foam at the surface of the liquid, depending on specific requirements, including Induced Gas Flotation (a fast process characterized by turbulent operating conditions, low retention time, and small bubbles) or Dissolved Gas Flotation (microbubbles, quiescent regimes, longer retention time, and high required footprint) [105]. While DGF is widely used for WW treatment in many fields, IGF will be implemented more often in the future because of its increased contaminant loadings.

Usually, the flotation process is assisted by flocculation; however, to improve flotation performance in terms of removal efficiency and selectivity, especially in the case of large quantities of low-concentration metal ions WW, other new, integrated approaches have been studied in recent years, such as ion flotation, precipitation flotation, or sorptive flotation [106,107].

The precipitation-flotation method allows for the separation and recovery of HMs from wastewater streams using a non-surface-active agent to precipitate the target ions. The precipitate, depending on its hydrophobicity, is then removed by gas bubbles with or without the help of a surface-active agent [108,109].

An early study on Co²⁺ ion recovery by DGF (using air as a flotation gas) was conducted in acidic conditions to maintain ion-flotation conditions. Three different collectors were tested, and the results showed that only sodium dodecyl sulfate (SDS) promoted good recovery (around 90%) of Co²⁺. However, most of the Co²⁺ was just hydraulically retained in a voluminous froth, so overall, the method efficiency was low [110]. This study was continued with a different approach: the addition of a sulfide precipitation stage prior to flotation. Na₂S was used to precipitate Co²⁺ under stirring at an ambient temperature and variable pH values. As sulfate precipitation was found to be kinetically limited, 100% recovery was obtained only in the alkaline environment (at pH 10) when the Co precipitates as hydroxide. Flotation was further applied to achieve precipitate recovery. Disregarding the experimental conditions and the usage of different collectors, the overall results were not satisfactory [111].

More recently, laboratory experiments were reported for the simultaneous removal of Cd²⁺, Ni²⁺, or Co²⁺ from distilled water and tap water using potassium ethyl xanthate. The results showed the competitive removal of Cd²⁺ and the other ions (Ni²⁺ or Co²⁺) in the presence of a suitable collector. Removal percentages of over 96% were attained for all ion metals, for lower and for higher Ni²⁺ or Co²⁺ concentrations, indicating that this method could be effective in the simultaneous removal of HM, including Co^2+, from water and wastewater streams [109].

Selective co-recovery of Cu²⁺ and Co²⁺ in WW containing also Ni²⁺ was recently achieved using α-nitroso-β-naphthol as a selective chelating agent [106]. The hydrophobicity of formed insoluble precipitates was improved with the addition of a cationic surfactant (cetyltrimethylammonium bromide). When the most favorable operation conditions were selected (precipitation time 1 h, pH 2, temperature 65 °C, flotation 10 min, proper chemicals addition), more than 98% of Cu²⁺ was recovered and approximately 42% of Co²⁺. A remarkable separation factor of 598 was reported for zinc:cobalt.

Another interesting perspective in the pursuit of new, environmentally friendly solutions is the use of biosurfactants in the ion-flotation process. As an emerging alternative to chemical surfactants, biosurfactants offer biodegradability, biocompatibility, and sustainability, as they can be produced by various microorganisms using renewable resources [107]. Surfactin, isolated from Bacillus subtilis, is a cyclic lipopeptide with an amphiphilic character, and it has been found to be even more effective than SDS as a foaming agent. Due to its complex structure, surfactin has the ability to change its self-assembling properties in the presence of divalent counterions, such as some HM. Furthermore, its chemical structure offers potential binding sites for divalent cations in its structure, and the resulting metal-surfactin complexes were found to be stable [112,113]. These characteristics were exploited by Schlebusch et al. (2023) for the flotation of copper, nickel, and cadmium in single systems. The obtained results were very good in all cases. Specifically for Co²⁺, 98% was recovered after 50 min of flotation at pH 7 and 1:3 initial ion-to-surfactin ratio. The ion metals concentration in the solution was reduced from 100 µM to 3 µM [107].

2.6. Sorption Processes

Although adsorption and ion exchange could be historically regarded as two different processes due to their many similarities (operating cycles and design concepts), they are often discussed together under the name of sorption [114]. Considered cost-effective, environmentally friendly, and technically mature, sorption is regarded as the most efficient way to remove dissolved metal ions from water and wastewater [84]. It has become increasingly common to treat industrial and municipal WW in sorption-contacting systems. As briefly described below, sorption is quite simple to implement at an industrial scale and enables the efficient removal of pollutants from solutions, even at very low concentrations.

In the case of HM, adsorption takes place when sorbate (HM) accumulates by physical or chemical means on the surface of the adsorbent in the form of a molecular or atomic film [115,116]. The mechanisms are different, and different analytical techniques could be used for their elucidation (adapted after [117]):

-

Titration—Performed to evaluate the ion exchange capacity of the sorbent or to identify available/total functional groups, etc.;

-

FTIR—To identify available and/or active functional groups on sorbent surface (an example is presented in Figure 9(a1));

-

XRD—Used for nanoscale analysis to reveal structure, composition or crystallinity of adsorbent before and after adsorption(an example is presented in Figure 9(a2));

-

XPS—To examine adsorbent surface chemistry and/or determine the content of selected elements;

-

SEM-EDX—Used for visualization of macro/microstructural characteristics of sorbent, elemental mapping and concentration evaluation (Figure 9(b1,b2,c1–c3), BSED—backscattered electron detector).

There are many factors to be considered when adsorption is used. Among them, the most commonly studied are pH, mixing/shaking time and intensity, contact time, the co-existing ions type and concentration, sorbent dosage, and temperature [118]. In many cases, the economy of the process is determined by the availability of adsorbent, costs, and regeneration options [115]. The only environmental issue related to adsorption is the disposal and management of the used adsorbents.

When conducting an ion exchange process, as shown here by Equations (6) and (7), ions of positive or negative charge in a liquid solution, usually water-based, replace similar and displaceable ions, known as co-ions, of the same or equivalent charge contained in a solid. So, by applying an ion exchange technique, the pollutant ions are replaced by others that are harmless to the environment. HMs specifically are attached to solid particles to replace the solid particle cations and are thus removed from the liquid phase [84]. Ion exchanges can be applied to remove HMs from raw water and WW using various types of ion exchange materials [119,120].

$$M^{-}{A}^{+}+B^{+} \leftrightarrow M^{-}{\mathrm{B}}^{+}+A^{+}$$

(6)

solid (IE) solution (HM)    solid (HM)  solution

$$M^{+}{A}^{-}+B^{-} \leftrightarrow M^{+}{\mathrm{B}}^{-}+A^{-}$$

(7)

solid (IE) solution (HM)    solid (HM)  solution

The solid structure of the adsorbent is not permanently altered and can be regenerated mostly for economic reasons. Matrix stability and reusability are the main factors for process feasibility.

Figure 9. Techniques used in the adsorption process evaluation: (a1) Comparative FTIR of walnut shell powder before and after Co(II) adsorption [121]; (a2) XRD patterns of titanate nanoparticles before and after Co adsorption (TC—tetracycline and NH₄⁺ ammonia ion, here as co-contaminants) [122]; (b1,b2) EDX pattern and elemental mapping of polyimide-chitosan fibers after Co adsorption [123]; (c1–c3) BSED image of cow bone char after Co adsorption (c1) used for EDS analysis (c2) and corresponding EDS spectrum (c3) [124].

Figure 9. Techniques used in the adsorption process evaluation: (a1) Comparative FTIR of walnut shell powder before and after Co(II) adsorption [121]; (a2) XRD patterns of titanate nanoparticles before and after Co adsorption (TC—tetracycline and NH₄⁺ ammonia ion, here as co-contaminants) [122]; (b1,b2) EDX pattern and elemental mapping of polyimide-chitosan fibers after Co adsorption [123]; (c1–c3) BSED image of cow bone char after Co adsorption (c1) used for EDS analysis (c2) and corresponding EDS spectrum (c3) [124].

Numerous studies have been reported on cobalt sorption using various materials, especially in the last few years since it is clear that the growing demand for cobalt will soon exceed its existing natural resources. In a comprehensive review published in 2018, Islam et al. reported innovative adsorbents used for Co²⁺ removal from water and WW [36]. This review focuses on the latest findings, shown synthetically and chronologically in

Table 1. Co sorption: metal ion concentration in the working solutions, sorbent dosage, experimental parameters (time, pH, temperature, and Co²⁺ initial concentration) and the best-obtained results (sorption capacity and sorption efficiency).

Adsorbent	Co2+ sol. conc. (mg/L)	Dosage (g/L)	t (h)	pH	T (°C)	Sorption Capacity (mg/g)	Sorption Efficiency (%)	Year and Ref.
Adsorption
Friedel’s salt	2–16 mM	4	24	4–6	25	217.2	> 99	2019 [125]
MIL-100(Fe)	100–200	1	1.75	8.9	38	119	93.4	2019 [126]
Activated carbon–alginate-Cyanex 272 spheres	0.5 mM	5	1	6–7	Ambient	12.4	96.7	2019 [127]
MIL-101-glycine MIL-101-diglycine MIL-101-triglycine	10–50	0.1	24	8.3–9	25–35	185 227 232	> 99	2019 [128]
Si/Ti–based amino-functionalized hybrids	10–1000	20	>10		25–45	354	N.S.*	2019 [129]
Chitosan and glucose-based hydrogel	100	0.2	3	7	20	286	74.8	2019 [130]
3-(((5-ethoxy-benzene-thiol)imino)methyl)–salicylic acid immobilized onto mesoporous silica	2–72	0.33	1–3	8	N.S.	185.23	N.S.	2020 [131]
Lignin deposited on nanoporous silica	2–200	2.8	1.5	~7	22	18.5	63	2020 [132]
Apatite from phosphate waste rocks	10–100	1	3	3–7	25	8.64	85	2020 [133]
Glycine functionalized silica particles	2–280	0.5	3	2–8	22/40/60	2.81 mM/g	84.7	2020 [134]
Two sediments from water bodies	2.5–50	10	38	2–8	25–60	0.40/0.93	45/96	2021 [135]
Magnetic composite of hydroxyapatite/Fe3O4/polydopamine	5–70	0.25–5	2.2	2–8	25–50	49.32	>90	2021 [136]
Microcrystalline cellulose- magnesium hydroxide	100–500	2.5	1	2–8	20–50	153.8	97.7	2021 [137]
Ethylenediaminetetraacetic dianhydride-modified orange peel	20–150	1	2	3–7	Ambient	40.48	N.S. *	2022 [138]
Natural apatite (nano-structural)	5–370	0.1–8	1	1.7–8	25–75	13.52	85	2022 [139]
Amide-functionalized metal-organic framework (TMU-24)	5–125	0.1	0.5	7	Ambient	500	N.S.	2022 [140]
δ-MnO2	80	10	1	7	80	N.S.	>80	2022 [141]
Chitosan	5 × 10−2–10−4 M	2	0.5	2–6	25	62	>99	2022 [142]
Nanocellulose from discarded palm leaves	100–600	20	1.5	5	20	5.98	>99	2023 [143]
Mesoporous activated carbon	5–200	10	0.25	5–6	25	5.8	66	2023 [144]
Cow bone/cow bone char/activated cow bone carbon	100–500	10	4	2–8	30	52.5/58.8/ 64.5	49.5/53/40	2023 [124]
Carbon dot/chitin nanocrystal	0–1000	100 mL/L	1	2–13	25–60	152	N.S.	2023 [145]
Ca(OH)2 modified quartz rock	60	5	2	2–10	25	47.1	>80	2024 [146]
Carbon xerogels derived from tannin	10–50	1	0.5	1–11	25	9.9	>97	2024 [147]
Adsorption and/through ion exchange
Lewatit TP260 (Resin)	10	0.07–0.5	5	1–5	20–60	62	99.5	2019 [148]
Single and mixed pillared clays	10–100	0.1	9	6–10	40	7–16	53	2019 [149]
Hydroxyapatite	0.25–4.2 mM	10	24	9	30	22.5	>90	2019 [150]
Faujasite NaY (Zeolite)	0.1–10 mM/L	0.15	12	5–6	25	0.92 mM/g	N.S.	2020 [151]
Mesoporous silica	5–50	0.3	8	3–9	25	6.62	88	2020 [152]
Nitrogen and sulfur-containing PVC-based polyampholyte	0.01–0.25 mol/L	1	10	5	30–50	3.45 mM/g	N.S.	2021 [153]
Composite of Fe3O4 and zeolitic imidazole framework-8	5–100	0.1–2	5	2–7	28	71.29	99.2	2021 [154]
Functionalized mesoporous carbon	0.1–0.7	0.1–0.5	0.2	2–10	20–30	1.59	95.5	2021 [155]
Poly(ortho-aminophenol) doped glycerol	10–100	0.4–4	1	2–10	Ambient	117.9	96	2022 [156]
Modified walnut shell	3000	0.5	1	3–7	Ambient	1371	> 86	2023 [121]
Chitosan-grafted polyimide fibers	500	1	5	2–7	20–60	46	>89	2024 [123]
Hydroxyapatite(HAP)/HAP dopped with 2-hydroxy-phosphonoacetic acid	25–200	0.5	2.5	2–10	25	74.5/171.4	>86	2024 [157]

* N.S.—not specified in the study.

. The list presented in Table 1 is just a representation of this technique’s scope and span and cannot be considered exhaustive in any way (radio-contaminated waters, waste solutions, and leaching liquors were not considered). The cited publications were published between 2019 and early 2024, highlighting the increasing interest in this specific metal removal from WW.

It is worth mentioning that the results always depend on the environment in which they were obtained and that metal ion sorption is affected by WW composition in almost all cases (other metal ion presence, coexisting organic compounds, and so on). As seen in the literature, most of the available studies focus on mono-elemental solutions, and only a few approach complex adsorption in multi-component WW. An interesting report was recently published regarding the possible effects of cobalt-containing WW treatment. Thus, it was shown that coexisting elements in multi-component matrices could lead to synergistic, neutral, or antagonistic effects in simultaneous adsorption, depending on the case [122].

In a comprehensive review of adsorption innovations and challenges, Satyam and Patra (2024) emphasized that “the economic viability of adsorbents is intricately linked to their synthesis and operational costs” [158]. In the process of obtaining a novel sorbent for water and WW treatment, the technical performances are the first to be evaluated: stability in aqueous systems, reaction rates/adsorption kinetics, selectivity for certain pollutants, and adsorption capacity. A comprehensive study will also address at least the desorption capacity; regeneration and reusability; and (bio)degradation and preservation of the adsorption efficacy.

Following a technical assessment, numerous innovative adsorbents proved suitable for the application, as in the case of all the materials presented in Table 1. However, from the laboratory to industry, other steps must be pursued, and most of the investigations lack cost analyses in general. Thus, economic viability, i.e., aimed at low costs for high performance, must be further proved as the second step. In this respect, efforts are needed to unify the concepts of adsorbent costs and evaluation criteria for process effectiveness. Adsorption economic viability should be evaluated based on clear equations, counting chemicals and energy costs, and adsorption capacity together [159]. Last but certainly not least, the new biosorbent must fulfill all regional regulatory and environmental requirements, and in all cases, these requirements refer to adsorbent production methods, use, and disposal/recyclability. Of course, the implementation of regulations will strongly impact on process economy. Overall, the adsorbent needs to provide “the least possible environmental impact” and the “best technical and economic performance”, as synthetically presented in Figure 10 [160].

Figure 10. Consideration in developing effective and feasible sorbents [160].

Figure 10. Consideration in developing effective and feasible sorbents [160].

In any case, a short and limited analysis of the adsorbents presented in Table 1, used for Co removal from waster and WW, will show the researchers’ clear tendency towards “green” synthesis methods, “renewable” resources, “eco-friendly” solutions in developing high adsorptive biodegradable, and/or reusable materials.

Considering the possible drawbacks of conventional adsorption, as briefly presented above, a newer, developed method offered a “green” perspective—biosorption, defined as “metabolism-independent adsorption of compounds to the surface of the microbial cell wall of both living and dead microorganisms” [161]. Since this method has gained significant interest lately, it was not considered in this study.

2.7. Photocatalysis

As a relatively new technique, photocatalysis allows for the photo-reduction of HM compounds with the use of photocatalysts (usually a semiconductor structure) irradiated by light. Highly reactive radicals formed in the process can provide fast or very fast breakdown of recalcitrant metal complexes under UV or visible light with minimal production of secondary pollution [162]. Although this method is considered the most eco-friendly (as can be seen in Figure 11), there are still many obstacles in future applications, especially at the industrial scale, including photocatalyst efficiency, light utilization efficiency, and photocatalyst separation and regeneration [163,164].

Figure 11. A comprehensive comparison between various methods for HM removal [84].

Figure 11. A comprehensive comparison between various methods for HM removal [84].

Despite its impediments and difficulties, there are specific conditions when only photocatalytic oxidation can provide acceptable treatment objectives/results. This is the case, for instance, of ternary precursor WW resulting from ternary cathode materials manufacturing. With high levels of ammonia and metal ammonia complexes, managing these WWs proved challenging when any other technique has been applied because of secondary pollution or additional operational and chemical requirements. Following the successful preparation of a simple yet performant photocatalyst, Li et al. (2024) reported a feasible photocatalytic treatment of such WW [165]. The obtained catalyst consists of Al and N co-doped carbon as carriers for nano-TiO₂ (ANC/T), characterized by a specific area of over 360 m²/g with a large number of active sites and pore volume as high as 0.35 m³/g. The degradation mechanism is complex and consists of two major steps: (i) high-performance contaminant adsorption on the ANC surface and (ii) effective photocatalytic degradation of pollutants around the active component (TiO₂) under sufficient light energy to N₂ as the main product and ion metals. The ion metals (here Ni²⁺ and Co²⁺) were then adsorbed on the catalyst surface. Their experiments, using a high-pressure mercury lamp (125 W) and 3 g/L photocatalyst on concentrated WW (ammonia nitrogen 385.3 mg/L, Ni around 22 mg/L, and Co 15 mg/L) proved that 1.5 h of dark adsorption and 5 h of illumination yielded more than a 91% degradation of ammonia nitrogen, the degradation of Ni/Co complexes, the removal of more than 98% of metals.

Another similar study on WW resulting from activities related to Li-ion battery production was recently reported, proving photocatalytic technology’s effectiveness. The following results were obtained related to cobalt: nickel–cobalt-ammonia complex degradation with more than 97% cobalt ion removal by adsorption on the catalyst surface using 2 g/L MoS₂/g-C₃N₄ heterojunction photocatalysts after one hour of dark adsorption followed by four hours of illumination (125 W) [166].

3. Challenges and Future Perspectives

Contaminant removal from water and WW is not a new problem. Even though it has been addressed and studied for many years, it continues to provide new challenges as concerns about environmental and health issues arise every day. Despite all of the measures taken so far, it is clear that pollution is increasing as industry and society proliferate in many ways. Emerging pollutants are much more often identified, and emerging treatment technologies, together or alongside conventional methods, are tested or enforced to minimize their harmful effects. It is the case of Co, for which potential environmental risks and impacts on human health were considered to be low or neglected until recently.

The evaluation of existing methods that allow for Co recovery/removal from WW shows no positive or negative solutions exist. Each method has its pros and cons (as represented in Figure 11 and

Table 2. Comparison between different methods for HM removal in general (and Co in particular).

Method	Advantages	Limitations
Precipitation	Simple operation Technically mature and practical Applicable for a high concentration of HM Low capital cost, cost-effective operation	Subsequent separation of sludge is needed (sedimentation) Large quantities of sludge are generated Low selectivity Extra cost for chemical reagents and sludge disposal
Electrochemical	High recovery No chemicals required No secondary pollution is generated Available for large-scale applications	Relatively expensive technology Requires (expensive) electrodes High electrical energy consumption
Membrane separation	Mature as it is practical High metal ions extraction efficiency	Membrane fouling and biofouling Pretreatment operations are needed High costs: capital, maintenance, operation Concentrate streams for disposal/recycling/treatment
Coagulation/ Flocculation	High efficiency Simple operation Accessible inorganic precipitants	Low selectivity Extra cost for chemical reagents and sludge disposal Other treatment is required to achieve low HM concentration
Flotation	High efficiency Low sludge production	Toxic sludge for disposal High costs: capital, maintenance, operation Other treatment is required to achieve low HM concentration
Adsorption	Simple operation Strong applicability Allows selectivity	Adsorbents disposal Effectiveness strongly depends on adsorbent
Ion exchange	Low capital cost High efficiency	Varied selectivity depending on resin High maintenance and operation costs Pre-treatment needed
Photocatalysis	No (or less) chemical consumption Eco-friendly Less sludge production Clean products	High capital and operation costs Possible formation of by-products

).

In an individual assessment, each method can be further improved through innovative research. For instance, efficient regeneration processes for adsorbents should be developed, novel membrane materials with enhanced capacities are needed, and so on [167]. Green chemicals, green synthesis routes, or eco-friendly processes need to be implemented more and more often. Energy-efficient systems are required to provide cost-effectiveness for higher removal rates.

Most importantly, disregarding any economic and technical challenge, treatment methods should, in any case, guarantee water treatment according to increasingly stringent quality standards enforced worldwide. The key to meeting these regulations is represented by the hybrid/multistage processes, meaning that WW treatment will be achieved through the combination of different techniques. A good example is the separation and recovery of Co from real, heavily polluted WW in an integrated system, including diffusion dialysis, precipitation, liquid–liquid extraction and stripping, as depicted in Figure 12 [45].

Figure 12. Operation sequence to separate HMs from industrial effluent [45].

Figure 12. Operation sequence to separate HMs from industrial effluent [45].

4. Conclusions

This work represents a brief overview, but it also provides an in-depth look at the problem of HM pollutants generated in the environment by anthropogenic or natural sources. The toxic effects on human populations and the potential health hazards of HM require continuous studies, experimentally and by mathematical modeling, with respect to controlling the occurrence of these pollutants and their distribution in all environmental compartments.

It can be concluded that if treatment technologies for HM-polluted water and wastewater exist, they are mature and well-developed for higher or lower levels of HM concentration. Nevertheless, high specificity and a customized solution for each aqueous system based on jar tests and field investigations are needed, especially when new elements of concern are investigated, as in the case of cobalt.

Although well-known techniques like adsorption, chemisorption, or membrane separation prove effective in terms of technical and economic performance, it is clear that the future belongs to integrated processes that will allow for heavy metals (of any kind) removal to more and more stringent treatments and recovery targets with minimal drawbacks at an industrial scale.

Most importantly, it must be mentioned that it is also crucial to minimize emissions, as continuous sources for heavy metal release are in all environmental compartments. Five major directions aiming at emissions reductions can be described. The first direction is to use green(er) technologies to develop processes and materials. The second direction is energy from the sun, water, and wind, urging towards the development and implementation of more and more non-polluting energy solutions. The third direction, the use of alternatives to open burning, starts with the consideration that open burning is an important source of particulate matter (a major source of heavy metals). The fourth direction, drive less, considers that driving a car is likely a person’s single most polluting daily activity. Driving less reduces the amount of vehicles/time on the road, which helps to reduce air pollution. The fifth direction, drive smart, refers to better solutions/alternatives to operate our vehicles to keep reducing pollution.