Microalgae in Mitigating Industrial Pollution: Bioremediation Strategies and Biomagnification Potential

Abstract

The rapid growth of the human population and industrialization has intensified anthropogenic activities, leading to the release of various toxic chemicals into the environment, triggering significant risks to human health and ecosystem stability. One sustainable solution to remove toxic chemicals from various environmental matrices, such as water, air, and soil, is bioremediation, an approach utilizing biological agents. Microalgae, as the primary producers of the aquatic environment, offer a versatile bioremediation platform, where their metabolic processes break down and convert pollutants into less harmful substances, thereby mitigating the negative ecological impact. Besides the CO₂ sequestration potential, microalgae are a source of renewable energy and numerous high-value biomolecules. Additionally, microalgae can mitigate various toxic chemicals through biosorption, bioaccumulation, and biodegradation. These remediation strategies propose a sustainable and eco-friendly approach to address environmental pollution. This review evaluates the microalgal mitigation of major environmental contaminants—heavy metals, pharmaceuticals and personal care products (PPCPs), persistent organic pollutants (POPs), flue gases, microplastics, and nanoplastics—linking specific microalgae removal mechanisms to pollutant-induced cellular responses. Each section explicitly addresses the effects of these pollutants on microalgae, microalgal bioremediation potential, bioaccumulation process, the risks of trophic transfer, and biomagnification in the food web. Herein, we highlight the current status of the microalgae-based bioremediation prospects, pollutant-induced microalgal toxicity, bioaccumulation, and consequential biomagnification. The novelty of this review lies in integrating biomagnification risks with the bioremediation potential of microalgae, providing a comprehensive perspective not yet addressed in the existing literature. Finally, we identify major research gaps and outline prospective strategies to optimize microalgal bioremediation while minimizing the unintended trophic transfer risks.

1. Introduction

Environmental pollution remains a critical ongoing global issue, posing one of the greatest challenges toward human health and ecosystem stability [1,2,3]. The detrimental effects include severe health disorders, habitat destruction, biodiversity loss, and climate change. Since the Second Industrial Revolution in the 1870s, rapid industrialization and anthropogenic activities have accelerated the pollution levels exponentially [4]. Despite increasing global awareness, its long-term impacts remain profound, and yet in many cases, unpredictable. Therefore, developing sustainable solutions to address the pollution is of prime importance.

Environmental pollutants may be organic or inorganic in nature, arising from various sources like fossil fuels, chemical, agricultural, nuclear, and many other industries. Key categories include heavy metals, pharmaceuticals, personal care products, persistent organic pollutants, industrial gases, microplastics, nanoplastics, dyes, surfactants, etc. Although conventional treatment methods—primarily mechanical and chemical—are widely used, they often lack sustainability [5]. This highlights the importance of a sustainable solution for reducing the pollutant toxicity while promoting environmental restoration.

Biodegradation could be considered as nature’s recycling mechanism for degrading pollutants to less toxic or nontoxic compounds. This method depends on using living organisms to break down or eliminate pollutants. The biodegradation of environmental pollutants is performed by various plants, bacteria, fungi, microalgae, macro algae/seaweeds, etc. [6,7,8,9,10,11,12,13,14].

Being primary producers, microalgae have an essential role in aquatic ecosystems [15]. The term microalgae include both eukaryotic microalgae and prokaryotic cyanobacteria, with the potential for CO₂ fixation, N₂ fixation, and oxygen production by photosynthesis [16,17].

Microalgae have emerged as a preferred bioremediation system due to its multifaceted advantages over other biological systems, integrating complementary removal mechanisms [18]. These mechanisms, including biosorption, biodegradation, bioaccumulation, photodegradation, etc., enable the removal of a wide range of industrial pollutants [1,2]. Unlike bacteria, fungi, or plants, microalgae possess a high surface to volume ratio that enables the efficient interaction of the microalgae cell walls’ associated functional groups (carboxyl, hydroxyl, amino, phosphate etc.) with various pollutants. Additionally, microalgae’s extreme tolerance capacity to salinity, temperature, nutrient stress, and heavy metal concentrations makes them a suitable bioremediation system for treating industrial effluents [3,4]. Furthermore, their carbon sequestration capacity helps to mitigate the carbon footprint by removing the pollutant gases from the environment [19]. Microalgae exhibit metabolic versatility and operate under heterotrophic, photoautotrophic, or mixotrophic conditions with extreme operational flexibility [5,6]. Compared with terrestrial plants, they offer fast growth rates, require minimal land for cultivation, and valorize the biomass to value-added products (biomolecules, biofuels, bio oils, feed, biochar), thus supporting a sustainable circular economy model [7,8,9,20]. When utilizing microalgae, heavy metals are removed by ion exchange, complex formation, and precipitation mechanisms. PPCPs are mostly degraded by photodegradation pathways whereas POPs are degraded by mono oxygenases and dioxygenase enzymes. Micro- and nanoplastics are removed via biosorption, bioaccumulation, and biotransformation along with extracellular polymer-induced aggregation [10,11,12,13]. Integration of other microbial remediation systems like microalgae with a bacteria consortium is more efficient in removing various pollutants [14].

Integrated microalgal biorefinery is an excellent concept for the biodegradation of various industrial pollutants. It also contributes to a higher energy output with positive economic and environmental impacts [21]. Thus, such systems can sustainably handle various industrial side streams.

In terrestrial C3 plants, utilizing the field-based light, photosynthetic efficiency measurements often result in yield values of 1–2% [22] whereas microalgae show a 10–20% photosynthetic efficiency from lab and optimized photobioreactor studies conducted under similar continuous light and nutrient conditions, minimizing photo inhibition and light saturation [23]. Being photosynthetic organisms, they have an excellent capacity for the biofixation of CO₂ from various sources such as the atmosphere, industrial exhaust gases, soluble carbonate salts, etc. [24,25,26].

Apart from the biofixation of CO₂, algae are used for the production of biomaterials, biofuels, high-value products, biofertilizers, biostimulants, biohydrogen, functional foods, bioplastics, etc. [27,28,29,30,31]. The high-value biomolecules from microalgae include pigments, carbohydrates, proteins, lipids, vitamins, and antioxidants. Microalgal products are used in food, pharmacy, cosmetics, nutraceuticals, agriculture, feed, etc. [32,33].

Microalgal cultivation is influenced by the water quality. The impacts of global pollution have critical effects on water quality and microalgal growth. The majority of pollutants in aquatic systems are anthropogenic chemicals or emerging contaminants that are capable of causing acute/cumulative toxicity, triggering cancer or endocrine disruptions [34,35,36]. To sustainably remove these contaminants, microalgae are utilized globally due to their bioremediation potential [37].

This review focuses on the recent investigations in the field of various prominent pollutants capable of influencing the environment and their interaction with microalgae, such as toxicity responses, bioremediation potential, and biomagnification possibilities, where toxins are transferred to higher trophic levels.

2. Bioremediation by Microalgae

The bioremediation process by algae is called phycoremediation, where algae biotransform various pollutants into less toxic chemicals [38,39]. Algal bioremediation includes biosorption, bioaccumulation, and biodegradation of the toxic components, as shown in Figure 1. The various bioremediation steps are displayed in the schematic along with the type of interaction as an active or passive process, the location of the process, and the chemical reaction. The three-stage enzymatic reaction enzymes are also listed in the schematic.

In Figure 2, a detailed cellular mechanistic diagram is displayed, explaining the enzymatic stages and various biodegradation processes involved. Examples of two different pollutants—enrofloxacin and imidacloprid—are described therein.

Phycoremediation impacts vary significantly between freshwater and marine systems due to differences in algal physiology, environmental conditions, and pollutant characteristics. Freshwater microalgae, like Chlorella and Scenedesmus, prefer nutrient-rich and low salinity water bodies, where they display high growth rates, strong heavy metal biosorption, and bioaccumulation mediated by cell wall functional groups and extra cellular polymers [29]. However, marine microalgae, such as Phaeodactylum tricornutum, Nannochloropsis, and Dunaliella, are adapted to high-salinity, nutrient-variable conditions such as eutrophication and competing cation conditions that alter metal speciation. These adaptations, while sometimes reducing the biosorption efficiency, offer resilience to salinity fluctuations and enhance the biodegradation of organic pollutants [40]. Considering the growth conditions and cost, freshwater species are more cost-effective and easier to cultivate as they require simple media and exhibit strong efficiencies in removing pollutants such as heavy metals and pharmaceuticals. Marine species, in contrast, thrive under saline conditions, and their high lipid content makes them particularly valuable for biofuel production and other high-value bioproducts. Recognizing these complementary traits is essential for selecting the most suitable system based on both the pollutant type and local environmental context [40,41,42]. Diatoms are present in both freshwater and saline environments and have potential in pollutant removal through biosorption and active assimilation pathways [43]. Although the fundamental mechanisms of phycoremediation remain similar across systems, differences among freshwater and marine taxa necessitate tailored, species-specific strategies to achieve maximum effectiveness.

The main two categories of biodegradation are metabolic degradation and co-metabolism. Biodegradation can occur either extracellularly or intracellularly, or through a combination of both. This process involves the extracellular breakdown of toxins to intermediate forms, and then the intracellular disintegration of these intermediates. Abdelfattah et al. [44] reported that extracellular degradation is often facilitated by extracellular polymers (EPSs) produced by microalgae.

The microalgal biodegradation of organic pollutants is a three-stage complex enzyme-mediated process. As illustrated in Figure 1, Stage 1 involves cytochrome P450 enzymes that increase the pollutant hydrophilicity either by hydrolysis or oxidation—reduction reactions. Stage 2 involves conjugation enzymes that enhance the binding of glutathione with different compounds, thus protecting the cell against oxidative damage. Stage 3 employs detoxification enzymes to biotransform the molecules into less or nontoxic intermediates [44]. Furthermore, microalgal biomass offers additional economic advantages including applications in biofuel production, fertilizer synthesis, animal/aquaculture feed, etc. [37]. The presence of lipids with polyunsaturated fatty acids (PUFAs), proteins containing essential amino acids, carbohydrates, pigments, and other bioactive compounds makes microalgae an economical and sustainable aquaculture feedstock [45]. The establishment of microalgal biorefineries facilitates the valorization of industrial side streams, thereby offering a sustainable pathway toward a circular economy with minimal or zero waste generation [46,47].

3. Industrial Pollutants

The major types of industrial pollutants discussed in this review are heavy metals, pharmaceuticals, personal care products, persistent organic pollutants, industrial gases, microplastics, and nanoplastics. Figure 3 displays a schematic representation of the industrial pollutants and corresponding microalgal responses to these pollutants. Depending on the pollutant characteristics, the algal cell responses often lead to cytotoxicity and growth inhibitions. An overview of the different processes involved is also displayed such as biosorption, bioaccumulation, and biodegradation. Furthermore, the potential of biomagnification and trophic transfer to higher organisms are also schematically represented here.

3.1. Heavy Metals

Heavy metals are among the most significant industrial contaminants. They comprise metals and metalloids with a high atomic weight and a higher density than water (< 5 g/cm³) such as arsenic (As), chromium (Cr), cadmium (Cd), copper (Cu), iron (Fe), lead (Pb), mercury (Hg), silver (Ag), platinum (Pt), and zinc (Zn). Heavy metals possess numerous risk concerns due to their toxicity potential (even at parts per billion levels) and carcinogenic nature. The nonbiodegradability and bioaccumulative nature of heavy metals makes them more detrimental to the environment and ecosystems. Water pollution by heavy metals is a global challenge requiring serious public attention [48,49]. Metal released into water bodies affects the aquatic environment by disturbing the growth, metabolism, and reproduction of organisms, which will ultimately affect the entire trophic chain including humans. In aquatic environments, metals exist as micropollutants, ionic forms, dissolved complexes, and attached to inorganic matter or dead organisms [50]. To evaluate the toxicity levels in an aquatic environment, aquatic organisms are used as indicators [51,52].

Heavy Metals—Microalgal Interactions and Bioremediation

Heavy metals are persistent pollutants that accumulate in food webs because they cannot be degraded. Microalgae offer natural resilience, binding and detoxifying metals through biosorption, chelation, and intracellular sequestration, often supported by extracellular polymers and stress responsive peptides [53]. Microalgae have good metal binding capacity; they accumulate it in cells and transfer it to the next trophic level [54]. The metal uptake occurs via surface adsorption followed by endocytosis, passive diffusion, mediated transport (by amino acid transporters), precipitation, ion exchange, complexation, and chelation (e.g.,: metal link to Ca²⁺ channel ion carriers) [55].

Exposure to heavy metals often induces oxidative stress in microalgae, leading to the production of various biomolecules like antioxidants and enzymes such as ascorbate peroxidase, catalase, glutathione peroxidase, superoxide dismutase, etc. Additionally, the toxicity effect of heavy metals on algae often creates the production of reactive oxygen species (ROS) and associated cellular redox changes. Increased levels of ROS result in cell disruption via lipid peroxidation, protein and sugar oxidation, or nucleic acid damage [55]. Depending on the heavy metal type, microalgae can create hydroxyl radical, superoxide anion, hydrogen peroxide, and singlet oxygen. When the ROS responses exceed the cell capacity to manage, they deteriorate the algal cells.

Different microalgal genera exhibit distinct tolerance mechanisms to various ROS types and reflect genus-specific cellular tolerance, stress responses, and signaling pathways. For ROS-induced stress, marine microalgae show higher reactivity compared with freshwater species [21]. Cyanobacterium Anabaneba displayed tolerance to selenium compound exposure, selenite VI at 0.4 mM and selenite IV at the 0.1 mM level was found to be detrimental, affecting photosynthesis and enhancing endogenous ROS [56]. In diatoms (Phaedactylum tricornutum), hydrogen peroxide (H₂O₂) triggers dynamic chloroplast calcium signaling [57,58], while in Chlamydomonas reinhardtii, singlet oxygen mediated photosynthetic gene regulation [16]. Superoxide and hydroxyl radicals lead to severe oxidative stress, as shown in Raphidocelis subcapitata, where copper exposure-induced ROS levels were up to 33 times with enhanced catalase activity [59]. Likewise, sulfate radicals disrupted cell integrity in Microcystis aeruginosa [60]. Antioxidant tolerance/protection varies among genera: Chlamydomonas reinhardtii uses singlet oxygen as a signaling molecule for upregulating photoprotection, while Chlorella ohadii displays tolerance through antioxidant accumulation and improved cyclic electron flow [61,62]. Scenedesmus sp. enhances enzymatic upregulation with catalase [26], while Chlorella sorokiniana undergoes thylakoid remodeling with the downregulation of ROS quenching enzymes [63]. These genus-specific responses reflect evolutionary diversification in an antioxidant system with α-carotene derivatives specific to Chlorophyta, whereas diadinoxanthin and fucoxnthin are present in Heterokontophyta, Haptophyta, and Dinophyta [64].

Irrespective of the interactions that occur between cells and metals, phycoremediation depends on several factors: pH, temperature, metal concentrations, metal speciations, charges, presence of other metals, biomass concentrations, and regeneration/reuse potential [65]. The influence of arsenic (As(V), As(III)) and cadmium (Cd) ions and their toxicity on Chlamydomonas acidophila showed results of ROS generation, internal cellular damage, and intracellular metal depositions. TEM-XEDS characterization confirmed bioaccumulation and biosorption of the metals as intracellular deposits, with a high phosphorous content due to the polyphosphate metal resistance action. However, microalgae were sensitive to Cd but tolerant to the more toxic As [66].

Heavy metals can also impair photosystems. Photosystems (photosystem 1—PS1 and photosystem 2—PS2) are pigment–protein complexes involved in the photosynthesis process. PS1 is a chlorophyll protein complex (plastocyanin–ferredoxin oxidoreductase) that accepts electrons from plastocyanin and transfers them to ferredoxin, enabling NADPH production [67]. PS2 is a multi-subunit pigment–protein complex (water–plastoquinone oxidoreductase) in the thylakoid membrane that catalyzes water oxidation, releasing O₂, protons, and electrons and transfers electrons to plastoquinone [68]. PS1 activity is evaluated by monitoring the P700 redox state by the absorbance change at 820 nm, and the PS2 activity is evaluated by chlorophyll fluorescence measurements. The inhibition of PS1 and PS2 is a reflection of electron transport disruption and photochemical efficiency reduction. The growth and photosystem (PS1 and PS2) activities of Chlorella pyrenoidosa in the presence of cadmium (Cd) showed a decreased growth rate with the inhibition of photosynthesis. When compared, the Cd concentration caused more inhibition on PS2 compared with PS1 [69]. Raphidocelis subcapitata, in the presence of cadmium (Cd) and cobalt (Co), was evaluated by the growth rate, lipid quantification, chlorophyll content, quantum yield, and efficiency of the oxygen-evolving complex. Both Cd and Co influenced growth, increased free fatty acids, and reduced oxygen-evolving complexes [70]. Mucidosphaerium pulchellum and Micractinium pusillum were tested in the presence of chromium—Cr(VI), where it displayed concentration and exposure time-dependent photosynthesis and lipid accumulation in terms of hexadecenoic acid as well as ω3, ω6, and ω7 fatty acids [71]. R. subcapitata showed enhanced growth at higher temperatures (20–30 °C), which increased the water pH, where copper became toxic to the microalgae [72]. When tested on three different microalgae, iron oxide and zinc oxide engineered nanoparticles (ENPs) displayed morphological responses, internalization, and potential impacts on physiology and metabolism. Both types of ENPs affected Paracentrotus lividus larval development, but ZnO ENPs had a much stronger effect. The metal oxides inhibited the growth of the algae Micromonas commode, while the growth of Ostreococcus tauri or Nannochloris sp. was found to be unaffected [73].

For the influence of mercury, lead, and cadmium toward Nannochloropsis oculata, we inspected the acute and sub-acute toxicity profiles. The acute toxicity results showed that mercury was about twice as toxic as lead and about 5.7 times more toxic than cadmium. Lead showed 2.7 times more toxicity than cadmium. The chlorophyll and carotenoid contents were found to be reduced during the sub-acute toxicity analysis due to the bioaccumulation [74]. Chlorella vulgaris was tested in the presence of silver nanoparticles to evaluate the toxicity responses on the physiology and morphology of the cells. A time-dependent toxicity and growth inhibition was observed with a change in the pigment production and morphology. Furthermore, the biomolecule content of the cells (lipid, protein and carbohydrate) was altered [75]. The influence of platinum on Dunaliella salina exhibited growth inhibition and the toxicity increased with time [76].

Mercury (Hg) toxicity has been specifically studied. Fifteen (15) various marine microalgae were tested to evaluate the Hg toxicity, where six (6) species (Amphidinium carterae, Chaetoceros curvisetus, Prorocentrum donghaiense, Prymnesium parvumcarter, Prorocentrum triestinum, and Phaeodactylum tricornutum) displayed the demethylation of methylmercury (MeHg). The MeHg demethylation rate was of the same order of magnitude as that of photo demethylation (i.e., photo-induced demethylation) [77]. Similarly, Chlorella sorokiniana, when treated with organic Hg, biosynthesized Hg–phytochelatins as a defense mechanism during the exposure [78]. To evaluate the dynamics of MeHg in the marine environment, MeHg was introduced to four (4) marine microalgae lineages. When compared with Pelagophyceae, Prymnesiophyceae, and Bacillariophyceae (diatom), Cyanophyceae displayed a higher cellular MeHg accumulation. Also, live cells showed more MeHg accumulation than dead cells, and diatom cells maintained the MeHg without releasing it during cell division and the stationary phase, which supports the critical role of microalgae’s metabolic activity in MeHg biomagnification in marine food webs [79].

Bioremediation by microalgae consists of two phases: (1) biosorption—a rapid passive absorption (extracellular), and (2) bioaccumulation—a slow positive diffusion and accumulation (intracellular). During the biosorption process, metal ions are bound to the cell surface via electrostatic interaction. This is regulated by the pH, presence of other ionic entities’ functional groups, and the cell surface composition. Microalgal cell walls are made of peptides, polysaccharides, exopolysaccharides, lipids, proteins, high-affinity metal binding functional groups, etc. Microalgae possess cell surface binding groups like carboxylates, aminos, amides, and hydroxyls with heavy metal binding capacity. The chemical structure of the biosorbent could identify the heavy metal ion affinity [80,81]. Heavy metal removal by microalgal species works through different extracellular and intracellular mechanisms [82]. The influence of As(III) toward the growth and antioxidative responses of Chlorella thermophila SM01 and Leptolyngbya sp. XZMQ was investigated. An enhancement in the antioxidant was observed after 3 days of treatment, where longer exposure decreased the levels of the antioxidants, possibly indicating the As tolerance of the cells. Within 2 days of exposure, C. thermophila converted 94.37% of As(III) to As(V), while Leptolyngbya sp. accumulated ~2429.90 μg As/g dry weight. The dry weight of both microalgae increased due to the adsorption of As, which was reflected by the shift in the presence of various functional groups in the FTIR analysis, indicating a good bioremediation potential [83]. Microalgae-based heavy metal remediation has the advantages of profuse availability, low cost, eco-friendliness, commercial viability, and exceptional metal removal competence [84]. Contrary to various conventional remediation processes, microalgae utilize heavy metals as a nutrient source for their metabolic processes. Additionally, heavy metal removal efficiency can be tuned by the immobilization of algal cells, by creating an algal consortium, or by developing microalgae-based nanocomposite materials [85].

Molecular level evaluations of microalgal remediation strategies include metal binding peptides (metallothioneins, phytochelatins), antioxidant enzyme systems (superoxide dismutase, catalase, glutathione reductase), and specific metal transporters mediating pollutant uptake, sequestration, and detoxification. Reports on genomics, transcriptomics, and proteomics studies have revealed that metal exposure induces differential gene expressions of proteins involved in oxidative stress mitigation and intracellular compartmentalization while synthetic biology approaches improve the microalgal capacity for heavy metal removal [33,86]. For example, under Cd stress, Chlorella pyrenoidosa revealed the activation of sulfur metabolism and peroxisome linked genes, inducing the synthesis of phytochelatins and metallothioneins, key chelators of Cd sequestration [87]. Time-based metabolomics analysis revealed microalgal responses including the upregulation of extracellular polymeric substances, compartmentalization of Cd to chloroplasts, cytoplasm and vesicles, and transient modification in photosynthetic and antioxidant pathways to restore metabolic equilibrium [88]. In the case of As, microalgae enable detoxification through enzymes like arsenate reductases and methyl transferases, facilitating the reduction of As(V) to As(III), and following methylation into organic As species and intracellular accumulation mediated by aquaporin-like transporters [89].

Table 1. Microalgal responses to heavy metals.

Pollutant Name (Heavy Metals)	Microalgae	Observed Activity and Removal Efficiency	Analysis * Performed	References
Cu, Ni	Mixed culture (Spirogyra, Chlamydomonas, Eudorina, Spirulina)	Adsorption mechanism Removal efficiency: 90.20% Cu and 78.25% Ni.	Growth kinetics, adsorption efficiency, Dry weight analysis, SEM analysis and thermodynamic analysis.	[90]
As(III)	Biofilm (Chlorella sp., Cladophora sp., Microspora sp., Gomphonema sp., Nitzschia sp., Navicula sp., Scenedesmus spp.)	Biotransformation and bioaccumulation. 41–63% bioaccumulation.	Biofilm evaluation, biotransformation study, bioaccumulation evaluation if fish models, arsenic content analysis in serum and tissue, antioxidant gene expression analysis by PCR.	[91]
Cu, Mo	Chlorella vulgaris, Scenedesmus spinosus	Biosorption; C. vulgaris removal Cu 64.7%, Mo 99.9%, S. spinosus removal Cu (55%), Mo (80.3%).	Tolerance bioassays, bio removal assays, SEM-EDX, CLSM, ash content, lipid analysis, GCMS analysis.	[92]
Cd, Ni, Pb	Chlorella vulgaris, Scenedesmus quadricuda, Spirulina platensis	Bioremediation. The consortium removed Pb 89%, Cd 88% and Ni 95%.	Growth kinetics.	[93]
Cd, Pb, Cu	Spirulina platensis, Chlorella vulgaris	Biosorption. Removal efficiency: S. platensis Cd 47.84%, Pb 47.39% and Cu 45.04% C. vulgaris Cd 48.54%, Pb 48.30% and Cu 47.72%.	FTIR analysis, adsorption kinetics.	[94]
Cd, Cr.	Parachlorella kessleri Bh-2	Bioaccumulation. Removal efficiency: 94.80% removal.	Genomic DNA extraction, PCR sequencing, phylogenetic analysis, AAS, TEM, polysaccharide analysis, toxicity analysis.	[95]
Cu, Cr, Pb, Cd	Chlorella vulgaris	Removal efficiency: 79% Cr, 93% Cd, 72% Cu, and 79% Pb.	Growth kinetics, dry weight analysis, decolorization assay.	[96]
Zn, Cd	Chlamydomonas reinhardtii	Biosorption; bioremediation.	Cellular tolerance analysis, cellular uptake analysis.	[97]
Cu2+, Na+, Ca2+, Fe3+	Chlamydomonas microsphaera	Biosorption.	Absorption kinetics, SEM, EDAX, absorption isotherm, pH influence study.	[98]
Cd2+, Cu2+, Ni2+, Pb2+	Costaria costata, Hizikia fusiformis, Gracilaria verrucosa, Codium fragile	Biosorption. Highest adsorption capacity by C. costata.	Solution chemistry effect, SEM, FTIR, and kinetic sorption analysis.	[99]
Hg2+	Nannochloropsis sp.	Bioaccumulation; The maximum adsorption rate 6.96 μg/(g·day) at day 1. For 0.7 ppb concentration of Hg2+ 50% adsorption in 30 min.	Kinetics, morphology by SEM, adsorption isotherms, FTIR spectrometry, fatty acid analysis by GC-MS.	[100]
Cu	Nannochloropsis oculata	Cu remediation; adsorption. Removal efficiency: 99.92%.	Cell density, growth rate analysis, lipid and fatty acid profile analysis, FTIR spectrometry.	[101]
Cu	Nannochloropsis sp.	Absorption. The adsorption capacity of Cu(II) 5.32 × 10− 1 mmol/g.	FTIR, SEM, EDX, XRD, particle size, charge, adsorption kinetics, sequential desorption.	[102]
Cu, Cd	Chlorella salina, Nannochloropsis salina	Bioremediation.	Growth inhibition, protein pattern analysis, gene analysis (RAPD-PCR).	[103]

* SEM—scanning electron microscopy; PCR—polymerase chain reaction; GC-MS—gas chromatography-mass spectrometry; EDX—energy-dispersive X-ray spectroscopy; CLSM—confocal laser scanning microscopy; AAS—atomic absorption spectroscopy; TEM—transmission electron microscopy; FTIR—Fourier transform infrared spectroscopy; XRD—X-ray diffraction; RAPD-PCR—random amplified polymorphic DNA-polymerase chain reaction.

describes the interactions of various algal strains with different heavy metals [90,91,92,93,94,95,96,97,98,99,100,101,102,103].

While metals readily bioaccumulate in aquatic organisms, concerns regarding their potential to biomagnify along trophic levels require careful consideration.

3.2. Pharmaceuticals and Personal Care Products (PPCPs)

Pharmaceuticals and personal care products (PPCPs) include various chemical substances such as drugs, cosmetics, disinfectants, skin care products, fragrances, cleaning products, insect repellants, daily use household chemicals, etc. PPCPs are mostly biologically active, hormone-disruptive, and toxic; they create serious problems for the aquatic environment worldwide [104]. PPCPs are persistent and bioaccumulative in nature. In water bodies, they can exist as acidic, neutral, and basic groups. PPCPs include both human and veterinary medicines. As shown in Figure 4, PPCPs consist of antibiotics (chloramphenicol, ciprofloxacin, sulfamethoxazole, trimethoprim), analgesics (diclofenac, ibuprofen, ketoprofen, paracetamol, naproxen), psychiatric drugs (carbamazepine, primidone), antihyperlipidermis, stimulants (caffeine), lipid regulators (atenolol, bezafibrate, gemfibrozil, metoprolol propranolol), etc. In addition to these drugs, PPCPs also include synthetic nitropolycyclic fragrances, antimicrobial compounds like triclosan, UV blockers like methylbenzylidene camphor, antioxidants, preservatives like phenols and parabens, hormones (estrone E1, estradiol E2, ethynlestradiol EE2), insect repellents (N,N-diethyl-m-toluamide (DEET)), etc. The physiochemical properties of PPCPs show that they are moderately polar, soluble in water, nonvolatile, thermally unstable, and exist in a wide range of pH (acidic to basic). Except for caffeine and hormones, most of the PPCPs are manmade or produced by human activities [105].

Bu et al. reported a screening level risk assessment (SLERA) regarding the aquatic environment in China. Among the 112 PPCPs detected, 6 important PPCPs in surface waters were listed as erythromycin, roxithromycin, diclofenac, ibuprofen, salicylic acid, and sulfamethoxazole [106]. In China, an evaluation of the exposure, ecotoxicity, and environment risk assessment of 50 PPCPs from the surface waters was performed, where 15 chemicals were priority compounds out of the 50 PPCPs tested [107]. A study along the Ganges River Basin in India displayed the presence and distribution of 15 PPCPs. A long-term health risk assessment for both human and aquatic organisms from exposure to diverse PPCPs was evaluated, where the PPCPs in the Ganges River water ranged between 54.7 and 826 ng/L, where caffeine (743 ng/L) displayed the highest concentration [108].

In a similar study, George et al. investigated the presence of endocrine disruptors per- and polyfluoroalkyl substances (PFASs) in the aquatic food web and the potential human health implications. Forty (40) PFASs were evaluated from muscle biopsy and serum samples of fish representing different trophic levels along Lake Huron. Fishes (walleye, yellow perch and round gobies) collected from other contaminated rivers showed results of PFASs with bioaccumulation and biomagnification [109]. PPCPs, especially perfluoroalkyl substances (PFASs), polychlorinated biphenyls (PCBs), and polybrominated diphenyl ethers (PBDEs), were developed to maximize their activity, even at low doses, for targeting different metabolic, enzymatic, or cell-signaling mechanisms. This includes their potential to passively interfere with the endocrine system [110,111]. Despite the lower concentration, these PPCPs are capable of interfering with nontargeted aquatic life forms at the molecular, cellular, and individual levels. Long-term exposure can induce unfavorable effects like genotoxicity, carcinogenicity, and infertility [112]. As PPCPs are tolerant to biodegradation, they pass through the water cycle and reach aquatic organisms, the food chain, and eventually humans, thus regular monitoring and the evaluation of the environmental effects of PPCPs must be estimated globally.

PPCPs—Microalgal Interactions and Bioremediation

PPCPs challenge remediation efforts because of their structural diversity, ranging from polar antibiotics to hydrophobic aromatics. Microalgae address this complexity by combining biosorption, bioaccumulation, and enzymatic biodegradation with uptake strongly shaped by molecular size, polarity, and functional groups [113].

In comparison with traditional methods, algae-based systems offer an eco-friendly and more efficient PPCP removal possibility. The growth of Chlorella sorokiniana, Chlorella pyrenoidosa Chlorella vulgaris, Chlamydomonas acidophila, and Scenedesmus obliquus was unaffected in the presence of various PPCPs. However, Cymbella sp., Scenedesmus quadricauda, and Navicula sp. showed some inhibitory effects, while Haematococcus pluvialis, Selenastrum capricornutum, Scenedesmus quadricauda, and Chlorella vulgaris showed an increased growth [114]. The growth enhancement or inhibition is specific to the strain and the interaction with PPCPs. The rbcL gene corresponds to the RuBisCO large subunit, while psbA, psbB, and psbC are genes that encode for protein subunits of PS2. The genes psbA, psbB, and psbC encode the D1, CP47, and CP43 proteins, respectively. Expression of the rbcL and psbA genes decreases during the exposure of cyanobacteria to organic pollutants [115], whereas psaB and psbC gene expression reduced with long-term exposure to organic pollutants [116]. The long-term exposure of algae to high concentrations of PPCPs often resulted in a decrease in the protein and chlorophyll content because of the ROS synthesis, and alterations in the gene expression of chlorophyll (rbcL, psbA, psaB, and psbC) [114]. For example, when subjected to six different wild microalgae species (Chlamydomonas mexicana, Chlamydomonas pitschmannii, Chlorella vulgaris, Ourococcus multisporus, Micractinium resseri, and Tribonema aequale) to evaluate the response to levofloxacin (LEV) removal capacity, C. vulgaris showed the highest removal capacity. The ecotoxicological effects of LEV and the acclimation of C. vulgaris proved the simultaneous bioaccumulation and biodegradation processes [117].

Microalgae interact with PPCPs through the processes of bioadsorption, bioaccumulation, biodegradation, photodegradation, and co-metabolism [118,119]. During biosorption, soluble forms of PPCPs attach to the microalgal cell wall or its extracellular polymeric substance (EPS). These interactions can be either ionic or hydrophobic. The biosorption of PPCPs to microalgae depends upon their physiochemical features and the microalgal species. Positively charged PPCPs can easily create electrostatic interactions with the carboxyl, hydroxyl, and phosphoryl functional groups present in the microalgal cell walls [120]. Chlorella sp. was studied in the removal of nonsteroidal anti-inflammatory drugs—ketoprofen (KET) and diclofenac (DIF)—from aqueous solutions with varying parameters (time, pH, drug concentration, and algal concentration). The FTIR analysis showed strong H-bonding and n–π interactions in the biosorption process of the microalgae, with a maximum adsorption efficiency at pH 6 [121]. Haematococcus pluvialis, Selenastrum capricornutum, Scenedesmus quadricauda, and Chlorella vulgaris were investigated for analyzing the interaction against the antibiotics sulfamerazine, sulfamethoxazole, sulfamonomethoxine, trimethoprim, clarithromycin, azithromycin, roxithromycin, lomefloxacin, levofloxacin, and flumequine. From the investigation, about 10 transformation products with lower toxicity and corresponding pathways were observed. Biodegradation was the main antibiotic removal mechanism, with minor contributions of bioadsorption, bioaccumulation, and abiotic factors [122].

Nannochloris sp., was tested for 14 days of interaction with trimethoprim (TMP), sulfamethoxazole (SMX), and triclosan (TCS) in aquatic environments. Compared with the hydrophilic antibiotics TMP and SMX, a 100% removal was observed for the lipophilic antimicrobial agent TCS from the medium. These algae promoted photolysis and showed the efficiency of Nanochloropsis-mediated TCS removal in water bodies [123]. Ciprofloxacin and sulfadiazine removal by Chlamydomonas sp. Tai-03 demonstrated the elimination of PPCPs by biodegradation and photolysis [124]. Similarly, for the removal of cephalexin (CEP) and erythromycin (ERY) from a wastewater treatment plant, a microalgae-bacteria consortium made of Chlorella sorokiniana and Brevundimonas basaltis was used, where more than 90% of chemical removal was observed [125]. Chlamydomonas reinhardtii, Chlorella sorokiniana, Dunaliella tertiolecta, and Pseudokirchneriella subcapitata were tested to remove nine antibiotics—azithromycin, clarithromycin, erythromycin, ciprofloxacin, ofloxacin, norfloxacin, trimethroprim, sulfapyridine, and pipemidic acid—and the antidepressant venlafaxine. From the analysis, the C. reinhardtii and D. tertiolecta cultures displayed a high growth profile and remarkable antibiotic removal. Ciprofloxacin, norfloxacin, ofloxacin, and pipemidic acid showed good photodegradation, while the photodegradation competence of azithromycin, clarithromycin, and erythromycin was found to be less than 5%. Trimethoprim and venlafaxine remained stable in the solution without much removal; sulfapyridine was eliminated mainly by algae biodegradation. The transformed products from azithromycin, erythromycin, and sulfapyridine were also subjected to biodegradation with algae [126]. In another study, Chlorella vulgaris and Scenedesmus obliquus were used to evaluate the interactions with sulfamethazine and enrofloxacin. The degradation and metabolic pathways of the drugs were evaluated in the tested microalgae, where the presence of sulfamethazine and enrofloxacin did not inhibit the growth. Enrofloxacin degradation involved dealkylation, decarboxylation, and defluorination. Chlorella showed a higher degradation potential toward both sulfamethazine and enrofloxacin. From the analysis, 40–50% of drug removal was observed due to a combination of biodegradation and photolysis [127]. To remove a mixture of three antibiotics—tetracycline, ciprofloxacin and sulfamethoxazole,—a mixed culture combining two cyanobacteria (Microcystis aeruginosa and Synechocystis sp.) and two microalgal species (Raphidocelis subcapitata and Tetradesmus obliquus) were used, where specific stimulated growth along with photosynthetic activity was observed in the cyanobacteria without inhibiting the microalgal growth. In M. aeruginosa, an antibiotic mixture enhanced the growth by gene regulation associated with ribosomes, photosynthesis, and nutrient metabolism. This resulted in the dominance of M. aeruginosa in the mixed culture under antibiotic exposure. Due to the selective promotion of cyanobacterial competitiveness against eukaryotic microalgae, a cyanobacteria bloom occurred in response to the antibiotic exposure [128].

Oxybenzone (benzophenone-3), the UV filter, is an active ingredient in many pharmaceutical products. The impacts of oxybenzone on the growth and molecular composition of different microalgae were investigated in Chlorella UMACC 400, Chlorella UMACC 401, Chlorella sp., Chlamydomonas reinhardtii, and Scenedesmus quadricauda. Upon exposure to various concentrations of oxybenzone, the highest concentrations were found to have adverse effects on the growth rate and biomass of these microalgae. However, exposure to oxybenzone at higher concentrations (200 mg/L to 400 mg/L) did not influence S. quadricauda growth, and higher concentrations of oxybenzone resulted in cell structure changes and increased protein and carbohydrate content [129]. Similarly, veterinary antibiotic removal by microalgae is through the degradation of tetracycline, oxytetracycline, chlortetracycline, and doxycycline in swine wastewater treated with phycoremediation. The carbohydrate-rich phycoremediation biomass is a rich source for bioenergy conversion [130].

There are several consistent structure-property patterns in the microalgal uptake of PPCPs. Microalgal interactions are influenced by molecular properties including hydrophobicity, molecular weight, and the presence of functional groups. Compounds with a higher hydrophobicity or halogenated aromatics (e.g.,: triclosan) exhibit greater algal uptake via hydrophobic partitioning and π-π interactions, with cell surface proteins or extra cellular polymeric substances. However, hydrophilic molecules adhere weakly, which can be observed among various algae and mixed consortia systems, enabling biosorption on aromatic surfaces as hydrophobic domains on algal cell walls [44,131,132]. Ionization state (pKa) and functional groups regulate the electrostatic interactions. Amine bearing protonated PPCPs like fluoroquinolones and macrolides interact strongly with anionic algal surfaces through electrostatic interactions and cation bridging, where anionic interactions are regulated by the hydrophobic moieties of algae [44]. Certain functional groups enhance covalent binding through aromatic rings and heterocyclic favored π-π stacking, and phenols and amides connect through hydrogen bonds to extracellular polymeric polysaccharides and proteins utilizing the algal surface amine, phosphate, carboxyl groups as complementary sites [133]. Steric structural features of PPCPs (e.g., carbamazepine), prevent easy algal uptake and biodegradation unless supported by immobilization or engineered biocatalysts like cytochrome P450 [134,135,136]. Immobilized algal consortia increase the contact time and oxidative capacity, enabling the removal of anionic, hydrophobic PPCPs like diclofenac [46]. PPCPs with cationic or amine functionality and higher aromatic hydrophobicity are easily captured by microalgae, whereas neutral, small, polar, and rigid structured PPCPs require additional modifications or process intensification for algal uptake [47]. These patterns suggest the integration or modification of different physiochemical characteristics of PPCPs at the molecular level for improved algal bioremediation.

Table 2. Microalgal responses to PPCPs.

Pollutant Name (PPCPs)	Microalgae	Observed Activity and Removal Efficiency	Analysis Performed	References
Diclofenac	Chlorella vulgaris, Nannochloropsis oculata, Scenedesmus acutus, and Scenedesmus obliquus	Biosorption and bioremediation.	Growth kinetics, chlorophyll evaluation, removal efficiency comparison analysis.	[137]
Sulfonamides, fluoroquinolones	Chlorella vulgaris, cyanobacterium (Chrysosporum ovalisporum)	Biosorption/bioaccumulation.	Growth inhibition assay, growth kinetics, pigment fluorescence, antioxidant enzyme evaluation, LPO evaluation residual antibiotics.	[138]
Ibuprofen, salicylic acid, acetaminophen, diclofenac, tetracycline	Coelastrum sp.	Pharmaceutical removal (mechanism not mentioned). Ibuprofen, salicylic acid, and acetaminophen—89.4–99.8% removal, diclofenac 55%, and tetracycline 100%.	Evaluation of removal efficiency, conventional treatment line efficiency.	[139]
EE2—ethinylestradiol, E2—estradiol, ibuprofen, estrone, gemfibrozil, bisphenol A	Chlorella vulgaris, Scenedesmus sp., Westella botryoides, and diatoms species	Photodegradation, bioadsorption and biodegradation. EE2—25.12%, E2—84.91%, ibuprofen 64.8%, estrone 95%, gemfibrozil 39%, and bisphenol A 43%.	Sewage quality analyses, GC-MS analysis, behavior and fate of micropollutants.	[140]
Ibuprofen, diclofenac	Parachlorella kessleri	Bioaccumulation and adsorption, growth inhibition, photosynthetic imbalance, and chlorophyll variation. Ibuprofen 51.3% and diclofenac 55.7%.	Growth kinetics, chlorophyll content, adsorption rate/removal efficiency, photolysis analysis.	[141]
Salicylic acid, ibuprofen	Scenedesmus Obliquus	Biosorption. Adsorptive removal. Maximum adsorption capacities for salicylic acid 60 mg/g and ibuprofen 12 mg/g.	Adsorption kinetics, thermodynamic adsorption, FTIR, point of zero charge determination, SEM.	[142]
Sulfamethazine, sulfamethoxazole	Scenedesmus obliquus	Biosorption, bioaccumulation, and biodegradation as an adaptive mechanism to antibiotics. sulfamethazine 31.4–62.3%, sulfamethoxazole 27.7–46.8%.	Growth kinetics, analysis of biochemical content, elemental analysis, FTIR, HPLC.	[143]
Estradiol, Diclofenac, Triclosan	Microalgal consortium—Chlorella sp., Merismopedia sp., Closteriopsis sp., and Scenedesmus sp.	Biodegradation. Growth inhibition and variations in chlorophyll content. Estradiol 91.73%, diclofenac 74.68%, and triclosan 78.47%.	Growth kinetics, chlorophyll analysis IC50 assays, drug adsorption HPLC, SEM, and degradation studies.	[144]
Cefradine	Chlorella sp. L166 and Scenedesmus quadricauda	Biodegradation. Three degradation pathways—decarboxylation, hydroxylation, demethylation. 97.27% by Chlorella sp. and 98.50% by Scenedesmus quadricauda.	Growth kinetics, pigment changes, antioxidant enzymes evaluation, removal efficiency, metabolic identification.	[145]
Ofloxacin	Scenedesmus obliquus	Ofloxacin stress influenced microalgae photosynthetic system, leading to carbon redistribution, increased lipid accumulation. Ofloxacin 39.24% removal by Scenedesmus obliquus.	HPLC, biomass productivity, biomolecule analysis, pigment content analysis.	[146]
Carbamazepine	Chlorella vulgaris	Biodegradation; bioaccumulation; biosorption. Removal efficiency: 79.16%.	Toxicity analysis, photosynthesis influence, antioxidant enzymes activity, removal mechanism.	[147]
Diclofenac	Picocystis sp. and Graesiella sp.	Biodegradation; biotransformation. Removal efficiency: Picocystis and Graesiella 73% and 52% of 25 mg/L initial diclofenac concentration.	Growth inhibition, photosynthetic activity, diclofenac removal and degradation intermediates, LCMS analysis.	[148]
Diclofenac	Chlamydomonas reinhardtii and Scenedesmus obliquus	Biodegradation; bioaccumulation; biosorption processes. Removal efficiency: Chlamydomonas 78%, Scenedesmus 80.1% (3 days); mixed culture: 91.4–92.3% (3 days), 100% (6 days).	Growth kinetics, diclofenac removal mechanisms (biodegradation, bioaccumulation, abiotic degradation) analysis.	[149]
Diclofenac, ibuprofen, metronidazole	Chlorella variabilis	Bioaccumulation; growth inhibition and oxidative stress.	Growth inhibition analysis, ROS production, antioxidant activity, photosynthetic activity, zeta potential analysis, hydrodynamic size measurements.	[150]

BPA—bisphenol A; FTIR—Fourier transform infrared spectroscopy; SEM—scanning electron microscopy; IC₅₀—half-maximal inhibitory concentration; LC-MS—liquid chromatography-mass spectrometry; ROS—reactive oxygen species; HPLC—high performance liquid chromatography; STA—simultaneous thermal analysis; GC-MS—gas chromatography-mass spectrometry; LPO—lipid peroxidation; UPLC-MS—ultra-performance liquid chromatography-mass spectrometry.

shows further examples of microalgal interactions with PPCPs [137,138,139,140,141,142,143,144,145,146,147,148,149,150].

At the organism level, PPCPs can accumulate depending on their physiochemical properties. Although consistent evidence for biomagnification across trophic levels remains limited, certain PPCPs with higher persistence and lipophilicity raise legitimate concerns regarding trophic transfer.

3.3. Persistent Organic Pollutants (POPs)

Persistent organic pollutants (POPs) are highly toxic synthetic organic chemicals with a particular combination of physical and chemical properties that resist photolytic, biological, and chemical degradation. Because of their hydrophobic and lipophilic semi-volatile nature, POPs will persist and remain unchanged for a long period of time in the environment (soil, water, and air). POPs accumulate in the fatty tissues of living organisms and enter into the food chain. POPs are categorized into three groups, as shown in Figure 5, as (1) pesticides such as aldrin, chlordane, deildrin, dicholrodiphenyltrichloroetane (DDT), endrin, heptachlor, hexachlorobenzene (HCBs), mirex, and toxaphene; (2) industrial chemical products like flame retardants (FRs); and (3) unintentional by-products such as polychlorinated biphenyls (PCBs), polycyclic aromatic hydrocarbons (PAHs), dioxins, and furans [151,152]. POPs are also released into the environment through volcanic activities, forest fires, agriculture runoff, industrial effluents, motorized vehicle exhaust, and other waste streams [153]. POPs cause health issues such as allergies, hypersensitivity, developmental changes, damage to the central and peripheral nervous systems, disruption of the endocrine, reproductive, and immune systems, cancers, and even death [154]. Therefore, the accumulation of POPs in the environment is a serious problem and needs urgent solutions. One of the potential solutions for this problem is microalgae, as microalgae can degrade POPs enzymatically and use them as nutrients.

POPs—Microalgal Interactions and Bioremediation

POPs persist for decades due to their lipophilic and chemically stable nature, making them especially prone to biomagnification. Microalgae mitigate these risks by assimilating POPs into lipid stores and initiating enzymatic dehalogenation, though their resistance to breakdown still remains a major challenge [75].

The main route of toxic POPs to microalgae is through the aquatic environment, and the concentration ranges from picograms to nanograms per liter. These toxic effects on microalgae vary according to microalgal species and POPs. PCBs displayed cell cycle arrest and apoptosis in dinoflagellate Prorocentrum minimum [155]. Toxicity studies of PAHs mostly measure the microalgal growth or biomass. Higher concentrations of PAHs showed declined growth and biomass, supporting the dose–effect relationship [156]. However, in oceanic Prochlorococcus populations, PAHs and organophosphorus pesticides (OPs) induced a reduction in the expression of photosynthetic genes rbc L (RuBisCO large subunit) and psb A (PSII D1 protein) [115]. Another toxicity study of an environmental mixture extracted from the Polar Organic Chemical Integrative Sampler (POCIS), using a microalgal bioassay based on the photosynthetic capacities of diatom (Nitzschia palea) cultures, showed >85% inhibition on the photosynthetic capacities. This assay showed the overall toxicity of the mixture of herbicides [157].

A similar toxicity analysis of one of the POPs—polybrominated diphenyl ether (BDE-47)—was evaluated on five species of cyanobacteria such as Synechocystis sp., Oscillatoria planctonica, Microcystis flos-aquae, Nostoc sp., and Pseudanabaena sp. Concentrations of 0.05 to 1.0 mg/L BDE-47 were used, where the growth and photosynthesis inhibition of Pseudanabaena was observed at concentrations over 0.1 mg/L. However, Synechocystis sp., Oscillatoria planctonica, Microcystis flos-aquae, and Nostoc sp. were tolerant to BDE-47, and Pseudanabaena developed a resistance to BDE-47 after 30 days of exposure [158].

Growth inhibition, photosynthetic activity inhibition, and oxidative stress are also reported by FRs. The toxicity of tetrabromobisphenol A (TBBPA), Tris(1,3-dichloro-2-propyl) phosphate (TDCPP), and tributyl phosphate (TBP) present in FRs was studied in Chlorella sorokiniana to evaluate the ecological risk of FRs to aquatic organisms. Results showed the highest toxicity by TBBPA with a 96 h EC50 value of 7.606 and photosynthetic activity inhibition. Increasing concentrations of TDCPP and TBP exposure increased the Chl fluorescence at first, and then decreased after the utilization of excess phosphorus by algal cells. TBBPA, TDCPP, and TBP enhanced ROS, malondialdehyde (MDA), and superoxide dismutase (SOD) activity, thus indicating increased oxidative stress [159].

A similar study of Chlorella sorokiniana with TBBPA and bivalent cadmium ions (Cd (II)) showed <1 mg/L of Cd(II) promoting growth, while TBBPA showed inhibitory effects. With a Cd(II) concentration of ≤1 mg/L and increasing TBBPA concentration (0–1.5 mg/L), an increase in chlorophylls and carotenoid was observed. Increasing the Cd(II) concentration further to 2 mg/L led to an increase in chlorophyll a content (4.58 mg/g) at first, and later decreased with the increase in the TBBPA concentration. The transcriptome analysis of C. sorokiniana shows an upregulation of protein synthesis-related genes in the presence of both Cd(II) and TBBPA [160]. The mechanism of toxicity toward POPs may vary in cyanobacteria and microalgae but may cause growth inhibition, oxidative damage, photosynthetic inhibition, cell cycle arrest, and apoptosis. Photosynthetic inhibition is the reduction in rate or the efficiency of photosynthesis due to stress factors like changes in light intensity/temperature, nutrient availability, and the presence of pollutants or oxidative damage leading to the suppression of photosystems. Chlorophyll parameters, such as ETo/RC, RC/ABS, ETo/TRo, Fv/Fo, and Fv/Fm, were measured once in two days using a fluorescence meter, and the photosynthetic efficiency was calculated by evaluating the concentration of photosynthetic pigment Chl a [160].

For a better understanding, further research on the toxicity and tolerance of microalgae and cyanobacteria toward POPs is needed.

Microalgal bioremediation is mediated through three pathways: bioadsorption, bio-uptake or bioaccumulation, and biodegradation [161]. In bioadsorption, microalgae act like sponges, where the POPs adhere or attach to the cell wall components or to the organic molecules secreted from the cells through different interactions such as covalent bonding, electrostatic forces, and molecular forces. In bio-uptake, however, POPs are transported by active transport, diffusion, or facilitated diffusion into the microalgal cells. Inside the cells, POPs bind to intracellular proteins and other chemicals. Catalytic metabolic degradation of the POPs by microalgae into small nonhazardous or less toxic components is biodegradation, which can happen extracellularly, intracellularly, or a combination of both [162,163]. Various microalgae genera involved in bioremediation are Arthrospira, Botryococcus, Chlamydomonas, Chlorella, Cyanothece, Desmodesmus, Nodularia, Oscillatoria, Phormidium, Scenedesmus, and Spirulina [162].

In the bioremediation process, the bioaccumulation of PCBs is influenced by the physicochemical properties, cell density, nutrient status, exudates produced by the algae, algal growth stage, membrane permeability, hydrophilicity, and chemical nature of PCBs [164].

Acetamiprid, bentazone, and propanil are polar pesticides found in fresh water. The removal efficiency by Scenedesmus sp. and Chlorella sp. in a batch lab-scale reactor showed the complete removal of propanil and acetamiprid after 7 days. Two and four transformational products, respectively, were generated in the biodegradation process. A simultaneous pesticide removal in an outdoor photobioreactor achieved 99% propanil and 71% acetamiprid removal efficiencies. Harvested biomass was used for the biogas production, and there was no inhibition for anaerobic digestion caused by pesticides. In this study, the authors coupled pesticide degradation, nutrient recovery, and biofuel production using microalgae [165]. Enzymes involved in pesticide degradation metabolism are polyphenol oxidase, superoxide dismutase, esterase, glutathione S-transferase, and catalase [164]. Multiple factors affect the bioremediation of POPs, and it is a slow process requiring from a few days to weeks for the degradation of pollutants. With the improving knowledge of cell biochemistry and molecular biology and omics studies, it is possible to achieve better degradation results.

Table 3. Microalgal responses to POPs.

Pollutant Name (POPs)	Microalgae	Observed Activity and Removal Efficiency	Analysis Performed	References
Atrazine	Chlorella sp.	Growth inhibition; Degradation created three by-products—desisopropyl-atrazine, desethyl-atrazine, and dsethyl-desisopropyl-atrazine. Removal efficiency: 64.3–83.0%.	Photocatalytic degradation evaluation, growth kinetics, atrazine removal, GCMS analysis, photosynthetic activity.	[166]
Paraoxon, malathion and diazinon	Coccomyxa subellipsoidea	Degradation; inhibition of photosystem II; mitochondrial ROS generation.	Degradation profile testing, GCMS analysis, singlet oxygen degradation assay, ROS-dependent degradation, photosynthesis efficiency.	[167]
Imidacloprid	Chlorella sp.	Degradation by hydroxylation and oxidation. Removal efficiency: 57.20–61.66%.	Cell growth kinetics. Assays (SOD, ROS and MDA), pigment evaluation.	[168]
BDE-47 or BDE-209	Chlorella sp.	Debromination; bioaccumulation; degradation; cell growth inhibition. Removal efficiency:82.2–93.5%.	Toxicity profile, bioaccumulation analysis, debromination process, extraction from microalgae and analysis by GC. Evaluation of biomolecules.	[169]
Lindane	Scenedesmus sp. ISTGA1	Bioremediation; synergistic effect on the degradation and detoxification.	Growth kinetics, evaluation of ions, GCMS analysis, detoxification analysis, cytotoxicity and EROD activity analysis by MTT and EROD assays.	[170]
Naphthalene, anthracene, benzo[a]pyrene	Chlamydomonas reinhardtii	Adsorption; degradation Removal efficiency: naphthalene 85.5%, anthracene 89.5%, and benzo[a]pyrene 16.9% by degradation, naphthalene 13.1%, anthracene 9.8%, and benzo[a]pyrene 82.0% by adsorption.	Growth kinetics, HPLC analysis, chlorophyll quantification.	[171]
Nonylphenol	Chlorella pyrenoidosa	Growth inhibition; synergistic toxicity.	Growth kinetics, chlorophyll fluorescence analysis, toxicity analysis, antioxidant assays (SOD, MDA, and CAT) assays.	[172]
PFOA and GenX	Chlorella pyrenoidosa	Growth inhibition; gene down regulation; photosynthesis variations; physical damage; metabolic disorders.	Growth kinetics, photosynthetic parameters evaluation, gene analysis.	[173]
BDE-47	Chlorella sp.	Growth inhibition; photosynthetic damage; low antioxidant level; oxidative stress; programmed cell death. EC50 (96 h)—64.7 μg/L.	Growth kinetics, photosynthetic parameters evaluation, antioxidant evaluation, TEM analysis, ROS detection.	[174]
BDE-47	Chlorella sp.	Extra cellular adsorption; intracellular absorption; bioaccumulation. Removal efficiency: 82.1–84.2%.	Adsorption and absorption kinetics, effects of EPS, light limitation, nitrogen starvation, FTIR analysis, GCMS analysis.	[175]
Bisphenol A	Chlamydomonas reinhardtii	Growth inhibition; cytotoxicity; oxidative stress. Bioaccumulation—0.16 pg BPA/cell.	Growth inhibition analysis, toxicity assays, ROS assays, FCM analysis biomarkers analysis.	[176]
PCB	Chlorella Sorokiniana	Cytotoxicity; chlorophyll degradation; growth inhibition; recovery of after stress.	Growth kinetics, fluorescent measurement, chlorophyll analysis, oxidation kinetics.	[177]
Polychlorinated diphenyl ethers	Scenedesmus obliquus	Low bioavailability; toxicity responses.	Concentration exposure studies, toxicity analysis, toxicity comparison with similar structure chemicals, oxidative stress evaluation.	[178]
Trichlorfon	Chlamydomonas reinhardtii	Biotransformation; biodegradation; toxic metabolite; chlorophyll level reduction.	Biodegradation analysis, growth kinetics, pigment analysis, antioxidant enzyme evaluation, HPLC analysis, GCMS analysis, chlorophyll fluorescence, and antioxidant enzymes.	[179]
Benz(a)anthracene	Chlamydomonas reinhardtii CC-503	Sorption and biodegradation. Upregulation of 4 genes and 5 enzymes—degradation process. Removal efficiency: 80% degradation efficiency 10 mg/L.	Growth kinetics, biomolecules and intermediate metabolites analysis by GC, gene regulation evaluation by qRT-PCR, enzyme activity analysis.	[180]
Atrazine, endosulfan	Scenedesmus arcuatus, Chlorella sp., Pseudokirchneriella subcapitata	Cytotoxicity; oxidative stress; ROS production; lipid peroxidation.	Oxidative stress biomarkers, photosynthetic biomarkers and morphological biomarkers.	[181]
Clothianidin	Chlamydomonas reinhardtii	Biosorption; biodegradation. Cytotoxicity; decrease in photochemical activity; resistance to the insecticide. Removal efficiency: 50% in 10 days.	Growth kinetics, chlorophyll analysis, microscopic evaluation, and pigment analysis.	[182]

GC–MS—gas chromatography-mass spectrometry; ROS—reactive oxygen species; SOD—superoxide dismutase; MDA—malondialdehyde; BDE-47—2,2′,4,4′-tetrabromodiphenyl ether; BDE-209—decabromodiphenyl ether; GC—gas chromatography; MTT—3-(4,5-dimethylthiazol-2-yl)-2,5-diphenyltetrazolium bromide assay; EROD—ethoxyresorufin-O-deethylase; HPLC—high-performance liquid chromatography; CAT—catalase; PFOA—perfluorooctanoic acid; TEM—transmission electron microscopy; EPS—extracellular polymeric substance; EC₅₀—median effective concentration; BPA—bisphenol A; FCM—flow cytometry; PCB—polychlorinated biphenyl; qRT-PCR—real-time quantitative reverse transcription-PCR; FTIR—Fourier transform infrared spectroscopy.

shows more examples of POP interaction with microalgae [166,167,168,169,170,171,172,173,174,175,176,177,178,179,180,181,182].

Beyond the organism level bioaccumulation, the lipophilic and recalcitrant nature of POPs raises concerns of biomagnification across food chains, thus remaining a critical but underexplored risk.

3.4. Flue Gas

Flue gases are a mixture of combustion products generated either by incineration (direct combustion) or an indirect oxidation of a combustible material or syngas [183]. Flue gas mainly contains nitrogen, moisture (water vapor), carbon dioxide, particulate matter (dust), sulfur oxide, nitrogen oxides, and carbon monoxide, as shown in Figure 6. Apart from these constituents, flue gas contains hydrogen fluoride, hydrogen chloride, hydrocarbon derivatives, and heavy metal derivatives, depending on the burning materials and combustor [184]. Power plants and cement production plants are a major source of flue gas. Other sources include industries such as lime production and use, soda ash production and use as well as chemical and metal production industries [185].

Photosynthetic microalgae are considered as a solution for the mitigation of flue gas-related emissions. The main advantages of using microalgae for flue gas mitigation include the photosynthetic utilization of CO₂, high biomass, and lipid accumulation as well as various value-added products like pigments and proteins [161]. Apart from these advantages, flue gas is considered as a free source of CO₂ for the cultivation of microalgae. One of the economic barriers for the industrialization of microalgae is the cost of sufficient CO₂ supply, where flue gas can be sustainably utilized as a carbon source [186].

Flue Gas—Microalgal Interactions and Bioremediation

Industrial flue gases contribute heavily to climate change, carrying both concentrated CO₂ and acidic co-contaminants. Microalgae transform this liability into opportunity by sequestering CO₂ through photosynthesis while stabilizing the pH and capturing other emissions in integrated bioenergy systems [187].

Depending on the source, flue gas has multiple impacts on microalgal cells. For example, in Chlorella sp. AE10, a simulated flue gas, especially SO_x and NO_x, had significant negative impacts on its PS1 and PS2, which led to changes in the photosynthetic activity of the culture. In the presence of flue gas components, the photosynthetic activity (Fv/Fm) of Chlorella sp. AE10 decreased from 0.7 to 0.2 on day 1. Even in 0.5 FG (containing 10% CO₂, 100 ppm NO_x and 50 ppm SO_x), a significant decrease in photosynthetic activity was observed from day 2 [188]. This clearly proved that Chlorella Sp. AE could not tolerate the simulated flue gas. The simulated flue gas decreased the pH of the medium from 7 to 3.3, which inhibited the growth of Chlorella sp. AE10. To overcome these problems, an adaptive evolution of 46 cycles was conducted that obtained Chlorella sp. Chlorella vulgaris, which can fix CO₂ at a rate of 1.2 g/Ld with a biomass concentration of 2.7 g/L and a 68.4% carbohydrate content. The comparative transcriptomic analysis of Chlorella vulgaris also proved that the evolved species could tolerate more flue gas [188]. The tolerance level of microalgae to CO₂ varies from 5 to 100%, according to the species and growth conditions [189]. One of the main reasons for the growth inhibition is the increasing acidity with higher concentrations of CO₂. Therefore, by maintaining the pH of the culture medium, it is possible to improve the tolerance of the microalgae. Apart from pH, the CO₂ tolerance in microalgae is also influenced by the cell density, microalgal species, light intensity, and nutrients. Furthermore, the biochemical composition and enzymatic reactions of the microalgae are also changed by CO₂ enrichment [190,191]. Similarly, the NO_x tolerance of microalgae depends on the microalgal cell density, NO_x concentration, NO_x gas flow rate, reactor type, and species [192].

The toxic effects of sulfur dioxide (SO₂) vary greatly among microalgal species. The pH and bisulfite concentrations affect the tolerance of microalgae to SO₂. Membrane transport of α-amino-¹⁴C-isobutyric acid was blocked in Ankistrodesmus sp. at pH 5 and not at pH 7, showing the influence of pH. In Botryococcus braunii, higher concentrations of bisulfites caused an inhibition in growth by means of the peroxidation of membrane lipids and bleaching of chlorophyll [193].

Through the development of efficient technologies, like metabolic engineering, screening and the selection of suitable strains, using mixed cultures, and sustainable strategies, it is possible to utilize microalgae for the biosequestration of industrial flue gas.

The strategy of developing a transgenic Chlamydomonas reinhardtii shows that the microalgal valorization of CO₂ in the industrial exhaust is a potential option. Transgenic Chlamydomonas reinhardtii cultivated in toxic flue gas (13 vol% CO₂, 20 ppm NO_x, and 32 ppm SO_x) provided double the amount of CO₂-based bioproducts compared with that of the wild-type. In the study, the gene expression patterns of wild-type Chlamydomonas reinhardtii were inspected. A low expression of plasma membrane H+-ATPases (PMAs) accompanied by intracellular acidification in response to the extremely high concentrations of CO₂ was observed. Based on these results, the wild-type strains were expressed with a universally expressible PMA. This improved the tolerance of the strain to acidic environments and increased photoautotrophic production 3.2-fold by maintaining a higher cytoplasmic pH [194]. When using direct flue gas, nitrogen and sulfur oxides in the flue gas have some toxic effects on microalgae. However, few species can remove it efficiently and use it as a building block of the cell constituent. An isolated Scenedesmus obliquus PF3 strain can accumulate biomass in 15% CO₂ or 500 ppm NO or 50 ppm SO₂. This strain can grow at different temperatures (15 to 30 °C) and pH (4.5 to 10.5) range and is tolerant up to 2 mM bisulfite. Scenedesmus obliquus PF3 can remove 96.9 ± 0.03% (2.77 ± 0.08 mg/Ld) and 87.7 ± 6.22% (1.29 ± 0.01 mg/Ld) of NO and CO₂ at a rate of 2.86 ± 0.23 mg/Ld and 1.48 ± 0.12 g/Ld, respectively [192]. In another study, 15 Chlorella strains were acclimated to higher levels of NO_x. Results of the experiments showed that a higher concentration of nitrite inhibited the growth and photosynthesis of the Chlorella strains, but some strains among the 15 strains had a higher ability to acclimate to higher amounts of nitrites. These demonstrate that the biological detoxification of NO_x is possible with the selection of suitable strains [195].

Using flue gas for the cultivation of microalgae reduces heavy metal pollution to air from flue gas. Microalgal biomass with accumulated heavy metals can be used as a feedstock for biofuel production. Mixed microalgal cultures were grown in nutrient-rich medium with Al, B, Cu, Fe, Mn, and Zn, and three concentrations of CO₂ (1%, 3%, and 5.5%) were evaluated and compared with flue gas and control systems in a photobioreactor for heavy metal accumulation. The cells internalized mainly 46.8 ± 9.45 g/L of B, 253.66 ± 40.62 g/L of Mn, and 355.5 ± 50.69 g/L of Zn during the cultivation. The possible integration of microalgal cultivation with CO₂ and heavy metal removal was demonstrated in this study [196].

For the implementation of microalgae for the bioremediation of flue gas, a life cycle assessment (LCA) of the whole process should be undertaken. An outdoor raceway pond system for biomass production and the sequestration of flue gas CO₂ under different conditions showed promising results. This batch and semi-continuous cultivation with and without flue gas supplementation showed that cultivation was the main environmental impacting step in microalgal biomass production and that the location of the thermal power plant and cultivation system should be close to each other for a better result. As per the LCA analysis, flue gas utilization by microalgae had a lower carbon footprint in semi-continuous cultivation. This system can reduce greenhouse gas emissions and other impacts by 45–50% utilizing the mass and energy requirements, operation efficiency, emissions, and performance [197].

The bioremediation of flue gas by microalgae needs to be intelligently applied and practiced for better results. The integration of wastewater valorization, industrial facilities, and microalgal production is a promising method. Apart from this, metabolic engineering and cost-effective downstream processing are also needed for economically viable sustainable process development.

Table 4. Microalgal responses to flue gas components.

Pollutant Name (Flue Gas Elements)	Microalgae	Observed Activity and Removal Efficiency	Analysis Performed	References
CO2, SO2, and NO	Chlorella pyrenoidosa	Active CO2 biofixation. SO2 neutralization and NO detoxification. Improved growth and lipid production. Removal efficiency: CO2—95.9%, SO2—100% and NO—84.2%.	Growth kinetics, chlorophyll evaluation, lipid/biomolecule evaluation by GCMS and efficiency calculation.	[198]
CO2	Chlorella pyrenoidosa Chick (IPPAS C2)	Biofixation of CO2, improved biomass production. CO2 bio-fixation 0.790 g/Ld (10% CO2).	Growth evaluation, CO2 sequestration, dry weight analysis.	[199]
SO2	Chlorella sp.	Spermidine induced SO2 detoxification, cell protection, enhanced antioxidant activity, improved photosynthesis, and improved biomolecule production.	Growth analysis, chlorophyll fluorescence analysis, maximum photochemical efficiency evaluation, antioxidant assays, biomolecule e valuation by GC, effect of spermidine in presence of SO2.	[200]
NO	Chlorella sp.	Spermidine-induced NO assimilation: antioxidant enzymes prevented peroxidation damage, proteomic/metabolomic adjustments improved photosynthesis.	Growth kinetics, spermidine influence analysis, photosynthetic pigment concentration, enzyme assays, proteomic analysis, metabolomics analysis.	[201]
CO2, NO3	Chlorella vulgaris	Bioaccumulation, fixation of CO2, nitrate removal and biomass improvement.	Growth kinetics, photosynthetic activity analysis and carbonic anhydrase activity.	[202]
CO2, NOx, SOx	Chlorella sp. AE10; Chlorella sp. Cv	CO2 fixation and tolerance evaluation of CO2, NOx, SOx.	Evaluation of growth kinetics, tolerance levels of CO2, NOx, SOx and photosynthetic activity.	[203]
CO2, NO	Asterarcys quadricellulare; Chlorella sorokiniana	Bio-fixation of CO2 and NO.	Non-photochemical quenching analysis, photosynthetic oxygen evolution measurement, fluorescence measurements.	[204]
CO2, SO2, NO	Chlorella sp.	Reduced biomass yield and the content of carbohydrate, down regulation of 3 enzymes, growth inhibition.	Growth kinetics, flue gas effect on biomass and biomolecules, gene expressions evaluation.	[205]
CO2	Chlorella vulgaris	Biofixation. 5% CO2 showed best growth capability with enhanced biomass. Removal efficiency: 76.92% CO2. CO2 fixation rate: 0.318 ± 0.009 (g/Ld) for 5% (v/v) CO2 input.	Growth kinetics, morphology analysis, functional group analysis, CO2 removal efficiency, CO2 fixation rate, biodiesel properties.	[206]
CO2	Scenedesmus sp. (UKM9); Chlorella sp. (UKM2)	Fixation of carbon, nitrogen and phosphate. Chlorella sp. recovered CO2 12.435 g/L after 15 days of CO2 fixation.	Growth rate analysis, biomass productivity evaluation, biochemical evaluation, CO2 fixation efficiency.	[207]
CO2, NO	Chlorella vulgaris	Carbon fixation. Improved biomass and enhanced biomolecule content CO2 fixation rate 47.51 mg/Ld.	Growth kinetics, biomass productivity, fixation of CO2 and nitrogen, biomolecule analysis with HPLC.	[208]
CO2	Botryococcus braunii; Scenedesmus sp.	Carbon fixation. Optimum CO2 removal efficiency 10% for B. braunii and 20% for Scenedesmus sp.	Growth evaluation, CO2 sequestration analysis, biomolecule evaluation.	[209]
CO2	Chlamydomonas reinhardtii	Carbon fixation. Maximum CO2 sequestration observed at 30% CO2.	Growth analysis, pigment evaluation, CO2 sequestration analysis.	[210]
CO2	Scenedesmus obliquus	Carbon fixation. Improved protein kinase and ATPase activity and oxidative phosphorylation process.	Growth analysis, oxidative stress response evaluation, enzyme activity evaluation and proteomics analysis.	[211]
CO2, N2, O2	Chlorella sp.	Carbon fixation. CO2 fixation rate 0.90 g/L·day. Improved lipids, proteins, and carbohydrates 20.95%, 26.48%, and 9.3%, respectively.	Growth kinetics, biomass productivity and biochemical composition evaluations.	[212]

GC-MS—gas chromatography-mass spectrometry; SEM—scanning electron microscopy; FTIR—Fourier transform infrared spectroscopy.

shows the flue gas interactions with microalgae [198,199,200,201,202,203,204,205,206,207,208,209,210,211,212], with an emphasis on the reported efficiencies of SO₂ and NO_x removal, alongside CO₂ biofixation. This evidence suggests that flue gas grown algal biomass may carry lower toxicological burdens, since SO_x and NO_x can be transformed rather than accumulated, enabling downstream conversion into biofuels, bioplastics, or biochemical feedstocks with minimal contamination risks.

Although flue gas-derived contaminants can bioaccumulate in lower trophic organisms, their potential for biomagnification across food chains remains a critical concern.

3.5. Microplastics and Nanoplastics

Microplastics and nanoplastics are small pieces of synthetic polymers found in the environment as emerging particulate anthropogenic pollutants. They are commonly found in aquatic organisms, water, soil, sediments, foods, and even in the air. Micro- and nanoplastic particles have extremely intricate and different sizes, morphology, composition, density, surface properties, etc. [213]. Polyethylene (PE), polyethylene terephthalate (PET), polypropylene (PP), polystyrene (PS), and polyvinyl chloride (PVC) are the most common plastics found in the environment [214]. Other relevant plastics produced in bulk are polybutylene terephthalate (PBT), polycarbonate (PC), polystyrene (PS), polyurethane (PUR), polymethylmethacrylate (PMMA), polytetrafluoroethylene (PTFE), acrylonitrile butadiene styrene (ABS), etc. [215].

Depending on the size, various classifications are present in the case of plastic contaminants. Large debris of plastics with a size > 25 mm is called macroplastics, mesoplastics are 25–5 mm, and microplastics (MPs) have a size limit up to 5 mm, whereas plastics with a particle size < 100 nm (sometimes < 1 μm) are defined as nanoplastics. Micro/nanoplastics (MNPs) can be of primary and secondary origin. Primary MPs include plastic beads used in exfoliants, cosmetic products, specific purpose nanomaterials (3D printing, tissue engineering, drug delivery), industrial abrasives, accidental spills, etc. Secondary plastics are the inadvertent products of plastic degradation due to environmental weathering processes (e.g., hydrolysis, UV photodegradation, mechanical abrasion, and biodegradation). Secondary MPs exist in irregular shapes and in complex chemical forms [213,216].

In an aquatic ecosystem, MNPs interfere with nutrient efficiency and cycling. It causes physiological stress in organisms, which can lead to behavioral alterations, immune responses, abnormal metabolism, and changes to energy budgets, thus creating an imbalance in the ecosystem. This can lead to the defective development of species, populations, and the creation of nonfunctioning communities. The occurrence of MNPs in wetlands, mangroves, and deep ocean seafloor ecosystems has a higher risk of deteriorating susceptible aquatic environments [217].

Micro/Nanoplastics—Microalgal Interactions and Bioremediation

Micro- and nanoplastics are long lasting contaminants that stress macroalgae through oxidative damage and light interception. Microalgae enable the aggregation of these particles through extracellular polymers, absorbing them on the cell surfaces and forming biofloccules that reduce their mobility [218,219].

MNPs influence aquatic systems at different trophic levels. The major influence from the cell to the community level is the inhibitory effect on growth, followed by other effects such as photosynthesis depression, enzymatic alterations, ROS production-induced oxidative stress effects, structural damage, etc. The toxicity impacts depend on the microalgal cell size, microplastic features (type, surface charge, concentration, and time of exposure), microalgal species, and synergistic toxicity effects with other contaminants [215]. To evaluate the effect of polystyrene microplastics (PS-MPs) on microalgae, Xiao et al. studied a model of the freshwater microalgae Euglena gracilis exposed to 1 mg/L PS-MPs (size—5 μm and 0.1 μm) for 24 h. From the results, a reduction in pigment value and SOD activity was observed, corresponding to the oxidative stress. Pigment value refers to the quantitative content of photosynthetic pigments—chlorophylls, carotenoids, and sometimes phycobiliproteins—present in algal biomass, expressed in units of dry weight per cell volume, and is an indicator of the light harvesting capacity and physiological status of the algae [94]. The molecular level evaluation showed that gene alteration was observed for the cellular activities (cell, gene, organism level) and metabolisms. Moreover, activation of the KCS gene and CTR1 gene was observed at the 5 μm PS-MPs exposure, showing the pathways to influence further adverse effects on the microalgae [220]. In a similar investigation, the effect of microplastic PVC (mPVC, average diameter 1 μm) vs. plastic debris (average diameter 1 mm) was compared on the marine microalgae Skeletonema costatum. For a 96 h exposure, an obvious inhibition of up to 39.7% was observed with microplastic PVC, while no inhibition was observed for plastic debris. A higher concentration of 50 mg/L mPVC showed an inhibitory effect. The interaction between the mPVC and microalgal cell was observed by SEM analysis, which showed adsorption and aggregation. The shading effect of the mPVC on the cells by absorption is another factor. This interaction limits the transfer of nutrients and signals, thus limiting the growth of the microalgae [221].

Song et al. investigated the interaction between four microplastics (PP, PE, PET, and PVC) and two microalgae (Chlorella sp. L38 and Phaeodactylum tricornutum MASCC-0025). From the evaluation, microplastics displayed an inhibition ratio of up to 21.1% on the growth of the marine algae P. tricornutum. However, Chlorella sp. showed a strong adaptive capacity toward microplastics. To evaluate the possible toxic effect of microplastics and associated reactive oxygen species generation, antioxidant enzyme superoxide dismutase (SOD) and catalase (CAT) assays were performed. The growth enhancement could also be due to the leaching mechanism of microplastics. Analysis of the active enzymes’ concentration variation and morphological characterization by SEM and TEM confirmed the toxicity effects on both microalgae and also showed the EPS production, which could lead to adsorption and aggregation with microplastics as a biosolution for microplastic treatment via bioelimination [222].

Anabaena sp. PCC7120, a filamentous cyanobacterium, was utilized to compare the toxicity of polystyrene nanoplastics (PS-NPs) and a combination of PS-NPs with 7th generation PAMAM dendrimers (G7). The interaction with PS-NPs prompted the induction of ROS and associated cellular activities (lipid peroxidation, membrane disruptions, intracellular acidification), displaying a negative influence on photosynthetic activity and cell internalization whereas PS-NPs combined G7 interaction on anabaena displayed a lower toxicity and no internalization, possibly due to the aggregate formation. Additionally, the evaluations showed the presence of different protein molecular biomarkers corresponding to the nanoplastic interaction [223].

Another study by Khoshnamvand et al. assessed the effects of amino-functionalized polystyrene nanoplastics (PS–NH2) on Chlorella vulgaris. For the three sizes of PS-NH2 (with diameters of 90, 200, and 300 nm) in C. vulgaris, a dose-dependent toxicity was observed for the PS-NH2 sizes of 90 and 200; end-points of biomass and photosynthetic pigment (chlorophyll a) were also observed. However, different concentrations of PS-NH2-300 had no significant toxic effect on the biomass and chlorophyll-a end-points was observed. Smaller-sized nanoplastics showed a higher toxicity. With decreasing size, higher morphological changes, less algal cell density, and cell aggregates were observed. The electrostatic interaction between C. vulgaris and PS-NH2, or the potential of PS-NH2 to behave as a mediator for connecting algal cells could be the reason for the complex formation. From the morphological analysis, PS-NH2 aggregates found to be adsorbed on the microalgal cells prevented the nutrient, energy, and gas transfer, thus causing toxicity [224]. Similarly, polystyrene nanoparticles (PSNPs) with different surface charges (negative and positive charge) and different sizes (50, 300, and 500 nm) were investigated on Chlamydomonas reinhardtii to evaluate the hetero-aggregation, cellular internalization, and cellular responses. From the studies for different concentrations, 300 and 500 nm sized PSNPs caused an apparent toxicity to C. reinhardtii by the PSNP-microalgae hetero-aggregation and shading effect. The smaller, 50 nm sized PSNPs showed a dose and surface charge-dependent toxicity. Photosynthesis damage-induced toxicity, which is an impairment of the photosynthetic apparatus, specifically in photosystems and the electron transport chain, was observed due to the direct physical or physiochemical interactions between the algal cells and nanoparticles. The positively charged PS-NH₂ were hetero-aggregated with microalgae C. reinhardtii more than the negatively charged PS-COOH, leading to the shading of light harvesting pigments and disruption of thylakoid membrane function, thus reducing photosynthetic efficiency. This reduced photosynthetic capacity led to impaired growth and viability. Also, EPS production was observed in the case of higher concentrations of 50 nm sized PSNPs by the cell surface adsorption and cell internalization [225].

To evaluate the continuous and cumulative exposure of microplastics to microalgae, polystyrene (PS) nanoplastics of 80 nm were introduced to Chlorella pyrenoidosa. From the results, C. pyrenoidosa exhibited a defense mechanism of self-recovery against the toxicity impacts and the inhibitory effect due to accumulation [226].

In a similar study, 80 nm polystyrene nanoplastics were tested on C. pyrenoidosa to evaluate the biochemical and molecular level responses [227]. Upon the analysis, both inhibitory and stimulating effects were detected for the chronic exposure of nanoplastics on C. pyrenoidosa. The initial inhibitory response, observed for 2 weeks, gradually decreased, and an enhanced algal growth was detected during days 15–21. From the molecular analysis, a gene downregulation of aminoacyl-tRNA synthetase was observed, which showed the possibility of growth inhibition. Later, C. pyrenoidosa developed a defense mechanism against polystyrene-induced stress by endorsing cell proliferation and speeding up the degradation of defective proteins, which resulted in the enhanced growth [227]. Exposure to low concentrations (1–10 mg/L) of polystyrene nanoplastics activated cellular deference responses in C. pyrenoidosa, which are characterized by the upregulation of genes involved in DNA replication and cell cycle progression, thereby promoting cell proliferation. The observed accelerated defective protein degradation reflects the enhanced proteostasis through increased intracellular protein quality control, regulating damaged proteins. These coordinated responses enable the maintenance of cellular homeostasis by removing nonfunctional components, regulating metabolic efficiency under nanoplastic-induced stress conditions. This adaptive mechanism leads to a stimulatory effect on algal growth over prolonged exposure to nanoplastics.

To evaluate the response of Scenedesmus obliquus toward EPS-treated functionalized PSNPs (plain, aminated, and carboxylated), an investigation was conducted by Giri et al. From the results, compared with the pristine forms, the aging caused lesser toxic effects on the microalgae and induced agglomeration. Depending on the surface functionalization and aging duration, the ROS levels were reduced, the oxidative stress markers increased, and cell viability improved with a higher photosynthetic efficiency [228].

Similarly, to evaluate the bioremediation potential of microalgae toward the nano- and microplastics, an EPS producing freshwater algae Cyanothece sp. was used. For two different concentrations (1 and 10 mg/L) of the polystyrene nano- and microplastics tested, an inhibitory effect of the microalgal growth was observed, along with EPS production as a defense mechanism. However, an excellent bioflocculant activity was observed for the EPS produced by Cyanothece sp., even in low concentrations. Moreover, EPS showed enhancing characteristics for the aggregation, where the complex was formed with microalgae, EPS, and the micro- and nanoplastics, which ensures the excellent bioremediation potential for microplastic pollution markers [229].

The majority of phycoremediation studies on MNPs involve physical removal by adsorption and hetero-aggregation processes [230], and few studies are listed on the biodegradation of micro-nanoplastics by enzymatic degradation [91]. Though the phycoremediation of MNPs includes both physical and chemical processes, the majority of studies have referred to the physical bioremediation process rather than actual biodegradation, which is quite limited and species dependent. Most microalgal species capture MNPs through cell surface interactions, hetero aggregation, extracellular polymer binding, and physical removal. However, a recent study by Gowthami et al. investigated six marine microalgal strains—Amphora sp., Dunaliella salina, Limnospira indica, Navicula sp., Picochlorum maculatum, and Synechocystis sp.—for their polystyrene biodegradation capability [91]. Over a 45-day incubation period, polystyrene microplastics underwent enzymatic degradation by laccase enzyme, as confirmed via weight reduction, structural-physiochemical characterizations, and molecular docking analysis. Microalgae facilitated this biodegradation process by forming biofilms on microplastic surfaces, causing structural alterations, and promoting laccase-driven enzymatic breakdown. Molecular docking provided insights into the specific individual interactions between the laccase enzyme and styrene ligand. Further studies employing molecular docking could further explore the potential of other microalgal enzymes for the phycoremediation of nano- and microplastics.

Table 5. Microalgal responses to micro- and nanoplastics.

Pollutant Name (Micro- and Nanoplastics)	Microalgae	Observed Activity and Removal Efficiency	Analysis Performed	References
Polyvinyl chloride	Chlorella vulgaris	Adsorption; aggregation; precipitation in cells. Dose dependent growth inhibition; high oxidative stress; improved enzymatic activity.	Toxicity analysis, growth kinetics, SEM, biomass production, antioxidant enzyme analysis.	[231]
Polystyrene, dibutyl phthalate	Chlorella pyrenoidosa	Growth inhibition; cytotoxicity. Synergistic toxicity effect lead to morphological variations and total pigment content changes. IC50 value (96 h) of dibutyl phthalate was 2.41 mg/L and for polystyrene 6.90–7.19 mg/L.	Growth inhibition analysis, morphology, cytotoxicity analysis.	[232]
Polystyrene, poly(methyl methacrylate), polylactide	Scenedesmus abundans	Adsorption; hetero-aggregation Removal efficiency > 84%.	Growth kinetics, flow cytometry, removal efficiency calculations, SEM, EPS quantification.	[233]
Polyethylene, polyethylene terephthalate, polypropylene, polystyrene	Scenedesmus vacuolatus	Growth inhibition, oxidative stress, toxicity.	Growth inhibition study, microplastic leachates/degradation products effects, cytotoxicity evaluation.	[234]
Virgin polyvinyl chloride, UV-aged polyvinyl chloride	Chlamydomonas reinhardtii	Aged polyvinyl chloride more toxic; growth inhibition; reduction in chlorophyll-a content’ oxidative stress; antioxidant enzyme activation.	Growth kinetics, chlorophyll content evaluation, antioxidant enzyme activity assays, FTIR.	[235]
Polystyrene (nano- and microsizes)	Chlamydomonas reinhardtii	Nanoplastics showed higher adsorption capacity and lower desorption rate to lead ions. Growth inhibition, oxidative stress and cytotoxicity.	Adsorption/desorption tests, particle charge evaluation, toxicity analysis, biochemical analysis, elemental analysis and morphological analysis.	[236]
Polystyrene (1, 5 µm)	Chlamydomonas reinhardtii	Growth inhibition; cytotoxicity; antioxidant enzyme activity; chlorophyll a content decrease; hetero-aggregation of microplastics.	Growth measurement, antioxidant enzyme analysis, toxicity analysis, Pigment quantification and photosynthetic activity measurement, SEM.	[237]
Polystyrene microplastics	Chlorella vulgaris	Leaching depends on time, pH, water sources. High microplastic concentration lead to inhibitory effects.	Growth kinetics, chlorophyll fluorescence, cell distribution analysis.	[238]
Polystyrene microplastics (6 μm)—plain polystyrene, aminated polystyrene, carboxylated polystyrene	Chlorella sp.	Toxicity; growth inhibition; oxidative stress; hetero aggregation.	Growth analysis, toxicity evaluation, SEM, oxidative stress determination.	[239]
Polyethylene, polypropylene, polystyrene, polyvinyl chloride, polyethylene terephthalate (200–600 µm)	Scenedesmus sp.	Growth inhibition; Smaller size particles induced higher inhibition.	Microplastics preparation, Characterization, growth kinetics, growth inhibition analysis.	[240]
Polystyrene	Scenedesmus quadricauda	Cell wall thickening; internalization; aggregation. Highest concentration 200 mg/L—enhanced growth, improved levels of chlorophyll, polysaccharide and soluble proteins.	Growth kinetics, chlorophyll content analysis antioxidant enzyme activity assays, biomolecule quantifications.	[241]
Amino-modified polystyrene nanoplastics (50 nm)	Microcystis aeruginosa FACHB 905 (single cells), FACHB 1327 (small colonies), and FACHB 1338 (large colonies)	Enhanced microcystin production, inhibition of photosystem II efficiency, oxidative stress, gene expression, protein up regulation and decreased cell membrane integrity.	Algal growth evaluation, chlorophyll a content analysis, evaluation of cell membrane integrity, antioxidant enzyme activity biomolecules, evaluation of gene expression and proteomic response.	[242]
Polystyrene microbeads (100 nm, 100 μm)	Chlamydomonas reinhardtii	Nanosized plastics showed higher growth inhibition via shading effect, oxidative stress and cell damage. Higher antioxidant enzyme activity and cell toxicity also observed.	Algal cultivation evaluation, chlorophyll content analysis, evaluation of biomolecules, cell morphology analysis by flow cytometry and SEM analysis, antioxidant enzyme activity and lipid peroxidation assays.	[243]
Polystyrene (20, 50 and 500 nm)	Chlorella vulgaris	Growth inhibition; ROS generation; antioxidant enzyme activity; membrane damage; stress conditions.	Growth kinetics, growth inhibition analysis, chlorophyll a content analysis, LDH assay, FTIR, TEM/SEM, ROS associated oxidative stress determination.	[244]
Polystyrene nanoplastics (200 nm)	Chlorella sp.	Cytotoxicity; cell viability reduced; loss of membrane integrity; reduced photosynthetic activity; enhanced ROS level; oxidative stress; EPS formation.	Growth kinetics, growth inhibition by cell viability, elevated intracellular ROS and specific radicals (hydroxyl and superoxide), membrane integrity analysis, percentage of maximum quantum yield of PSII evaluation.	[245]
polyvinyl chloride (≤85 µm)	Chlorella vulgaris	Growth inhibition; biofilm inhibition; chlorophyll reduction; EPS production.	Growth kinetics, biofilm formation analysis, FTIR, FESEM, chlorophyll concentration analysis, EPS analysis.	[246]

SEM—scanning electron microscopy; DBP—dibutyl phthalate; IC₅₀—half-maximal inhibitory concentration; PS—polystyrene; PMMA—poly(methyl methacrylate); PLA—polylactide; EPS—extracellular polymeric substance; PE—polyethylene; PET—polyethylene terephthalate; PP—polypropylene; FTIR—Fourier transform infrared spectroscopy; Pb²⁺—lead ion; μm—micrometer; nm—nanometer; PSII—photosystem II; FESEM—field emission scanning electron microscopy; ROS—reactive oxygen species; LDH—lactate dehydrogenase.

includes the interaction of microalgae with nano- and microplastics [226,227,228,229,230,231,232,233,234,235,236,237,238,239,240,241].

Overall, the toxicological behavior of MNPs is largely governed by factors such as particle size, surface charge, polymer type, and environmental aging. These factors modulate their bioavailability and interaction with the organism. Current evidence indicates that physical processes, including adsorption and hetero-aggregation, constitute the predominant mechanism, whereas biodegradation remains limited and less explored. These insights show the need for more systematic studies to explain the long-term transformation dynamics and develop removal strategies under realistic conditions.

As per current evidence though, MNPs can be transferred across trophic levels, strong evidence about biomagnification is lacking, and the possibility remains as an emerging ecological concern.

4. Biomagnification and Transfer of Toxins in Food Chain

Microalgae are widely recognized as natural bioremediators that are capable of sequestering, transforming, and degrading pollutants. Also, as primary producers, they are the first trophic step in food webs, making them potential gateways for contaminants to higher organisms. This duality raises serious concerns about the hidden risks of trophic transfer and biomagnification. In aquatic ecosystems, various scientific and regulatory programs are employed to assess chemical hazards. Bioconcentration factors (BCFs), bioaccumulation factors (BAFs), biomagnification factors (BMFs), and trophic magnification factors (TMFs) are used as pollutant quantification tools to measure the bioaccumulation hazard, risk of exposure, and trophic transfer possibilities in ecosystems [247]. Evidence from recent trophic migration factor (TMF) studies has shown some leads, but the evidence remains fragmented across various pollutant classes [248,249,250,251,252].

Microalgae are widely used for food and feed production due to the high protein and omega-3 fatty acid contents, which offer a suitable aquaculture feed option [253,254,255]. Aquaculture requires huge volumes of protein rich feed material to improve fish health, and the nutritional quality influences fish health and market value. High quality aqua feed is an area that is still expanding, where microalgae offer a reliable and less volatile solution [256]. However, this promising solution carries a hidden risk of the bioaccumulation of pollutants, especially when microalgae are grown in wastewater biorefineries or industrial side streams, due to their excellent bioremediation potential, which retains the pollutants in the biomass [161]. Although pollutant transfer investigations from algae to higher trophic levels are limited, ecotoxicology evidence indicates the biomagnification possibilities [257].

Heavy metals efficiently accumulate in microalgal cell wall ligands, extracellular polymers, and intercellular sequestration. Once ingested, they could propagate to higher trophic levels. Reports have confirmed the trophic transfer patters in aquatic environments with biomagnification and the trophic transfer of Pb (TMF = 1.56 and Cd (TMF = 1.31), whereas Cu (TMF = 0.64) and Cr (TMF = 0.73) often dilute depending on the speciation and physiology. The highest BMF of 3.89 was observed in the case of Pb [258].

The trophic transfer of titanium dioxide nanoparticles (TiO₂ NPs)—a major component of PPCPs—from the microalga Nitzschia closterium to scallop Chlamys farreri was observed with biomagnification in the gills, digestive gland, and mantle of scallops. Along with the transfer of TiO₂ NPs, toxicity impacts in terms of lysosomal membrane damage, DNA damage, and histopathological effects were also observed [259]. The toxicological impact of TiO₂ NPs in intensified results were also observed in higher trophic levels [260].

A neurotoxin β-N-methylamino-L-alanine (BMAA) was biomagnified to advanced trophic levels in a diatom-dominated marine ecosystem. The presence of BMAA was detected in zooplankton, bivalve mollusks, carnivorous crustaceans, and carnivorous gastropod mollusks with a TMF of 4.58, 30.1, 42.5, and 74.4, respectively. Among the 56 tested diatom strains, 21 strains contained BMAA (concentration 0.11 to 3.95 µg/g dry weights) [261]. Ciprofloxacin exhibited trophic magnification in cold water webs (TMF = 2.8), with strong electrostatic binding onto the algal surface or extracellular polymers [262].

The presence of silver nanoparticles (AgNPs) in the food chain was through the microalga Dunaliella salina to brine shrimp Artemia salina and to guppy fish Poecilia reticulate. The accumulation, distribution, concentration, and biomagnification factors and trophic toxicity of AgNPs were evaluated in the trophic food chains. Overall, silver accumulation was amplified in guppy fish with exposure concentration and reducing trophic levels. Bioconcentration factor values for microalgae, shrimp, and fish were recorded as 826, 131, and ≈1000, respectively. Additionally, the trophic exposure of AgNPs also induced oxidative stress and reproductive toxicity [263].

POPs, specifically PFAS, biomagnify protein affinity, making protein-rich algal biomass an entry pathway with consistent trophic increases, where short chain analogues dilute more. In a comprehensive analysis of 24 PFASs tested in the coastal food web (not on low trophic levels) of the U.S. North Atlantic in 18 marine species, 19 PFASs were detected across the samples, indicating that the concentration of the transferred PFAS depends on the species, habitat, body size, feeding guild, and location [264].

Micro- and nanoplastics readily adsorb to algal surfaces and aggregate via extracellular polymers, enabling the ingestion by zooplankton and transfer to fish. Beyond plastic transfer, the co-transfer of additives heightens the risk [260].

The uses of algae-derived biomass depend on the cultivation source and contaminant presence. Biomass grown in clean or fully treated waters are routed for food, feed, nutraceutical applications. Strict regulatory monitoring adhering to the standards is provided, since microalgae supplies high-quality protein and omega-3 fatty acids for aquaculture and human diets [18,265,266] whereas the microalgae grown in municipal or industrial wastewater is usually directed to biofuels, biochar, or fertilizer due to the trophic transfer risk of pollutants like heavy metals, PPCPs, POPs, micro-nanoplastics, and cyanotoxins, which can accumulate in algal cells [267,268,269,270]. Fuel pathways, particularly hydrothermal liquefaction (HTL), offer a low risk outlet as most metals partition into solid parts, although the incomplete destruction of some pollutants has raised concerns requiring downstream controls. Biofuels and fertilizers are preferred routes of wastewater grown biomass, while feed and food applications require contaminant free growth conditions and strict safety validation, underscoring the importance of end user risk assessment in microalgal bioremediation strategies.

Despite these insights, systemic evidence for biomagnification from algae to higher trophic levels remains limited, and more detailed, multi-trophic, long-term studies are needed to better understand the ecological and human health implications.

5. Challenges and Future Prospects

Extensive research is required for a better understanding of pollutant toxicity and algal bioremediation. The interdependent and intelligent integration of these two studies can solve many challenges caused by these pollutants. The bioremediation process faces numerous environmental challenges such as chemical complexity, formation of new complex molecules, pH, salinity, temperature, oxygen and nutrient availability, microbial dynamics, product toxicity, etc. Therefore, developing a universal solution for this complex issue is not an easy process [271]. Another challenge is the generation of toxic or harmful by-products and bioaccumulation after bioremediation, which can disrupt ecosystems and cause selective algal blooms [272]. Apart from these scientific challenges, industrial deployment also faces substantial techno-economic difficulties. Large-scale algal systems are constrained by light penetration, self-shading, seasonal variability, and the instability of culture/contaminations, while harvesting and dewatering remain energy intensive and cost dominant steps. Recent techno-economic analysis has proven that large-scale bioharvesting operations, like flocculation–dissolved air flotation, can lower costs compared with flocculation sedimentation, however, the overall treatment costs remain high [273]. Moreover, recent trends on microalgal biorefinery value chains underline the trade-offs between CAPEX and OPEX to realistically assess the commercial viability [274]. Techno-economic analyses consistently show that the financial viability of algal wastewater treatment depends on reducing the downstream processing costs and coupling treatment with the generation of value-added bioproducts. While considering these challenges, they are highly interconnected and can lead to more complex scenarios.

To resolve these issues, the integration of different technologies and scientific methodologies (physical, chemical and biological) has to be used in a sustainable way. Advances in metabolic engineering and omics can provide more insights into the underlying processes and enhance organism level understanding [275] Advances in metabolic engineering and omics—a collective term for high throughput biological data evaluations including genomics (DNA), transcriptomics (RNA), proteomics (proteins), and metabolomics(metabolites)—are providing deeper insights into microalgal physiology and pollutant removal mechanisms. Multi-omics evaluations, when applied in bioremediation strategies, have resulted in improved selected gene expressions, protein production, and metabolite generation in response to the contaminants and reactor conditions by identifying and focusing on targeted enzymatic pathways for biosorption, biodegradation, and oxidative stress management [275,276]. Additionally, new gene editing tools (RNA interference (RNAi), transcription activator-like effector nucleases (TALENs), zinc-finger nucleases (ZFNs)) and targeted metabolic engineering using CRISPR/Cas9 mediated gene edits are universal toolkits for broader species adaptation.

Phycoremediation of various pollutants is facilitated by the biodegradation action of key enzymes. The main enzymes involved in pollutant degradation by microalgae are laccase, peroxidase, cytochrome P450 monooxygenase, dehydrogenase, carboxylase, esterase, and glutathione S transferase. The efficiency and activity spectrum of these enzymes depend on multiple factors like the microalgal species, pollutant structure, and environmental conditions [92]. The biotransformation of heavy metals and metalloids is supported by various reductase enzymes like arsenate reductase, mercuric reductase, and chromate reductases. Chlorella vulgaris utilizes chromate reductase for the reduction of highly toxic chromium, Cr(IV) to less toxic Cr(III). Similarly, Chlorella fusca, Galdiera sulphuraria, and Selenastrum minutum mediate the biotransformation of Hg²⁺ into elemental mercury and metacinnabar (HgS) via the mercuric reductase enzyme, and Chlamydomonas reinhardtii utilizes arsenate reductase for the reduction of arsenate As(V) to arsenite As(III) [277]. Polyphenol oxidase and laccase from Chlamydomonas moewusii are effective in the biodegradation of phenolic compounds that come under PPCPs [278]. Aphanocapsa spp., Chlorella spp., and Scenedesmus spp. are capable of biodegrading various synthetic organic dyes that come under POPs by azo reductase enzymes [279]. Similarly, six marine microalgal strains—Amphora sp., Dunaliella salina, Limnospira indica, Navicula sp., Picochlorum maculatum, and Synechocystis sp.—successfully degraded polystyrene by the laccase enzyme [280]. Molino et al. investigated the expression and secretion of PHL7, an enzyme by Chlamydomonas reinhardtii that breaks down polyethylene terephthalate (PET) plastics [281]. They engineered C. reinhardtii for recombinant enzyme PHL7 production and tested the efficacy in degrading ester bond containing plastics such as PET and polyurethanes. The enzymatic degradation generates PET monomers, which can be recycled to new PET plastics.

These genetic engineering strategies could be tailored to target specific metabolic pathways for enhanced pollutant uptake, improved biomolecule production, higher tolerance, and resilience to regulate the phycoremediation performance [282,283].

In microalgal bioremediation, a significant gap remains specifically in understanding trophic transfer and the fate of bioaccumulated toxins and biodegradation pathways. Most of the studies that focused on biosorption neglected the biodegradation outcomes and the trophic transfer through food chains. Multidimensional research, including human-related risk assessments, is required to evaluate the trophic transfer of pollutants to higher organisms. The storage and maintenance of live cultures preserving high biomass and nutrient yields on a commercial scale is an expensive task, with the challenge of reduced cell viability, quality, and nutrient losses over time. The development of economically feasible biorefineries is important for sustainable microalgae cultivation. Compared with traditional agriculture, microalgal cultivation is still a developing area, which needs the transference of technology and well-established farming equipment facilities for high-yield production, depending on the intended final product. For the production of high-value biomolecules or biofuels, the cost associated will include the complex culturing setup, specific equipment, and energy consumption for the downstream processes [284]. Economic feasibility and life cycle assessments must remain central to evaluating algal bioremediation strategies, ensuring that environmental benefits are not outweighed by prohibitive costs or energy demands, as recent reviews highlight critical trade-offs between capital expenditure (CAPEX), downstream processing energy, and ecosystem impact across the value chain [274].

The integration of multiple bioremediation organisms, such as microalgae-bacteria consortia or mixed algal co-culture techniques, could be utilized for bioremediation/biodegradation in the event of mixed contaminant environments. The synergistic effect will be more efficient in limiting the pollutants through an accelerated biodegradation mechanism [285]. Microalgae are considered as a promising solution for sustainable development, turning waste streams into valuable products. The integration of microalgae with industries in a rational way will lead to a positive carbon footprint and the zero emission goal being attained. In addition, the generation of oxygen and nutrients by microalgae facilitates biodegradation by other microbial communities. However, concerns about biodegradation and the bioaccumulation of toxic components remain unanswered. Thus, further research should be carried out to clearly understand the biodegradation, bioaccumulation mechanism, and the fate of the final products. Integrating fields like nanotechnology, gene editing, and strain improvement with microalgal bioprocessing can improve the bioremediation potential of microalgae. A combination of new and existing technologies in a sustainable framework will support the effective valorization of various industrial by-products, and thus circular economy goals.

6. Conclusions

Unregulated industrial activity has generated a myriad of contaminants of diverse nature, placing unprecedented stress on ecosystems and human well-being. Although various mitigation strategies have been developed, microalgae stand out as one of the most promising solutions due to their remarkable capacity to sequester pollutants and simultaneously generate value-added products. Multiple species have already been employed to remediate contaminants in air, water, and soil, highlighting the feasibility of algae-based interventions across diverse environmental matrices.

However, there is a critical paradox—while the efficiency of microalgae in pollutant removal is well-established, uncertainties remain regarding the metabolic transformations of the absorbed contaminants and the risks posed by their eventual entry into the food web through bioaccumulation or biomagnification. The limited characterization of metabolic end-products limits our ability to fully evaluate the ecological and human health implications. To ensure safe and responsible applications, future research must integrate systematic ecotoxicological assessments, metabolic pathway analysis, and life cycle evaluations into bioremediation frameworks.

Addressing these challenges will require the deployment of advanced approaches ranging from nanotechnology and gene editing to strain improvement and systems biology tools that can improve or optimize the contaminant degradation while minimizing the unintended risks. Importantly, certain applications, such as flue gas-based cultivation for CO₂ capture, offer pathways to generate algal biomass with comparatively lower toxicological burdens, making such biomass suitable for conversion into biofuels, bioplastics, biofertilizers, and other products without much concern or contaminant transfer.

A future vision must focus on algal bioremediation within the broader framework of a circular bioeconomy, where industrial waste streams are not only remediated, but also valorized into sustainable bioenergy and high-value products. Ultimately, the future of algae-based technologies will depend on balancing their extraordinary remediation potential with ecological safety. When these challenges are met, microalgae can evolve into a pillar of sustainable resource management, advancing both environmental restoration and global decarbonization.