Chapter 1: Advanced Chemical Extraction and Detection Methods Applied in the Photochemical Treatment Methods for Removal of CECs from Different Water Matrices

The study of the environmental impact of Contaminants of Emerging Concern (CECs) has drawn attention in recent years, due to the increase in their production and use of chemicals worldwide.¹  CECs are the group of naturally occurring or synthetic chemicals that enter into the environment and can cause adverse effects on human health and aquatic ecosystems.²  It comprises a group of contaminants such as pharmaceuticals and personal-care products (PPCPs), food additives, pesticides, industrial chemicals, engineered nanomaterials, flame retardants, endocrine-disrupting compounds (EDCs), and perifluorinated compounds. The prevalence of these CECs in the environment is relatively low in concentration, ranging from µg L⁻¹ to ng L⁻¹, and their detection poses challenges due to sample clean up followed by analytical methods development.^3,4  CECs may enter surface water bodies due to discharge of treated sewage/wastewaters into rivers and lakes. Wastewater treatment plants (WWTPs), sewage treatment plants (STPs) or effluent treatment plants (ETPs) consist of secondary treatments, such as the activated-sludge process, and tertiary treatments, such as filtration and disinfection, which inadequately remove CECs means there is an urgent need for advanced treatment technologies for their remediation.⁵  The other sources of CECs are nonpoint sources of pollution such as agriculture runoff, household and hospital waste as well as landfill leachates (Figure 1.1).

A wide range of advanced treatment methodologies are being studied for the elimination of CECs including adsorption based on activated carbon,⁶  membrane technologies,⁷  and advanced oxidation process (AOPs).⁸  AOPs are considered as one of the most promising and ecofriendly of the studied technologies due to generation of highly reactive nonselective radicals, which can partially or completely eliminate the organic pollutant from the water matrices.⁹  AOPs include photocatalytic, chemical oxidation, and photochemical processes, wherein photochemical treatment is widely preferred.¹⁰  The photochemical treatment generates reactive hydroxyl radicals (˙OH) in the presence of solar radiation/UV lamps and oxidants such as hydrogen peroxide/persulfate/catalyst, making the process more sustainable for future purpose.¹¹ 

The CECs, particularly PPCPs and their metabolites present in the environment undergo partial degradation through various physical, chemical, and biological process and form new chemical entities called transformed products (TPs). The TPs formed during AOPs possess properties that differ from the parent compound and their existence can pose hazards, which makes it essential to perform risk assessment of the CECs in the water matrices.¹²  Some of the examples of TPs are 4-isobutylacetophenone (4-IBAP) from ibuprofen that are toxic even at the concentration of 1 mM to erythrocytes and cultured fibroblast.¹³  The TP of carbamezapine and oxcarbazepine is acridine, which exhibits carcinogenic potential,¹⁴  while the phototransformed products of diclofenac are highly toxic approximately five times more than the parental compound to the green algae.¹⁵ 

The identification of the transformation pathways of the contaminants, through identifying the TPs, is used to evaluate their fate and also to assess the risk of the CECs in the water matrices, which is of the utmost importance. One of the most challenging and complex tasks is the detection of TPs using sophisticated analytical instruments such as liquid chromatography-mass spectrometry (LC-MS), gas chromatography-mass spectrometry (GC-MS), nuclear magnetic resonance (NMR), etc.¹⁶  Due to recent advancement in technology and development in the domain of LC-MS, it has been considered as one of the most popular techniques for the identification of the TPs. The LC-MS detection does not require chemical derivatization as in GC-MS, enabling detection of polar TPs from the complex matrices.¹⁷  The coupling of LC with high-resolution mass-spectroscopic analyzers such as quadrupole time-of-flight (QToF), quadrupole-linear ion trap (QqLIT), time-of-flight (ToF), quadrupole ion trap (IT), Orbitrap, are some of the most recently used instruments for the structural elucidation of TPs of the CECs. These high-resolution instruments characterize TPs and detect the elemental formulae, on the basis of the exact masses without any potential interference of the compounds with identical masses in the complex matrices.¹⁸ 

The chapter primarily focuses on the reliability of analytical techniques in measurement of ubiquitous CECs and their TPs formed during the AOP, and the recovery capabilities of popular analytical extraction techniques. The focus is on the use of the high-resolution MS-based techniques in combination with approaches to overcome matrix effects among coeluting components, and the role of solvents in detection of CECs in LC coupled with various mass analyzers.

Photochemical treatment is a significant mitigation pathway for numerous organic contaminants in water. It entails chemical reactions in the presence of a photocatalyst/oxidizing reagents under the influence of light (a photon (hv) is an elementary particle of light).¹⁹  Photochemical methods such as AOPs are effective because they involve the formation of the reactive oxygen species (ROS), like hydroxyl radicals (˙OH), superoxide anion radicals (˙O₂⁻), ions, etc. using solar/ultraviolet (UV) radiation.²⁰  The ROS is capable of reacting with organic compounds to form carbon dioxide, mineral acids, and water as a byproduct, however, complete mineralization of the organic pollutant is often not achieved.²¹ 

The photochemical process has been classified into (i) direct (with UV), (ii) indirect photochemical processes (with UV/O₃, UV/H₂O₂, UV/H₂O₂/O₃, photo-Fenton, and UV/Cl) and photocatalysis with TiO₂ or any other semiconductor. Direct and indirect photochemical processes are also known as homogeneous photochemical treatment, whereas photocatalysis with TiO₂ or any other semiconductor is termed a heterogeneous photochemical process (Figure 1.2).²² 

The process in which the oxidizing reagent and the reactant are present in the same phase is known as homogeneous photochemical process. It is driven by the exposure of the UV light and the compounds that absorb the UV radiation at lower wavelength are the best examples of photodegradation.²³  The free radicals are generated from the interaction of the homogeneous molecule of the oxidizing agent such as ozone and hydrogen peroxide. The most common homogeneous photochemical processes are UV/H₂O₂, ozonation (UV/O₃), UV/H₂O₂/O₃, UV/Cl and photo-Fenton (Fe²⁺; Fe²⁺/H₂O₂).²⁴ 

UV radiation is often used for the eradication of CECs across the various water matrices. The UV alone can remove the photosensitive compounds from river, drinking waters, etc., with low concentration of the contaminants. However, the water matrices with higher concentration of contaminants such as effluents of treatment plants can hinder the process and require alternate methodology for the remediation of the organic contaminant.²⁵  UV/H₂O₂ is an AOP used for the oxidation and effective destruction of the hazardous organic matter from complex matrices. The H₂O₂ photolyzed into two (˙OH) radical under the UV radiation as shown in eqn (1.1)–(1.6), then it reacts with the organic contaminant to undergo decomposition and it absorbs in the range of 200–300 nm, i.e., low absorption wavelength:

The H₂O₂ molecules in the reaction act as scavengers in order to consume hydroxyl radicals and generate water and oxygen molecules as shown in eqn (1.7). As a result, the demand for H₂O₂ must be strong in order to generate a sufficient amount of ˙OH for the breakdown and mineralization of organic target pollutants.²⁶ 

For UV radiation, mostly two lamps have been used commercially i.e., low-pressure (LP) and medium-pressure (MP) mercury vapor lamps. The UV radiation from the LP lamps emit in the narrow region at around 254 nm, while MP emits the radiation in the broad region from 200 to 800 nm, respectively.²⁷  The LP lamps generate (˙OH) radicals more effectively as compared to MP, while it was found that the photodegradation of the organic contaminant is more effective as well as superior using MP as compared to LP.²⁸  The studies proved that as the concentration of H₂O₂ increases, the (˙OH) radical generation increases leading to an increase in the rate of decomposition. However, the excessive use of H₂O₂ can enable the reaction between the formed (˙OH) radical and the oxidant results into the formation of the radical, which could be less reactive than the originally formed (˙OH) radical.²⁹  This oxidation process is widely used for the removal of endocrine-disrupting compounds (EDCs) such as estrone, 17-α ethinylestradiol (EE2), 17-β estradiol, bisphenol-A, polycyclic aromatic hydrocarbon (PAHs), pesticides, pharmaceuticals, etc.³⁰ 

Ozonation (UV/O₃) is another viable oxidation process using ozone as a powerful oxidant in the combination of UV radiation for degradation of the organic contaminant. Ozone alone is not able to generate sufficient (˙OH) radicals for the degradation or mineralization of the contaminant.³¹  Ozone is frequently used for the treatment of municipal and industrial treatment, drinking-water disinfection, water and wastewater treatments, chemical synthesis, agriculture, etc. The dissolved ozone on photolysis (λ ≈ 260 nm) first produces H₂O₂ molecules, then the H₂O₂ reacts with the ozone or UV radiation to generate (˙OH) radical as expressed in eqn (1.8)–(1.10).³² 

The major advantage of ozonation is the short half-life of ozone of approximately 10 min, which leads to rapid degradation of the organic contaminant. It also produces large amounts of (˙OH) radicals as compared to UV/H₂O₂, and depends on the type of UV lamp used for the treatment. Additionally, the combination of UV and ozone inhibits the production of bromate.³³  The UV/O₃ degrades most of the organic contaminant such as amoxicillin, azithromycin, norfloxacin, ciprofloxacin, sulfamethoxazole, and trimethoprim.^32,33  However, due to the high energy requirement for the production of both ozone and UV light, the economic viability of this oxidation process is limited.

In the UV/H₂O₂/O₃ ternary process, there are several different processes, such as UV photolysis of either O₃ or H₂O₂, direct photolysis, ozonation or the combined impact of H₂O₂ and O₃ that can lead to the production of (˙OH) radicals that have been studied.³⁴  The most essential feature of this system is the existence of both UV radiation and H₂O₂ that accelerates the ozone decomposition in order to yield increased (˙OH) radical generation,³⁵  as shown in eqn (1.11). Some of the antibiotics removed using this ternary system are chlortetracycline, penicillin, trimethoprim, ciprofloxacin, etc.³⁶  Nevertheless, the cost of the three elements of the system (i.e., O₃, UV radiation and H₂O₂) poses a major drawback and thus limits the wide application of the oxidation process. As a result, the ternary process is limited to highly contaminated effluents in order to achieve degradation or mineralization of the persistent pollutant.²⁵ 

The photo-Fenton process is the combination of Fe²⁺ or Fe³⁺ and H₂O₂, i.e., the Fenton reagent using UV radiation for the degradation and mineralization of the organic pollutant such as aliphatic and aromatic compounds.³⁷  The photoreduction of Fe³⁺ ions to Fe²⁺ ions, which in turn react with H₂O₂, generates (˙OH) radicals. The process will depend on the pH of the system, which should be acidic (pH ∼3), and studies have shown the efficient removal of PPCPs, specifically antibiotics from the water matrices. The presence of UV radiation enhances the generation of (˙OH) radicals and the consumption of H₂O₂ reduces compared to the traditional Fenton process, as shown in eqn (1.12) and (1.13).³⁸  This is one of the most promising approaches for the removal of some of the antibiotics such as trimethoprim, ciprofloxacin, oxacillin, chloramphenicol, tetracycline, etc. from the water and wastewater matrices.^37–39  The ferrous ion formed during the reaction will react with H₂O₂ to generate ferric ions and hydroxyl radicals and the process continues.

The UV/chlorine method generates active radical species such as ˙OH, ClO˙, Cl˙, and Cl₂˙⁻. Amongst the active radicals formed in this treatment, Cl˙ is widely used as an oxidant due to its high oxidation potential of 2.47 V for remediation.⁴²  It is highly reactive with the ECs comprising electron-rich anilines, phenolic, aromatic, secondary and tertiary amine moieties.⁴³  The pH of the system as well as the oxidant used for the treatment can have an adverse effect on the degradation of the pollutant. An acidic pH, less than 6.5, is suitable for the treatment, while an increase in the amount of chlorine enhances the generation of (˙OH) radicals as well as degradation of the organic pollutant, as shown in eqn (1.14)–(1.16).⁴⁴ 

The heterogeneous photochemical process includes photocatalysis with TiO₂ or any semiconductor. Photocatalysis is the phenomenon of accelerating the rate of photochemical reaction using a semiconductor or metal oxide that acts as a photocatalyst in the presence of solar or UV radiation.^35,36  The heterogeneous photochemical process comprises of four steps for the degradation. The first step is the generation of a negative electron–positive hole pairs (e⁻ and h⁺), by the absorption of UV radiation by the semiconductor or the photocatalyst.⁴⁷  The second and third steps include separation and migration of photogenerated charge carrier and formation of (˙OH) and (˙O₂⁻) radicals by redox reaction, and the fourth and last step is the photodecomposition of the organic pollutant by means of reaction with a reactive active species on the surface of the photocatalyst.⁴⁸  The mechanism involved in the process is shown in Figure 1.3.

There are several photocatalysts used for the remediation of the environmental organic pollutant such as TiO₂, WO₃, ZnO, Nb₂O₅, Fe₂O₃, ZrO₂, and V₂O₅, etc. and they also were evaluated.^39,40  TiO₂ is a widely used photocatalyst due to its high photocatalytic activity, chemical stability, nontoxic, low cost, and it can easily become activated by UV radiation.⁴⁹  The metal-oxide-based photocatalyst possesses a large energy band gap, which ultimately requires UV light for activation. The electronic band gap can be modified and the photocatalytic performance could be enhanced by doping the metal oxide with phosphate in Bi₂WO₆, Mo/TiO₂, N-doped TiO₂/ZnFe₂O₄ hybrid that can efficiently degrade antibiotics as well as phenol in water matrices.^41,42  The key advantage of this phenomenon is that it can work at room temperature and under pressure and the photocatalyst used is cost effective and environmentally friendly (i.e., it is recyclable and irradiating the catalyst can be done by using solar radiation). This mechanism is also capable of accomplishing either partial or complete mineralization of the organic pollutant in the water matrices into carbon dioxide and water molecules. Meanwhile, it has considerable drawbacks such as difficulty of achieving homogeneous radiation across the entire catalyst surface, and recovery of the photocatalyst requires additional extraction technique, which increases the cost of the process.⁵⁰ 

The presence of CECs in the various water matrices such as groundwater, surface water, and drinking water has been reported.^45,46  Several AOP methods have been used for the remediation as discussed above, such as UV/H₂O₂, UV/H₂O₂/O₃ ozonation (UV/O₃), photo-Fenton (Fe²⁺; Fe²⁺/H₂O₂), UV/chlorine, heterogeneous photocatalysis, etc., whereas the selection of methods depends on the targeted organic contaminant in the various water matrices. The comparative removal efficiency of various treatment process along with their TPs is shown in Table 1.1. This section describes some of the studies published in the literature during the years 2009 to 2020.

The UV/H₂O₂ process is studied for removal of carbamazepine, gemfibrozil, ibuprofen, naproxen, and diclofenac from the wastewater matrices⁵²  wherein only triclosan is completely removed. The degradation of carbamazepine by UV/Cl forms 10,11-epoxy carbamazepine and acridine substituted compounds as TPs of carbamazepine.⁵³  The photocatalytic degradation of carbamazepine using UV/modified TiO₂ is reported to be up to 78%.⁵⁴  The TPs of Ibuprofen by UV/H₂O₂ are 4-isobutylacetophenone (4-IBAP), a highly toxic transformed product that affects the red blood cells, the central nervous system as well as connective tissue cells of the human body.⁵⁵  It was reported that the concentration of the TPs is sometimes much more than the parent compound in the effluent of the wastewater.⁵⁶  Some of the PPCPs such as sulfamethoxazole, trimethoprim, ofloxacin are reported to be completely removed using the UV/H₂O₂ process. The removal of caffeine and indomethacin from water matrices is reported to be complete by the UV/H₂O₂ process.⁵⁷ 

Studies on remediation of CECs using UV/Cl have been widely reported by researchers. The degradation of caffeine, carbamazepine was evaluated at different pH, i.e., 7.5 and 8.5. It was found that carbamazepine was effectively removed by the combination of UV and Cl rather than using UV or Cl alone. The Cl˙ radical contributes to the degradation of the carbamazepine, while no Cl added or substituted TP was reported and also shown in Figure 1.4. The ˙OH radical reacts with carbamazepine on the N-heterocyclic ring and generates hydroxylates and epoxide, which is further oxidized by ˙OH and Cl˙ radicals forming di-hydroxyl carbamazepine and intermediates through hydrogen abstraction. The amine or acrylamido cleavage results in the formation of acridine-9-caraboxaldehyde that was later hydroxylated. The acridone or (9-OH-acridine) was formed either by the carboxylation or ketonization of the formed oxidation products, whereas de-ketonization of acridone and de-carboxylation of (9-OH-acridine) forms acridine as shown in Table 1.1.⁵⁷ 

The oxidation reaction of ibuprofen by UV/Cl and UV/H₂O₂ follows a pseudofirst-order reaction kinetics wherein the rate constant of UV/Cl was 3.3 times higher than UV/H₂O₂ due to a large amount of radicals, hence enhancing the process of remediation at pH 6.^57,58  However, in yet another study with UV/H₂O₂ and UV/Cl, ˙OH radicals contributed to 75% degradation of Ibuprofen as compared to 22% with Cl˙ radical at pH 6. On increasing the pH from 6 to 9 the efficiency of degradation by ˙OH radical decreases to 53% and Cl˙ radical increases from 30 to 17%.⁵⁹  It has been reported that as compared to UV/H₂O₂, UV/Cl efficiently removes sulfamethoxazole, trimethoprim, 17-α ethinylestradiol, diclofenac, etc.^59–61  The results obtained by UV/Cl were compared with UV/H₂O₂ and it was found that the H₂O₂ is not totally consumed and subsequently requires another process for removal, making the cost of UV/H₂O₂ more than that of UV/Cl.^52,53 

The photo-Fenton method (Fe²⁺/Fe²⁺/H₂O₂) eliminates some of the organic contaminants such as ciprofloxacin, oxacillin, chloramphenicol, amoxicillin, ampicillin, and trimethoprim with complete mineralization or partial mineralization, which forms TPs.²⁵  The chemical reactions involved in the degradation of the ciprofloxacin by the photo-Fenton method and UV/H₂O₂ take place at the piperazinyl ring and forms a degraded product, as shown in Figure 1.5. The oxidation of the quinolone moiety results in defluorination and a hydroxyl-substitution reaction, wherein the fluorine atom is replaced by a hydroxyl group.⁶²  The oxidation of the cyclopropyl group led to ring cleavage and forms 2 oxidized products, i.e., the two dimeric products. These dimeric products are the result of C–C covalent bond formation and involve a series of transformations such as breaking of piperazinyl group, hydroxylation, and removal of the cyclopropyl group.⁶³ 

The UV/O₃ AOP process with PPCPs such as caffeine, mefenamic acid, amoxicillin, azithromycin, sulfamethoxazole, ketoprofen can result in formation of TPs. The degradation of the ketoprofen by UV/O₃ forms major and minor transformed products with their m/z ratio as shown in Figure 1.6. The degradation begins with decarboxylation of the ketoprofen molecule forming 3-ethylbenzophenone, which undergoes hydroxylation at the benzyl position, forming 3-(1-hydroxyethyl)benzophenone. Double hydroxylation of the aromatic ring of 3-ethylbenzophenone or mono hydroxylation at 3-(1-hydroxyethyl) benzophenone, forms 3-(1-hydroxyethyl) hydroxy benzophenone. The cleavage of the hydroxyethyl group and oxidation of the formed intermediates led to the removal of the acetyl group and oxidative ring opening of the parent molecule forming 2-[3-(1-hydroxy-2-methylpropyl)phenyl]propanoic acid.⁶⁴  The formed TPs along with their m/z are shown in Table 1.1. UV/O₃ effectively produces degradations of 95% of ketoprofen and its TPs within the exposure of an hour. The 3-(1-hydroperoxyethyl)benzophenone is completely removed after an hour of UV/O₃ treatment, wherein the remaining aromatic TPs were still detected in the range of 10 to 30% after the treatment.⁶⁵  However, due to the large amount of energy required for ozone generation and UV lamps, the UV/O₃ has not been used for wide-scale application.

The heterogeneous AOP process that includes a semiconductor photocatalyst was applied for degradation of caffeine, ibuprofen, and sulfamethoxazole. It was found that the degradation of caffeine was not completely achieved, some of the oxidation product after the treatment was detected.⁶⁶  The photocatalytic removal of ibuprofen shows rapid mineralization on TiO₂ with the formation of small oligomeric species in the form of intermediates during the reaction. The photocatalytic property of the semiconductor can be increased by either doping or compositing (coupling of one or more semiconductors). Doping of TiO₂ with Fe, Bi, Ni, S, Ag, and Au has been carried out. Reduced graphene oxide has been used for the removal of carbamazepine, ciprofloxacin, naproxen, sulfamethoxazole, and ibuprofen efficiently. The doping of Fe into TiO₂ reduces the band gap of the pure TiO₂ from 3.20 eV to 2.40 eV, which improves the photocatalytic activity by increasing the charge-carrier trapping sites of the semiconductor.⁶⁷ 

When PPCPs enter into an environment or water matrices, a set of various chemical reactions and biological processes takes place. PPCPs are present in the environment in the form of parent compounds and their metabolites or TPs.⁷⁷  At the time of the transformation, the PPCPs undergo various reactions such as decarboxylation (–CO₂), dehydration (–H₂O), reductive displacement of chlorine (–Cl, +H), loss of nitro group (–NO₂ + H), oxidative displacement of chlorine (–Cl +OH), epoxidation (+O), glucuronide conjugation (+C₆H₈O₆), N-acetylcysteine conjugation (+C₆ H₈NO₃S), and acetylation (+C₂H₂O), as shown in the earlier section.⁷⁸  The detection of the formed TPs in the environmental sample is important as these chemicals can be more toxic than the parent compounds.

The structural elucidation of TPs requires sophisticated analytical instrumental techniques, like nuclear magnetic resonance (NMR), liquid chromatography-mass spectrometry (LC-MS) and gas chromatography-mass spectrometry (GC-MS), etc.⁷⁹  For a few decades, the use of MS detectors has become more prominent due to advancement in the analytical development and thus elucidation of the structure of TPs. LC-MS coupled with atmospheric pressure ionization (API) interfaces, electrospray ionization (ESI) or atmospheric pressure chemical ionization (APCI) has been used for the analysis of the polar organic compounds. The detection of the compounds with low molecular mass is widely performed by LC coupled online with ESI-MS.⁷⁷ 

High-resolution-mass spectrometry (HRMS) has become a powerful tool for qualitative as well as quantitative analysis of the TPs. The most widely used HRMS analyzers are time-of-flight (ToF), quadrupole ion trap (IT), quadrupole time-of-flight (QToF), quadrupole linear-ion trap (QqLIT), Fourier-transform ion cyclotron resonance (FT-ICR-MS), and the Orbitrap.⁸⁰  Amongst the analyzers, the Orbitrap is the widely preferred analyzer for the structural analysis of the TPs.⁸¹  It has been reported that the HRMS such as TOF, QTIOF, etc. provides detailed characterization of the TPs, along with their elemental formulae and exact mass of the compound in the complex matrices comprising potentially interfering contaminants with similar apparent masses.⁸²  The characteristics of MS analyzers used for the identification of TPs are shown in Table 1.2.

Overall, toxicological data on the impacts of target TPs on ecosystems is lacking, in particular, few systematic investigations have produced records on their influence on the environment. As a result, more study in this area is necessary. Analytical methods for TPs have been divided into three classes: target quantitative analysis with reference standards, suspect screening without references, and nontarget screening.⁶⁹ 

• Target quantitative analysis with reference standards: the analysis requires a reference standard to estimate the concentration of the TPs in the sample. The reference standard could be an isotope labeled as an internal standard, for each target whose TPs is to be analyzed. This technique ensures unequivocal confirmation of a TPs carried out with low-resolution tandem MS systems or full scan-based methodology such as HRMS instruments. Identification of TPs involves a balance between the target analysis and screening methods.⁸³ 

• Suspect screening without references: specific information on substances expected to be in the samples should be accessible for tentative identification in the second technique. Suspect screening determines the availability of comprehensive datasets with retention periods and precise HR mass spectra of both the molecule and fragment ions associated with TPs. The exact mass of their predicted ions, derived from the chemical formula, can be used to screen suspects. For further confirmation, additional evidence such as isotope-pattern match is frequently used and the characteristic mass defect is compared with compounds for the analysis.⁸⁴ 

• Nontarget screening: this nontarget screening method starts with no prior data of the underlying chemicals. As a result, a complete identification of the nontarget mass is challenging, with no assurance of success as found in the suspect screening, like high accuracy, high resolution, and isotopic-pattern data, and increased formula assignment to identified masses. Prior to suspect and nontarget screening, researchers usually recommend a complete target screening to collect enough data for a secure structural elucidation and better identification attempts for “known unknown” substances.^78,84,85 

High sensitivity, accuracy, and rapid complete scanning makes HRMS a powerful tool for the monitoring of suspected and nontargeted compounds.⁸⁴  The instrumental analysis is usually performed twice for the screening of suspected and unknown metabolites. The first step is to perform preliminary screening to obtain full-scan precise mass data. This is followed by processing of data from the initial instrumental analysis to determine suspected intermediates using a database compiled from a large number of prior accumulations and literature reviews. The second analysis is general scanning under MS/MS conditions, with the candidates that are eventually being verified or initially confirmed on the basis of the reference standards that are available.⁸⁶  In an acquisition mode, the hybrid equipment can scan continuously and obtain full scanned spectra with continuous application of excitation energy.⁸⁷ 

Linear-ion trap high resolution coupled with Fourier-transform (LTQ-FT) Orbitrap-MS can excellently detect and identify molecules at low concentration in complex matrices.⁸⁵  It has also been proved that comprehensive analysis of the TPs can be carried by LTQ-FT Orbitrap-MS in the sample with no need of further analysis.⁸⁸  This technique is widely used for the rapid detection and quantification of PPCPs and their TPs in the sewage, as Orbitrap provides the accurate full-scan spectra, whereas LTQ simultaneously enable MS/MS measurements. HRMS is also combined with nuclear magnetic resonance (HRMS-NMR), for the rapid screening and structural elucidation of the TPs or unidentified metabolites in water matrices.⁸⁹ 

The steps in identifying the TPs are: the first phase is to identify the difference in elemental composition between the parent and the structural TP, considering the exact masses of both compounds. The second step is to compare their MS/MS spectra in order to distinguish between the structural moieties that have been transformed throughout the degrading process from those that have remained untransformed in the TPs. Instead of manually analyzing the MS/MS spectra, software-assisted algorithms can be used to accelerate the process of identification and make the process spontaneous.⁹⁰  Some of the calculation software has been used for the analyses of the MS/MS spectra and also for the determination of the structural formula of the analytes such as mass frontier, profile analysis, mass fragment, and chromalynx.⁹¹  The transformed pathway can be predicted using innovative software such as the University of Minnesota Pathway Prediction System (UM-PPS) and the Meteor Environmental Pathway Prediction System (Lhasa Limited, UK).⁹²  Figure 1.7 shows the systematic workflow for the screening of targeted, suspected, and nontargeted compounds in the environment matrices using analytical instruments.

The three main analytical methods used for the detection of CECs and their TPs in the environmental matrices are suspect screening, target, and nontarget screening of the analytes as described in Section 1.4. These approaches are coupled with advanced analytical instruments, i.e., LC along with HRMS to perform qualitative as well as quantitative analysis of the environmental samples, while there is a matrix effect in using LC-MS/MS or GC-MS/MS along with an electronic spray ion-source (ESI) interface for the detection of the organic compounds from the complex matrices. The matrix effect is due to the presence of organic matter, ion-pairing agents, natural salts, target and nontarget pollutants in the matrices affecting the ionization process.⁹⁴  In order to reduce the matrix effect, isotopically labelled internal standards have been used at the time of separation and elution within the same conditions and retention time in the ionization source. Enhanced clean-up as well as extraction methods, lowering sample matrix flow rate, minimizing the injection volume could also help in reducing the matrix effect and making the process more precise.²⁷ 

Regardless of the limitation, the LC-MS/MS is more frequently used in CECs analysis as it allows the identification and isolation of the molecules with distinct product ions but the same molecular mass if they coelute. As a result, in composite samples, the MS/MS detection increases analytical selectivity and sensitivity.⁹⁵  Monitoring of TPs in the wastewater was done by LC-QTOF-MS combined with an artificial neural network and confirmed the presence of 34 TPs.⁹⁶  Turbulent-flow chromatography (TFC) coupled with high-resolution hybrid LTQ (TFC-LTQ) Orbitrap was employed to analyze unidentified TPs of metoprolol. The ionized masses of the two new TPs are 226.1437 Da and 282.1707 Da, as confirmed by the secondary scanning by MS/MS. The UPLC coupled with TFC consist of a mixture of three solvents, i.e., pure water, methanol, and ammonium formate of 15 mM used as gradient. The analytes were eluted through a HSS T3 column with an internal diameter of 50 × 2.1 mm and a particle size of 1.8 µm. Orbitrap was used to obtain structural information of the TPs by obtaining fragmentation spectra at a resolution of 30 000.⁹⁷ 

The TPs of the triclosan, i.e., 5-hydroxy-triclosan and triclosan-O-sulfate were detected and elucidated by HRMS coupled with NMR. The preliminary detection was carried out by LC-MS/MS using an RP column of SynergiPolar with a particle size of 4 µm and 150 mm × 2 mm I.D. internal dimension. The binary solvents used for the analysis comprised of 0.2% formic acid in HPLC grade water and methanol, and LC analysis was followed by triple quadrupole API 4000 mass spectrometry maintaining ESI in the negative mode, using nitrogen gas for collision as well as a nebulizing gas and later quantified by using a multireaction monitoring scan (MRM). The structural elucidation of the TPs was analyzed by HRMS coupled with LTQ-Orbitrap-MS with a resolution of 100 000, maintaining ESI in the negative mode.⁹⁸  2,4-Dichlorophenol, 4,5-dichlror-2-(2,4-dichlrophenoxy)phenol and 5,6-dichloro-2-(2,4-dichlorophenoxy)phenol were detected by LC-MS/MS. LC-MS/MS with C-18 column (5 µm, 150 × 2.1 mm) was used and a binary mobile phase of ammonium acetate of 5 mM in distilled water and acetonitrile was used for the separation of the analyte. ESI (−ve) in MRM mode was maintained to analyze the TPs of triclosan.⁹⁹ 

Carbamazepine after oxidation by UV/Cl, generated TPs, as mentioned in Table 1.1. These were analyzed by HPLC-QToF-MS/MS. The separation column used for the identification of the TPs was an RP Atlantis T3 with 3 µm particle size and dimensions: 100 mm × 2.1 mm. The mobile phase used comprised of 0.1% formic acid in ultrapure water and acetonitrile containing 0.1% formic acid and 10% ultrapure water, respectively. The ESI was maintained in ESI (+ ve) as well as ESI (−ve) modes for the detection of the analytes.⁵⁷ 

The TPs of ketoprofen generated after ozonation, was detected using UPLC coupled with hybrid QTOF mass spectrometry in the negative mode and the column used was C18 column (30 × 2.1 mm; in dimensions and a particle size of 3.5 µm), the mobile phase used was 0.05% formic acid in ultrapure water and methanol and a flow of 0.4 mL per min. The operational conditions of the QTOF were maintained at 400 °C and 12 L per min as nitrogen gas temperature and rate of flow, respectively. The nebulizer pressure was maintained at 20 psig, the voltages of the fragmentor and the skimmer at 100 and 45 V, respectively, and the gas temperature at 325 °C. The mode for analyzing was either MS/MS or auto MS/MS keeping the collision energy fixed and mass ranges from 50–1000 m/z. The resolution and the acquisition mode of the instrument were kept at 4 GHz and 1.5 spectra per s, respectively.⁶⁴ 

The degraded products of the ciprofloxacin on treatment by ozonation was analyzed by LC coupled with ESI-MS/MS using a C18 column (150 × 4.6 mm: internal dimensions and a particle size of 5 µm). 0.1% formic acid and acetonitrile was used as the mobile phase in the gradient flow. The MS/MS instrument was scanned in positive mode and the spectra was scanned in the region from 150–1000 m/z.⁶³  A total of 81 pharmaceuticals was reported in various water matrices such as drinking water, ground water, surface, and wastewater and it has been found that the concentration of carbamazepine TP, i.e., 10,11-epoxycarbamezapine was predominant and found in higher concentration in a municipal wastewater sample.¹⁰⁰  The pharmaceuticals and its TPs were detected by UPLC-QqLIT-MS/MS using a HSS T₃ column with 50 mm × 2.1 mm dimensions and a particle size of 1.8 µm for ESI (+ve) mode, whereas a BEH C18 T₄ column with the same dimensions and a particle size of 1.7 µm, respectively, was used for ESI (–ve) mode analysis. 10 mM of ammonium acetate or formic acid and methanol was used as the binary solvent in the gradient for the positive mode and acetonitrile was coupled with 5 mM of ammonium acetate or formic acid instead of methanol in the negative ESI mode analysis. Multiple reaction monitoring scan mode was used for fragmentation monitoring in the MS/MS detector.¹⁰¹ 

Acetaminophen on treatment with UV/H₂O₂ generates TPs, as shown in Table 1.2. TPs were detected by HPLC coupled with QTOF-MS and an ESI interface (LC-ESI-QTOF-MS) using an RP column with 150 × 3 mm internal dimensions and a particle size of 4 µm. The chromatographic separation of TPs was achieved using 0.1% formic acid and mixture of 0.1% formic acid and methanol as binary solvent keeping flow rate of 400 µL per min in the gradient phase. The ESI was scanned in both the modes (+ ve and –ve), and the temperature was maintained at 50 °C, the iron spray floating voltage was kept at 4500 V, whereas the curtain gas and ion source gases (GS I &GS II) were maintained at 25 L per min and 30 and 40 psi, respectively. The resolution and mass range were set at 50–60 m/z and 40 000, respectively. The collision energy and declustering potential were set at 10 eV and 50 V, respectively.⁷⁵ 

The TPs of ibuprofen generated after treating with UV/H₂O₂ was analyzed using HPLC coupled with MS of a single quadrupole in APCI mode. An RP 18 column was used for chromatographic separation using deionized water and acetonitrile as the binary solvent (50 : 50; v/v), maintaining the flow rate and column temperature constant at 0.2 mL per min and 40 °C, respectively. An ion-chromatography system coupled with supressed conductivity detector and autosampler was used to measure the anion concentration. The analyte was eluted using a silica gel or a carboxyl-group-packed column with internal dimensions of 4 × 250 mm using 3.6 mM of sodium carbonate as a solvent at the flow rate of 0.7 mL per min.⁵⁵ 

AOPs have been studied widely as an alternative to overcome the limitation of conventional WWTPs and enhancing the removal efficiency of CECs or the organic contaminant from water that pose significant environmental and health implications. Amongst AOPs, photochemical treatment is widely preferred for treatment as it leads to degradation of the compound into carbon dioxide and water as byproducts along with some TPs. The TPs formed during AOP can be more hazardous than the parent compound and the data on TPs or metabolites is currently insufficient due to difficulties in analyzing the trace level of unknown TPs in the complex matrices and also a lack of standards. Advanced analytical instruments such as LC or GC coupled with mass spectrometers such as ToF, QqLIT, IT, HRMS, ESI, MS/MS are required for the analysis along with three analytical approaches: suspect screening, target, and nonsuspect screening.

There is a need for advancement of analytical approaches, within a single run. The development of a generic analytical procedure can enable the simultaneous identification of the parent compound and their transformed metabolites. The workflow for the identification of TP, software and the instruments used is still in progress. Prioritizing the peak and eliminating false-negative as well as false-positive peaks require suitable filtering and specific criteria. The HRMS used detect some of the suspect and nontarget analysis. Evaluation of the specific chemical structure is the difficult step in the assessment of nontarget pollutants and has demanded new research in this direction. Despite increased research efforts focused to the identification of TPs, only a few analytical approaches involving transformation-product quantification have been published. It is critical to create sophisticated instrumentation in order to identify unknown contaminants and to expand the database in order to meet the demands for new knowledge, complementary approaches, and genuine reference standards. Ecotoxicology and risk assessments of the TPs would enhance interest in this research studies areas.

The authors acknowledge the financial support of the European Union's Horizon 2020 research and innovation program in the frame of the PANIWATER project (GA 820718), funded under the Indo–EU International Water Cooperation sponsored jointly by the European Commission and the Department of Science and Technology, India. The chapter was checked for plagiarism using the iThenticate software and recorded in the Knowledge Resource Centre, CSIR-NEERI, Nagpur for anti-plagiarism (KRC No.: CSIR-NEERI/KRC/2022/APRIL/EBGD-WWTD/1).