Investigative Approaches for Pollutants in Water: Aligning with Water Framework Directive Maximum Allowable Concentrations

Abstract

In the wake of rapid advancements in the pharmaceutical, food, and agricultural industries, the environment faces an increasing influx of diverse compounds, both intentionally and unintentionally released. These compounds fall into two categories: persistent and emerging pollutants. Persistent pollutants, characterized by their resistance to degradation and potential to accumulate in the environment, pose serious ecological threats. The Water Framework Directive (WFD) plays a pivotal role in monitoring and regulating these substances. This review discusses various contemporary analytical approaches to determine problematic substances, including benzo(a)pyrene, cypermethrin, dichlorvos, heptachlor, and heptachlor epoxide, aligning with the priorities outlined in the 2013 WFD classification. This review focuses on diverse water sampling methods, sample preparation techniques, and analytical methods, encompassing chromatographic, spectroscopic, and electrochemical approaches, with the primary goal of achieving the requirement laid on analytical methods used for the determination of maximum allowable concentrations defined in the WFD. Chromatographic methods, utilizing diverse mass spectrometers, have achieved detection limits as low as 10⁻⁶ μg/L, while modern electroanalytical techniques reach levels as low as 10⁻¹³ μg/L, reflecting an ongoing collective effort to enhance monitoring and safeguard the health of aquatic ecosystems. From sampling methods, large-volume sampling and passive sampling devices have been shown to be a cost-effective and modern solution, addressing limitations in traditional sampling methods, even if both of them face important pros and cons in terms of quantitative analysis.

1. Introduction

As the pharmaceutical, food, and agricultural industries continue to advance, there is a growing proliferation of new compounds being generated and released into the environment, whether intentionally or unintentionally [1,2,3]. It is crucial to acknowledge that our modern world heavily relies on many of these compounds. Organic pollutants are commonly categorized into two groups: persistent and emerging pollutants. Emerging compounds refer to substances that endure in the environment, arising from varied and unregulated sources, posing potential ecological or human health risks, and are generally not subject to current environmental regulations. On the other hand, persistent pollutants are substances or compound groups known for their remarkable resistance to chemical, biological, and photolytic degradation, as well as their possibility to accumulate in both the environment and living organisms [1]. These persistent pollutants have the ability to travel over long distances, affecting regions where they have never been directly employed. One of the foremost concerns with such compounds lies in their profoundly harmful impact on the natural world. This realization has spurred the need for classifying and analyzing various compound groups that find their way into water, soil, and the atmosphere. In light of ongoing climate change, there is an imperative to continuously enhance and innovate analytical methods for evaluating different types of pollutants infiltrating the environment. Among the most prevalent water pollutants are pesticides and hydrocarbons. In the year 2000, the Water Framework Directive (WFD) introduced a comprehensive strategy for monitoring specific compounds, aiming to determine and evaluate their environmental impact [3,4,5,6,7]. This directive underscores the urgency of our efforts to safeguard the environment amid evolving environmental challenges.

The Priority Substances List, initially established in 2001 consisting of 33 compounds and groups of compounds (Annex X to the WFD), underwent a series of transformations [7]. It was first replaced by Annex II in 2008, as part of the Directive on Environmental Quality Standards [8], commonly known as the Priority Substances Directive. This update also set environmental quality standards for substances in surface waters. In 2013, the list underwent another revision, this time becoming Annex I to [9], which not only included environmental quality standards but also introduced additional provisions concerning chemical pollutants. In 2013, a mechanism known as the Watch List was introduced to enhance the accessibility of information for identifying the most concerning substances. Member states are required to conduct annual monitoring of the substances included in the list for a maximum of four years. The Watch List was initially established in 2015, subsequently updated in 2018, 2020, and once more in 2022. However, many compounds from these pollutant groups were not placed on the WFD list, which poses a serious challenge in further analysis. In this way, opportunities are opened for the application of different strategies of approach to the identification and analysis of different pollutants [5]. A diverse array of regulatory frameworks worldwide exists for overseeing pesticide usage and safeguarding groundwater reservoirs, along with various methodologies employed in the monitoring of groundwater [3,5,10]. Water and wastewater contain different types of pollutants that can interact in antagonistic or synergistic ways and thus have a detrimental and complex effect on the environment. Similar to solid samples, water samples are collected in proximity to suspected sources of contamination and areas where alterations in water composition are expected to take place. According to the recommendations, when it comes to rivers, samples are taken upstream and downstream from the source of contamination [11,12]. The fundamental distinction between the Watch List and the list of priority substances lies in the fact that priority substances are known to have established adverse effects, leading to specific regulatory measures such as environmental quality standards. In contrast, Watch List substances undergo more thorough scrutiny due to concerns about emerging risks, potentially without immediate regulatory mandates. Watch List substances undergo meticulous monitoring and evaluation, with the possibility of future regulatory actions contingent upon additional data and research findings. A specific challenge in this field pertains to the absence of a complete certified reference material (CRM) for primary organic substances, as outlined in the WFD. Despite years of efforts by reference material producers and proficiency testing providers, only partial solutions had been available until recently. The WFD defines “whole” water as the original water sample without separating solid matter and the liquid phase. Initial attempts to create a whole water reference material were made in 2008 with PT samples for interlaboratory comparison (IMEP-23). While reference values were assigned, these samples did not achieve CRM status due to limited stability assessment. Concurrently, other whole water reference materials were developed in the European Metrology Research Programme ENV08 project for polybrominated diphenyl ethers (PBDE), PAH, and tributyltin but also lacked certification due to limited value assignment and stability assessment. These materials differed from ERM-CA100 in matrix type and preparation method, offering various advantages and disadvantages [13].

In addition, the assessment of chemical persistence is a crucial factor in global regulations governing organic chemicals, influencing the prioritization of hazardous substances and the evaluation of the risk of exposure to living organisms. Persistence, defined as the ability of a substance to withstand environmental conditions before undergoing transformation, is evaluated through laboratory studies, environmental monitoring, and computational modeling within regulatory frameworks [14]. The challenges associated with current assessments of persistence, such as reliance on outdated tests and the need for increased accuracy, are comprehensively addressed in a review conducted by the European Centre for Ecotoxicology and Toxicology of Chemicals (ECETOC). For regulatory persistence, biodegradation tests form the foundation, including commonly used biocatalysts and microorganisms. Notably, the OECD and ISO have standardized various biodegradation tests that are widely accepted in numerous countries. One of these tests is the OECD TG 309 surface water simulation test, which can be conducted exclusively with surface water or modified with sediment to mimic water bodies with suspended solids [14]. This introduces uncertainties in degradation half-lives. In ready biodegradability tests, diverse test designs and sources of inocula lead to significant variations for a given substance. On the other hand, the biodegradation of volatile substances primarily occurs in aqueous environments. However, the role of interstitial air as a mass transfer medium for semi-volatile chemicals presents challenges in tests due to the competition between volatilization and biodegradation rates. Currently, there is no validated regulatory protocol for the routine evaluation of substances with significant volatilization rates [15]. However, despite the numerous conducted tests and databases on the half-life of individual substances, it is known that their half-life can vary significantly. This variability poses a challenge during exposure to these substances and in the risk assessment process. Influencing factors include variations such as the nature and characteristics of microbiological communities, which differ depending on environmental compartments. In addition to these conditions, other factors also play a role, such as pH, temperature, hydrolysis, photolytic processes, the concentration of O₂, and other important constituents [2,16,17,18,19,20]. Substances in the environment are often present in low concentrations, below their solubility threshold, which creates challenges in persistence testing. This is particularly true in ready biodegradability tests that typically measure biodegradation indirectly at high concentrations, leading to experimental challenges and potential artifacts. However, recent advancements in aligning biodegradation kinetic testing with modern analytics (chromatographic methods) enable testing at low, environmentally relevant concentrations [14,15]. Standardized tests are not only utilized for assessing chemicals for biodegradation purposes but can also be adapted in modified forms in other areas of environmental biotechnology. They can serve as pre-tests to evaluate potential biological treatment processes or even assist in designing the layout of actual treatment plants. While some ready biodegradability tests face criticism for their artificial test conditions, it is generally acknowledged that positive results hold significance for labeling substances as “ready biodegradable” and “non-persistent”.

The analysis of organic pollutants in water typically involves the utilization of different extractions and chromatographic techniques. Gas chromatography (GC) [21,22,23,24] and liquid chromatography (LC) [25,26,27], in conjunction with detectors such as mass spectrometers (MS) [28,29], flame ionization detectors (FIDs) [30,31], and electron capture detectors (ECDs) [24,32], are the most prevalent chromatographic methods employed for analyzing industrial pollutants and pesticides. High-resolution mass spectrometry (HRMS) techniques have the capability to detect approximately 2000–5000 substances and their transformation products in each environmental sample [33,34]. Achieving reliable qualitative and quantitative determination of water pollutants, even in trace amounts, hinges on efficient sample preparation. The most commonly employed water sample preparation method is solid-phase extraction (SPE) [28,35,36]. In addition to various SPE techniques, this review will delve into solid-phase microextraction (SPME) [23,37,38], liquid–liquid extraction (LLE) [39,40,41], and stir-bar sorptive extraction (SBSE) [42,43] methods. While SPME is environmentally friendly due to its solvent-free nature, its cost in terms of fibers is notably higher compared to solvents and sorbents used in SPE and LLE extractions. Consequently, SPME has found relatively less application in environmental pollutant research. Nonetheless, SPME offers full automation, from sample preparation to injection, resulting in generally clearer and “cleaner” chromatograms. For volatile and semi-volatile compounds, headspace SPME (HS-SPME) proves to be a more efficient and rapid method regarding mass transfer compared to direct immersion SPME (DI-SPME) [44]. Most of these modern methods aim to minimize the volume of solvents and samples used for analysis, aligning with the principles of green chemistry in environmental sample analysis. Method validation requires demonstrating satisfactory results in terms of accuracy, precision, limit of detection (LOD), limit of quantification (LOQ), and linearity [45]. When dealing with very low concentrations of compounds in a given sample, it is essential to find an optimal combination of sample preparation and analytical methodology [46,47]. The quick, easy, cheap, effective, rugged, and safe (QuEChERS) method has proven to be highly effective in extracting organochlorine pesticides (OCPs), polychlorinated biphenyls (PCBs), and PAH compounds from food and environmental samples [47,48]. Salem et al. compared the extraction efficiencies of different solvent mixtures, including acetonitrile, n-hexane/acetone (1:1, v/v), and dichloromethane/acetone (1:50, v/v), using the QuEChERS method and GC-MS analysis. The results demonstrated that the most efficient and reproducible solvent mixture is dichloromethane/acetone [48].

This review will explore different methods for the precise detection of trace concentrations of pollutants in compliance with the WFD. Furthermore,

Table 1. List of selected priority substances with annual average and maximum allowable concentrations in inland and other surface waters (μg/L) according to the WFD and US EPA.

No.	Name of Priority Substance	1 CAS Number	2 AA-EQS Inland Surface Waters	AA-EQS Other Surface Waters	3 MAC-EQS Inland Surface Waters	MAC-EQS Other Surface Waters	4 US EPA MCL
1	Dichlorvos	62-73-7	6 × 10−4	6 × 10−5	7 × 10−4	7 × 10−5	-
2	Heptachlor	76-44-8	2 × 10−7	1 × 10−8	3 × 10−4	3 × 10−5	0.4
3	Heptachlor epoxide	1024-57-3	2 × 10−7	1 × 10−8	3 × 10−4	3 × 10−5	0.2
4	Cypermethrin	52315-07-8	8 × 10−5	8 × 10−6	6 × 10−4	6 × 10−5	-
5	α-Cypermethrin	67375-30-8	8 × 10−5	8 × 10−6	6 × 10−4	6 × 10−5	-
6	β-Cypermethrin	65731-84-2	8 × 10−5	8 × 10−6	6 × 10−4	6 × 10−5	-
7	θ-Cypermethrin	71697-59-1	8 × 10−5	8 × 10−6	6 × 10−4	6 × 10−5	-
8	ζ-Cypermethrin	52315-07-8	8 × 10−5	8 × 10−6	6 × 10−4	6 × 10−5	-
9	Naphthalene	91-20-3	2	2	130	130	0.2
10	Fluoranthene	206-44-0	6.3 × 10−3	6.3 × 10−3	0.12	0.12	0.2
11	Anthracene	204-371-1	0.1	0.1	0.1	0.1	0.2
12	Benzo(a)pyrene	50-32-8	1.7 × 10−4	1.7 × 10−4	0.27	0.027	0.2
13	Benzo(b)fluoranthene	205-99-2	Footnote 5	Footnote 5	0.017	0.017	0.2
14	Benzo(g,h,i)perylene	191-24-2	Footnote 5	Footnote 5	8.2 × 10−3	8.2 × 10−4	0.2
15	Benzo(k)fluoranthene	207-08-9	Footnote 5	Footnote 5	0.017	0.017	0.2
16	Indeno(1,2,3-c,d)pyrene	193-39-5	Footnote 5	Footnote 5	Not applicable	Not applicable	0.2

Note: ¹ CAS number—Chemical Abstract Service number. ² Annual average environmental quality standards. ³ Maximum allowable concentration environmental quality standard. ⁴ United States Environmental Protection Agency (US EPA) maximum contaminant level (MCL)—the highest level of a contaminant that is allowed in drinking water. ⁵ In the case of PAH, monitoring benzo(a)pyrene alone suffices as it serves as a marker for toxicity in comparison with the biota environmental quality standards (EQS) or corresponding annual average EQS (AA-EQS) in water.

provides an overview of priority substances in accordance with the 2013 WFD classification. In this review, we will explore modern strategic methods for assessing persistent pollutants such as PAH compounds, heptachlor, and heptachlor epoxide. Additionally, we will discuss cypermethrin and dichlorvos, which are not categorized as persistent pollutants. All of these compounds possess the lowest maximum allowable concentrations (MACs) (see Table 1). This exploration will encompass various techniques of sample preparation, encompassing both sampling and extraction procedures, along with analytical methods involving GC and LC coupled to different detectors as well as electrochemical methods.

2. Literature Survey

A systematic search was conducted to identify articles that address or discuss the analysis of surface water samples, sometimes including sediments and biota, with regard to the content of PAH compounds and four pesticides: dichlorvos, heptachlor, heptachlor epoxide, and cypermethrin. These compounds are on the list of priority substances in the WFD. The journals under investigation spanned approximately 15 years, from 2008 to 2023, but the published results even date back to late 2005. The articles were located through searches on Google Scholar and websites of prominent publishing houses such as ScienceDirect, Elsevier, Wiley, and Springer. Data were gathered from locations influenced by human activities, such as industrial areas, megacities, and harbors, as well as from areas with minimal anthropogenic impact. To align with the research objectives, a systematic search model was established, and inclusion and exclusion criteria were defined. The inclusion criteria for the articles encompassed the following: (i) original research papers, (ii) articles published in English, (iii) articles exploring novel methods for the sampling, extraction, and analysis of all or some of the mentioned compounds, focusing on quantifying their content and concentration in water samples, and (iv) articles investigating new sampling, extraction, and analysis methods or modifications to existing methods to obtain more comprehensive, precise, efficient, and detailed results. The exclusion criteria comprised the following: (i) articles written in languages other than English; (ii) studies that solely focused on the analysis of soil, sediments, or living organisms; (iii) articles that remained incomplete even after attempts to obtain missing parts from the authors. Following a thorough review of the abstracts of all identified articles, those that did not meet the inclusion criteria were omitted. The selected studies were then analyzed based on various parameters including their year of publication, country of origin, objectives, variables, preparation methods, analytical techniques, results, and the conclusions and/or recommendations provided by the authors. In total, 154 articles meeting the criteria were identified and cited in this study. The search for relevant articles utilized a combination of both general and specific keywords related to the chemical analysis of water (as outlined in

Table 2. Search terms employed in the literature exploration.

Thematic Areas	Specific Terms
Water Framework Directive	Water pollutuion, water analysis, maximum allowable concentration extration methods, maximum allowable concentration chromatography, liquid chromatography, gas chromatography, mass spectrometry, spectroscopy, electrochemical methods, biosensors.
Polyycyclic aromatic hydrocarbons (PAH)	Water pollutuion, water analysis, maximum allowable concentration extration methods, maximum allowable concentration chromatography, liquid chromatography, gas chromatography, mass spectrometry, spectroscopy, electrochemical methods, biosensors, naphthalene, anthracene, fluoranthene, benzo(b)fluoranthene, benzo(k)fluoranthene, benzo(a)pyrene, benzo(g,h,i)perylene, indeno(1,2,3-c,d)pyrene.
Pesticides	Water pollutuion, water analysis, maximum allowable concentration extration methods, maximum allowable concentration chromatography, liquid chromatography, gas chromatography, mass spectrometry, spectroscopy, electrochemical methods, biosensors, cypermethryn, dichlorvos, heptachlor, heptachlor epoxide.
Passive samplers	Water pollutuion, water analysis, maximum allowable concentration extration methods, maximum allowable concentration chromatography, liquid chromatography, gas chromatography, mass spectrometry, spectroscopy, electrochemical methods, biosensors, polycyclic aromatic hydrocarbons, pesticides.

Note: Phrases within identical boxes were linked using “or” during the search process, while phrases located in separate rows were connected using “and” in the search.

) using the Boolean “and” and “or”.

In this study, the bibliometric mapping technique was utilized to visually represent scientific information. Bibliometric network maps were developed using VOSviewer software (version 1.6.5., Leiden University, Leiden, The Netherlands) by employing co-occurrence term data. The co-occurrence map in networking demonstrates the connection between relevant text data in the references used in the analysis. VOSviewer is a software tool employed for processing keywords and grouping analyses to create topographic network maps based on a coincidence matrix, which allows the formation of clusters determined by specific term co-occurrence [49]. In Figure 1, the co-occurrence term map network displays 11 obtained clusters with 84 selected co-occurrence terms and 515 links. Within each cluster, various circle sizes were employed to illustrate how often each keyword appeared within the subset of 154 references out of the total 162 examined. To clarify, a larger circle signifies that the keyword was more frequently used in these publications. The proximity of terms to each other indicates their topic similarity and relative importance. Circles of the same color imply a shared topic within the publications. Based on the results, it can be observed that the largest cluster centers around the words “pah”, “water”, and “pesticide” while the “water framework directive” cluster links 10 terms (Figure 1).

3. Polycyclic Aromatic Hydrocarbons (PAHs): Sources, Properties, and Migration

The Environmental Protection Agency (EPA) has assembled a list of priority pollutants, which encompasses 16 PAH compounds. These 16 compounds include naphthalene (Nap), acenaphthylene (Acy), acenaphthene (Ace), fluorene (Flu), phenanthrene (Phen), anthracene (Ant), fluoranthene (Flt), pyrene (Pyr), benzo(a)anthracene (B(a)A), chrysene (Chr), benzo(b)fluoranthene (B(b)F), benzo(k)fluoranthene (B(k)F), benzo(a)pyrene (B(a)P), benzo(g,h,i)perylene (B(g,h,i)P), indeno(1,2,3-c,d)pyrene (I(1,2,3-c,d)P), and dibenz(a,h)anthracene (D(a,h)A). Conversely, the list of priority substances within the WFD encompasses only 8 of the 16 PAH compounds originally proposed by the EPA. These eight PAH compounds consist of B(a)P, B(b)F, B(g,h,i)P, B(k)F, I(1,2,3-cd)P, Ant, Flt, and Nap (Table 1). It is noteworthy that B(a)P, regarded as the most toxic, serves as a marker for all the others as per the WFD [13].

Approximately 45 tons of PAH are estimated to enter the Mediterranean Sea each year through atmospheric deposition [50]. Dong and colleagues introduced an innovative hybrid model for predicting PAH concentrations in surface water, which combines a complete ensemble empirical mode decomposition with adaptive noise (CEEMDAN) and variational mode decomposition (VMD) with a deep learning approach of long short-term memory (LSTM). This model outperformed other prediction methods, exhibiting impressive performance metrics with a mean absolute error of 27.89, root mean square error of 37.92, and a coefficient of determination (R²) value of 0.85. These results highlight the model’s effectiveness in managing surface water quality in German rivers within the Saxony region [51]. PAH sources are typically categorized as pyrogenic, petrogenic, and diagenic. Pyrogenic sources stem from natural combustion processes or human activities, leading to the formation of complex mixtures of high-molecular-weight PAH [13,16]. These compounds result from the incomplete combustion of materials such as wood, coal, and fossil fuels and can be released into the environment during activities such as oil well construction, transportation, storage, and processing of oil [16,47]. However, natural events such as volcanic eruptions and forest fires can also contribute smaller quantities of PAH compounds to the environment. Petrogenic PAHs, on the other hand, are typically generated during the transportation and accidental spillage of crude oil and its derivatives. These PAHs often consist of two to four aromatic rings. When they enter aquatic systems, only a small percentage dissolves in water, with the majority being distributed among colloidal and suspended particles, sediments, and living organisms [52]. The distribution of petrogenic PAH is influenced by various physicochemical properties and can be described using distribution coefficients. Identifying the specific sources of PAH can be challenging due to the presence of multiple sources and the continuous transformation processes that PAHs undergo. To address this, researchers often employ correlation factors and ratios of individual PAH compounds as diagnostic tools such as the alkylphenanthrene/phenanthrene, phenanrene/anthracene, naphthalene/phenanthrene, and low-molecular-weight PAH/high-molecular-weight PAH ratios [16,52,53]. Correlation factors are based on the idea that each source emits a characteristic pattern of PAH compounds [30,47]. When they reach the environment, they have a distinct ability to migrate. In waters with limited circulation, the migration of PAH compounds is notably enhanced, allowing for increased interaction with the effectively contaminated sediment. Guigue and colleagues conducted a study in the Mediterranean Sea, specifically in the Toulon harbor, a key French Navy installation, to investigate the kinetics and concentration of PAH compounds. They employed dichloromethane as a solvent for extracting PAH from both water and sediment samples. Subsequently, they dissolved the extract in n-hexane following the application of a rotary evaporator. The presence of 34 PAH compounds was determined using a GC-MS method with an HP-5MS ultra-inert column. The composition of PAH compounds remained fairly consistent with sediment depth, while the isomeric ratio of individual members suggested that the PAHs were derived from forest or coal combustion rather than crude oil and its derivatives. In aqueous samples, PAH with two and three rings dominated, with lower concentrations of those with four and five rings and PAH with six rings falling below the detection limit. Their research revealed that a cubic function, a polynomial function of degree 3, provided the most accurate model to describe the kinetics of PAH migration from sediments to the aquatic environment. This migration was influenced by various factors, including the hydrophobicity of PAH, sediment grain size, organic matter quality, dissolved organic matter, and the duration of contact between water and sediments. Importantly, the concentrations of released PAH in water were found to be significantly below the recommendations set by the WFD [54].

Most industrial plants that use oil or petroleum products as a heating medium have a problem with the concentration of PAH compounds that occur during fuel combustion; thus, many regulations have been adopted that include the use of filtration agents [55,56,57]. When they reach the aquatic environment, PAH compounds are subject to chemical and biochemical transformation processes. The primary pathway of deterioration typically encompasses photooxidation and the breakdown facilitated by various bacterial strains (including Cycloclasticus, Pseudomonas, Halomonas, Pseudoalteromonas, Marinomonas, Bacillus, Dietzia, Colwellia, Acinetobacter, Alcanivorax, Salinisphaera, and Shewanella). Notably, Dietzia is the most prevalent among these bacteria and is consistently found in all sediment samples [50,53,56,58]. However, during these processes, it is necessary to take into account the possible formation of other pollutants such as halogenated hydrocarbons. The concentration and stability of PAH compounds in water are affected by several factors such as temperature, pH, oxygen concentration, and the presence of other substances. It has been shown that the concentration of PAH compounds decreases with increasing temperature as well as oxygen concentration [16,17]. The primary degradation product that frequently emerges is carbon monoxide. In aquatic ecosystems, PAHs tend to accumulate not only in fish and other organisms but also form intriguing aggregates of PAH particles. These aggregates are notable due to their limited solubility in water and their deposition in sedimentary environments [1,16,17]. Some studies suggest that the carcinogenic risk associated with PAH compounds increases as their molecular weight rises [17,27,55]. Since PAH compounds are typically found in mixtures, pinpointing the precise mechanism by which an individual PAH compound induces cancer can be challenging. Nonetheless, like most carcinogenic substances, it is believed that covalent binding to DNA is involved [17,53]. Furthermore, the accumulation of PAH compounds can take place on floating materials in water. Bouhroum and colleagues investigated the PAH and PCB content in plastic samples collected from aqueous environments using a solid–liquid extraction technique. PAH and PCB compounds exhibit hydrophobic interactions with plastic materials. However, exposure to solar radiation can lead to a reduction in the concentration of PAH compounds on polymeric materials due to photooxidation. The concentration of PAH compounds in plastic materials depends on several factors such as the duration of exposure and type of polymer [59]. In

Table 3. A concise overview of primary analytical methods used to detect water pollutant compounds.

1 Instrumentation	Analytes	Sample	Extraction Type	Mode	2 LOD	3 LOQ	References
GC-MS (Rxi-PAH)	PAH	Water	PAL SPME Arrow	Thermal desroption	0.1 × 10−3 to 0.8 × 10−3 μg/L	0.4 × 10−3 to 0.0026 μg/L	[21]
GC-HRMS (TR-DIOXIN-5MS)	PCB, PAH, and BDE	Surface water	SPME	Thermal desroption	/	0.1 × 10−3 to 0.05 μg/L	[22]
GC-MS (HP-5MS)	PAH	Water	LLE, sonication, PAL SPME Arrow	Liquid injection (splitless), thermal desorption	0.05 to 0.23 μg/kg (flavored tobaccos and their wastes); 0.03 to 0.17 μg/kg (traditional tobaccos and their wastes); 0.1 × 10−3 to 0.8 × 10−3 μg/L (liquid samples)	0.12 to 0.72 μg/kg (flavored tobaccos and their wastes); 0.11 to 0.55 μg/kg (traditional tobaccos and their wastes); 0.4 × 10−3 to 0.0025 μg/L (liquid samples)	[23]
GC-ECD (Zebron ZB-5MS)	OCP	Water	SPME	Liquid injection (splitless)	0.3 to 3.0 μg/L	1.0 to 10.0 μg/L	[24]
HPLC-UV (C18)	OPPs and their oxon-derivatives	Water	Selective SPE and MIP	Liquid injection	0.07 to 0.12 μg/L	0.23 to 0.41 μg/L	[25]
GC-FID/MS (5% phenylmethylsilicone)	PAH	Aqueous samples	Large-volume sampling	Large-volume injection using the through oven transfer adsorption–desorption	0.0215 to 0.211 μg/L	0.0717 to 0.7033 μg/L	[26]
GC×GC-TOF-MS (Rxi-5SilMS and Rxi-200)	PAH	Water and sediment	LLE followed by SPE	Liquid injection (splitless)	/	/	[27]
GC-QqQ-MS/MS (VF-17MS Factor Four)	PAH	Wastewater	SBSE and SPE	Large-volume injection (split/splitless)	0.001 to 0.01 μg/L	0.005 to 0.1 μg/L	[28]
GC-MS (Optima 5-MS) and LC-MS/MS	PAH, alkylated PAH, heterocyclic PAH (NSO-HET), and phenols	Water	LLE followed by SPE	Liquid injection	/	0.5 × 10−3 to 2.3 × 10−3 μg/L for EPA-PAH; 0.4 × 10−3 to 2.4 × 10−3 μg/L for alkylated PAH; 0.2 × 10−3 to 0.8 × 10−3 μg/L for S-HET; 0.6 × 10−3 to 9.8 × 10−3 μg/L for O-HET; 0.2 × 10−3 to 2.1 × 10−3 μg/L for basic N-HET	[29]
GC-FID (TR-5MS)	PAH	Seawater	Ultrasonic extraction followed by column chromatography	Liquid injection (splitless)	0.010 to 0.49 μg/L	/	[30]
GC-FID (Elite-5HT)	PAH	Water and sediment	LLE (water), Soxhlet extraction (sediment)	Liquid injection	44 to 4290 μg/L	/	[31]
GC-ECD (chrompack CP-Sil 8 CB fused-silica)	OCP	Water and fruit	HLLME	Liquid injection (split)	0.001 to 0.03 μg/L (water), 0.005 to 0.1 μg/L (fruit)	/	[32]
GC-MS (Rxi-5 Sil-MS)	PAH	Water	SPE	Liquid injection	0.004 to 0.026 μg/L	0.011 to 0.078 μg/L	[35]
GC-ECD (CPSil8 CB)	OCP and PCB	Water and sediment	Sonication (suspended particulate phase), SPE (water), sonication (sediments)	Liquid injection (split)	0.005 × 10−3 to 0.044 × 10−3 μg/L (suspended particulate phase), 0.004 × 10−3 to 0.110 × 10−3 μg/L (water), 0.2 × 10−3 to 4.4 μg/kg (sediments)	/	[36]
GC-MS (DB-5)	Dichlorvos, cybutryne, terbutryn, and quinoxyfen	Water	PAL DI-SPME Arrow	Liquid injection (splitless)	0.7 to 0.554 μg/L	0.23 to 1.960 μg/L	[37]
GC-MS (Rxi-PAH)	PAH	Water	PAL SPME Arrow	Thermal desroption	0.9 × 10−3 to 0.0036 μg/L	/	[38]
GC-APCI-MS/MS (Rtx-5MS)	OPP	Surface water	LLE	Liquid injection (splitless)	1.25 × 10−5 to 1.25 × 10−4 μg/L	/	[39]
HPLC-UV (LC-PAH Supelcosil)	PAH	Wastewater and sediment	LLE (wastewater), sonication (sediments)	Liquid injection	0.01 to 0.51 μg/L	0.03 to 1.71 μg/L	[41]
Microfluidic devices (PDMS) and HPLC with fluorescence detection	PAH	Water	SBSE	Liquid desorption	/	0.4 × 10−3 to 0.029 μg/L (microfluidic device)	[42]
GC×GC-TOF-MS (Rxi-5SilMS and Rxi-200)	OCP	Water	SBSE	Liquid injection (splitless)	0.01 × 10−3 to 0.044 × 10−3 μg/L	/	[43]
GC-MS/MS (DB-5MS)	33 WFD priority substances	Wastewater, surface and drinking water	DLLME	PTV	0.1 × 10−3 to 2.6 × 10−3 μg/L	0.3 × 10−3 to 8.8 × 10−3 μg/L	[60]
GC-MS (HP-5MS)	16 EPA PAH	Water	SPE	Liquid injection	/	/	[61]
GC-MS (Zebron ZB 5)	PAH	Aqueous samples and sediment	SPE (aqueous samples), LLE (sediments)	Liquid injection (splitless)	/	0.001 to 0.005 μg/L	[62]
GC-MS (DB-5MS)	16 EPA PAH	Water	SPE	Liquid injection (splitless)	0.002 to 0.0085 μg/L	/	[63]
GC-MS (DB-17MS)	PAH and PAH derivatives	Surface water	SPE	Liquid injection	0.1 × 10−3 to 0.4 × 10−3 μg/L	0.2 × 10−3 to 0.0188 μg/L	[64]
GC-MS (HP-5MS)	PCB, PBB, pesticides, PAH, phthalate esters, alkylphenols, and bisphenol A	Wastewater	SPE	PTV	/	0.001 to 0.198 μg/L	[65]
GC-MS (HP-5MS)	PAH	Surface water and sediment	SPE (water), microwave digestion (sediments) followed by column chromatography	Liquid injection	0.04 × 10−3 to 0.65 × 10−3 μg/L (water); 0.09 to 0.65 μg/kg (sediments)	/	[66]
UHPLC-ESI-qTOF-MS/MS (Acclaim RSLC C18 column), GC-EI-MS (HP-5MS), Headspace-GC-EI-MS (DB-624MS), GC-NCI-MS (HP-5MS), ICP-MS	Metals, bisphenol A, bis(2-ethylhexyl) phthalate, prometryn, BDE, PAH, PCB, and degradation products of DDT	Surface water and sediment	Ultrasonic-assisted extraction	Water and sediment	/	/	[67]
GC-ECD (Rtx-5MS)	OCP and pyrethroid pesticides	Water	SDME	Liquid injection (split ratio—1:5)	0.003 to 0.6 μg/L	0.01 to 2 μg/L	[68]
GC-ECD (Rtx-5MS)	PAH, nitro-PAH, and quinones	Surface and groundwater	SDME	Liquid injection (splitless)	0.01 to 0.02 μg/L	0.05 to 0.08 μg/L	[69]
GC-MS (Select PAH)	PAH	Tissues of freshwater mussels—Unionidae	FUSLE and DSPE	Liquid injection (splitless)	0.25 to 0.79 μg/kg	0.76 to 2.4 μg/kg	[70]
GC-MS (Select PAH)	PAH	Crustacean gammarids	FUSLE and DSPE	Liquid injection (splitless)	0.036 to 0.17 μg/kg	0.12 to 0.56 μg/kg	[71]
GC-FID	PAH	Seawater	Magnetic SPE (gold immobilized magnetic mesoporous silica nanoparticles) coupled with DLLME	Liquid injection	0.002 to 0.004 μg/L	/	[72]
GC-MS (SGE 25QC3/HT8)	PAH	Surface water	ASE	Liquid injection			[73]
GC-MS (SPB 20)	16 PAHs	Water and sediment	Column chromatography (suspended particulate phase), SPE (water), sonication (sediments)	Liquid injection	0.01 × 10−3 to 0.1 × 10−3 μg/L (water); 0.03 × 10−3 to 0.2 × 10−3 μg/L (suspended particulate matter); 0.01 to 0.15 μg/kg (sediment)	0.02 × 10−3 to 0.15 × 10−3 μg/L (water); 0.06 × 10−3 to 0.3 × 10−3 μg/L (suspended particulate matter); 0.03 to 0.2 μg/kg (sediment)	[74]
HPLC-UV with fluorescence detection (C18), GC-MS (XTI-5)	PAH	Water and sediment	SPE (water), sonication (sediment)	Liquid injection	0.0021 to 0.102 μg/L (water); 0.4 to 26.5 μg/kg (sediment)	/	[75]
Fluorescence spectroscopy and GC-MS	PAH and pesticide	Sub-surface water	SPE	Liquid injection	0.02 to 1.29 μg/L (fluorescence spectroscopy)	0.07 to 4.3 μg/L (fluorescence spectroscopy)	[76]
Microfluidic silicon/glass chips (PDMS, porous SiOCH), HPLC with fluorescence detection	PAH	Water	SBSE	Liquid desorption	/	0.01 to 1.0 μg/L (microfluidic device)	[77]
Electrochemical DNA biosensor, HPLC with fluorescence detection	PAH	Mytilus galloprovincialis	Solid–liquid extraction followed by column chromatography	Liquid desorption	/	/	[78]
Sensing platform based on π–π interaction with thionine-graphene composite electrochemical method	16 PAH with priority on phenanthrene and anthracene	Liquid sample containing 16 PAHs	/	/	1.7823 × 10−5 μg/L	1.7823 × 10−10 to 0.17823 μg/L	[79]
Polypyrrole-composite-film-based electrochemical method	PAH	0.1 mol/L sodium acetate-acetic acid solution (pH 5.0) containing PAH	/	/	1.0 × 10−13 μg/L	/	[80]
Electrochemical detection (polished glassy carbon electrode)	Benzo(a)pyrene	Acetonitrile–water binary medium	/	/	1.6905 × 10−6 μg/L	/	[81]
HPLC-DAD (Acclaim 120-C18)	Carbaryl and cypermethrin	Water and soil	LLE (water), liquid–solid extraction (soil)	Liquid injection	/	/	[82]
HPLC-UV (Alltima C18)	Dichlorvos	Water	DLLME	Liquid injection	0.2 μg/L	/	[83]
HPLC-UV (Alltima C18)	Dichlorvos	Water	Ionic-liquid-based dispersive liquid-phase microextraction	Liquid injection	0.2 μg/L	/	[83]
Fluorescence spectroscopy	OPP (chlorpyrifos, malathion, endosulfan, and heptachlor)	Water	/	Quartz cell, excitation wavelength at 365 nm	1.6478 × 10−4 to 4.1504 × 10−4 μg/L	/	[84]
GC-MS (for organic compounds, narrow-bore fused-silica capillary column), ICP-MS (for metals)	WFD 45 priority substances	Water	LLE (for organic substances), microwave extraction (for metals)	/	0.01 × 10−3 to 1 μg/L (for organic substances); 0.1 × 10−3 to 1 μg/L (for metals)	0.01 × 10−3 to 0.3 μg/L (for organic substances); 0.03 × 10−3 to 0.03 μg/L (for metals)	[85]
Methods for organic compounds: LC-MS/MS, GC-MS, GC-MS/MS; methods for elements: ICP-MS method	WFD Priority Substances (45 priority substances and 250 specific pollutants)	Water	SBSE for organic compounds	Liquid injections	0.005 × 10−3 to 2 μg/L (for organic compounds); 0.021 to 100 μg/L (for elements)	/	[86]
GC-MS/MS (HP-5MS)	117 endocrine disruptors according to the WFD	Water	SBSE	PTV with thermal desorption	0.01 × 10−3 to 0.005 μg/L	0.12 × 10−3 to 0.05 μg/L	[87]
GC×GC-TOF-MS (Rxi-5SilMS and Rxi-17SilMS)	53 pesticides	Surface and groundwater	SPE	Liquid injection (split ratio—1:10)	0.3 × 10−3 to 0.0022 μg/L	0.0011 to 0.0072 μg/L	[88]
HPLC (Waters Atlantis T3) with 484 tunable absorbance detector	Pyrethroid pesticides (tetramethrin, fenpropathrin, cypermethrin, deltamethrin, fenvalerate, and permethrin)	Fruit juice	DLLME	Liquid injection	2 to 5 μg/L	5 to 10 μg/L	[89]
GC-TOF-MS (DB-17MS), GC-ECD (RtxPesticides fused silica), HPLC-DAD (Terra RP18)	Synthetic pyrethroids, chlorpyriphos, and neonicotinoids	Surface water	LLE	Liquid injection (splitless for GC)	0.05 to 0.3 μg/L (for GC), 0.03 to 0.3 μg/L (for HPLC)	/	[90]
HPLC	Pesticide residue	Water	LLE	Liquid injection	0.01 μg/L	/	[91]
GC-MS (ZB-5)	OCP (lindane, heptachlor, and heptachlor epoxide isomers)	Groundwater	SPME	Liquid injection (splitless)	/	0.03 to 0.1 μg/L	[92]
GC-ECD (HP-5)	OCP	Fish	Miniaturized matrix solid-phase dispersion, homogeneous LLE	Liquid injection (split ratio—1:100)	0.4 to 1.2 μg/kg	1.3 to 4 μg/kg	[93]
GC-ECD (5% diphenylpolymethylsiloxane) and NI-GC-MS (5% phenyl methyl polysiloxane)	OCP	Water	SPE	Liquid injection (splitless)	/	/	[94]
GC-MS (TR-5MS)	10 OCP	Water	DLLME-SFO	Liquid injection (splitless)	0.06 to 3.00 μg/L	0.20 to 10.00 μg/L	[95]
GC-MS (HP-5MS)	OCP	Aqueous samples	DLLME	Liquid injection (pulse splitless)	0.4 × 10−3 to 0.0025 μg/L	/	[96]
GC-MS (Zebron ZB-Multiresidue-1 fused silica)	34 pesticides	Water	DLLME	Liquid injection (splitless)	0.0032 to 0.0174 μg/L	0.0096 to 0.052 μg/L	[97]
GC-MS (Rtx-5MS)	39 pesticides	Water	DMSPE	Liquid injection	0.51 to 0.0224 μg/L	0.00171 to 0.0745 μg/L	[98]
GC-MS (HP-5MS)	OCP	Water	Microextraction in packed syringe	Liquid injection (splitless)	0.02 to 0.19 μg/L	/	[99]
GC-MS (TR-5MS)	Organochlorine insecticide residues	Water	LLE	Liquid injection	/	/	[100]
HPLC-UV (ZORBAX Eclipse XDB-C18)	OPP (dichlorvos and malathion)	Water	SPE	Liquid injection	0.1 to 0.4 μg/L	/	[101]
Colorimetric analysis (quaternized carbon dots and Ellman’s test)	Dichlorvos	Juice	/	/	1.31 × 10−11 to 5.6 × 10−6 μg/L	/	[102]
GC-MS (HP-5MS)	OPP	Water	SPE	Liquid injection	0.002 to 0.597 μg/L	0.006 to 1.990 μg/L	[103]
GC-ECD (HP-5)	17 OCP	Water	SPE and HS-SPME	Liquid injection, thermal desroption	0.0018 to 0.027 μg/L	/	[104]
HPLC-DAD-CL (Kinetex C18)	Organothiophosphorus pesticides	Water	SPE	Liquid injection	0.8 to 6 μg/L (without SPE); 1.2 to 8 μg/L (without SPE; DAD); 0.015 to 0.08 μg/L (250 mL SPE); 0.006 to 0.03 μg/L (1000 mL SPE)	2.7 to 20 μg/L (without SPE); 4 to 27 μg/L (without SPE; DAD); 0.05 to 0.30 μg/L (250 mL SPE); 0.017 to 0.1 μg/L (1000 mL SPE)	[105]
UV-spectrophotometry (ascorbic acid functionalized gold nanoparticles)	Dichlorvos	Water, apple juice, and wheat	Solid–liquid extraction	/	9.4887 × 10−3 μg/L	/	[106]
Screen-printed, amperometric biosensor	OPP	Water	/	/	4 to 7 μg/L	/	[107]
Acetylcholinesterase-based biosensor	Dichlorvos	Potential for different samples	/	/	2.20976 × 10−12 μg//L	/	[108]
Flow injection analyzer (acetylcholinesterase micro-electrode arrays)	OPP (dichlorvos, parathion, and azinphos)	Pesticide solution	/	/	2.20976 × 10−12 to 3.17324 × 10−11 μg/L	/	[109]
Electrochemical measurements (graphene/CdSe@ZnS or graphene/CdSe@ZnS/AChE electrode as the working electrode)	OPP	Fruit	Solid–liquid extraction	/	2.75195 × 10−9 to 2.20976 × 10−7 μg/L	/	[110]
Acetylcholinesterase-based biosensor	OPP (fenitrothion, dichlorvos, and malathion)	Water	/	/	4.41952 × 10−8 to 2.31251 × 10−7 μg/L	/	[111]

Note: ¹ Analytical method employed in the analysis with details in parentheses (column information in case of chromatographic analysis). ² Limit of detection. ³ Limit of quantification. Abbreviations: GC—gas chromatography, LC—liquid chromatography, HPLC—high-performance liquid chromatography, DAD—diode array detector, CL—chemiluminescence, ECD—electron capture detector, APCI—atmospheric-pressure chemical ionization, MS—mass spectrometry, MS/MS—tandem mass spectrometry, TOF-MS—time-of-flight mass spectrometry, QqQ-MS—triple quadrupole mass spectrometry, ICP—inductively coupled plasma, PDMS—polydimethylsiloxane, HRMS—high-resolution mass spectrometry, UHPLC—ultra high performance liquid chromatography, ESI—electrospray ionization, qTOF—quadrupole time of flight, EI—electron impact, NCI—negative chemical ionization, EPA—Environmental Protection Agency, WFD—Water Framework Directive, PAH—polycyclic aromatic hydrocarbon, OCP—organochlorine pesticide, OPP—organophosphorus pesticide, PBB—polybrominated biphenyl, PCB—polychlorinated biphenyl, NSO-HET—nitrogen-, sulfur-, and oxygen-containing heterocycles, LLE—liquid–liquid extraction, SPE—solid-phase extraction, SPME—solid-phase microextraction, PAL—prep and load, DI-SPME—direct immersion solid-phase microextraction, SBSE—stir bar sorptive extraction, HS-SPME—headspace solid-phase microextraction, DLLME—dispersive liquid–liquid microextraction, DLLME-SFO—dispersive liquid–liquid microextraction based on the solidification of floating organic droplets, ASE—accelerated solvent extraction, FUSLE—focused ultrasound solid–liquid extraction, DSPE—dispersive solid-phase extraction, SDME—single-drop microextraction, DMSPE—dispersive micro-solid-phase extraction, PTV—programmable temperature vaporization, BDE—brominated diphenyl ether, FID—flame ionization detector, HLLME—homogeneous liquid–liquid microextraction.

, an overview of analytical methods commonly used for determining PAH compounds in water samples, as well as other types of samples, is presented. Additionally, the table displays the LOD and LOQ values achieved by these methods.

3.1. Sample Preparation Approaches and Gas Chromatographic Detection of PAH

Edokpayi and colleagues employed LLE to extract PAH compounds from water samples from South Africa’s Mvudi and Nzhelele rivers. After extraction, they purified the organic phase with SPE (florisil) and analyzed it using comprehensive GC time-of-flight mass spectrometry (GCxGC-TOF-MS). The GC system used Rxi-5SilMS as the primary column, Rxi-200 as the secondary column, and a thermal modulator with four jets. Their findings indicated an increase in PAH compounds with more rings, which are known to be more persistent. They concluded that the source of these compounds is anthropogenic based on the relationship between low- and high-molecular-weight PAH [27]. Nekhavhambe et al. examined the content of six PAH compounds (indene, azulene, dibenzothiophene, Ant, Flt, Pyr) in 19 samples of ode and river sediments with GC-FID with Elite-5HT capillary column [31]. Water samples were extracted by using the LLE method, while sediment samples were extracted by using the Soxhlet extraction method. The ratio of fluoranthene/(fluoranthene+pyrene) refers to the petrogenic source of these compounds, such as traffic density. The lowest determined concentration of PAH in water samples was 38.2 μg/L while the highest was 2553.1 μg/L. Dispersive liquid–liquid extraction (DLLME) provides an alternative to the traditional LLE method, as it entails dispersing tiny droplets of the extraction solvent within an aqueous sample. This method was used to extract 33 priority substances from the initial WFD list. This method consisted of taking a 35 mL water sample, adding sodium chloride, and employing a solvent mixture (3.2 mL of acetonitrile as a disperser and 75 μL of 1,1,2-trichloroethane as the extraction solvent). Extraction was carried out for 1 min; however, it was observed that this parameter had minimal impact on extraction efficiency. GC-MS/MS analysis was conducted using a DB-5MS stationary phase. The LOD and LOQ for four pesticides and PAH compounds were in the range from 0.2 × 10⁻³ μg/L to 8.8 × 10⁻³ μg/L and from 0.2 × 10⁻³ μg/L to 5.6 × 10⁻³ μg/L, respectively. Following method optimization, it was applied to samples from the Llobregat river in Spain, both before and after water treatment at three different locations. None of the mentioned compounds were detected in surface water samples, except for terbutin, lindane, and dichlorophen [60].

SPE employs various sorbent materials, including alumina, activated carbon, polymers, and nanostructure-based sorbents, to extract PAH from water. Porous activated carbons, typically derived from coal and high-carbon-content waste organic materials, are a common choice. In a study by Kamran and colleagues, they introduced a novel SPE sorbent made from asphalt, a crude oil derivative. Intriguingly, the quantity of sorbent in the SPE cartridges (ranging from 25 to 250 mg) did not significantly affect the results, making this sorbent appealing due to its reproducibility and potential for reuse with proper rinsing. The LOD and LOQ for individual PAH with ranges of 0.004 to 0.026 μg/L and 0.01 to 0.07 μg/L, respectively, fell within the WFD-mandated MAC values. Structural analysis of the sorbent revealed a notable concentration of hydroxy and carboxyl groups on its surface, which are believed to enhance its affinity for PAH compounds [35]. Kouzayha and their team fine-tuned the SPE process using a C18 cartridge, with a focus on reducing the method duration. Their optimization efforts encompassed various parameters, such as sample volumes, elution phase volume, centrifugation effects, solvent selection and volume, temperature, nitrogen flow rate, and the duration of sorbent drying [61]. In a separate study, Werres and colleagues developed an automated method for detecting PAH compounds under the WFD. They utilized an SPE disk containing a C18 sorbent phase, allowing for the extraction of PAH compounds whether they were dissolved in water or associated with other particles. The GC method achieved low LOQ ranging from 0.001 to 0.005 μg/L. Furthermore, the SPE disk approach demonstrated superior efficiency in terms of compound recovery compared to the conventional LLE method [62]. Regarding sample volume selection, the most favorable range was determined to be between 500 mL and 1000 mL. A sequence of investigations consisting of five steps demonstrated that drying the sorbent for more than 20 min caused substantial reductions in the presence of lower PAH compounds. Additionally, surpassing a drying time of 30 min resulted in almost complete absence. The research team identified an optimal drying duration of 10 min, yielding recoveries between 77% and 82%. Dichloromethane emerged as the preferred solvent, demonstrating remarkably high recoveries of up to 99% with a 9 mL volume, as reduced solvent volumes negatively impacted the recovery of higher PAH. To eliminate excess solvent prior to instrumental analysis, the recommended temperature range was established between 35 °C and 40 °C. However, even at these temperatures, losses of lower PAH were observed. In an innovative approach, carbon nanotubes were employed as a relatively new sorbent for SPE in the analysis of 16 PAH compounds. Carbon nanotubes consist of multiple closely bonded graphene layers, and n-hexane was used for extraction. Separation was carried out using the GC-MS method with a DB-5MS fused-silica column. Recoveries using nanotubes were approximately 10% higher compared to C18 cartridges, spanning from 75.3 to 125.7%. While C18 proved to be more efficient in extracting higher PAH due to hydrophobic interactions, nanotubes exhibited weaker physical absorptions and bonds based on delocalized π electrons in their interactions with PAH [63]. Qiao and colleagues collected water samples from different areas within the Chaobai river System near Beijing, China, including urban, agricultural, and mountainous zones, over four seasons. They prepared the samples by performing solid-phase extraction with a C18 column and elution using dichloromethane and n-hexane solvents. They analyzed the samples using GC-MS with a DB-17MS fused silica capillary column. The PAH concentration averaged 0.158 ± 0.105 µg/L, exceeding the WFD maximum allowable concentration. Surprisingly, despite the urban area’s higher population density and shift toward eco-friendly energy sources, all locations exhibited similar PAH concentrations during the winter season. The relationships of individual PAHs have confirmed the fact that the main sources in mountainous and agricultural parts are combustion [64]. Bizkarguenaga and co-workers developed a large-volume-injection-programmed temperature vaporization GC-MS (LVI-PTV-GC-MS) method for the analysis of phthalates, pesticides, PAHs, PCBs, PBDEs, polybrominated biphenyls (PBBs), alkylphenols, and bisphenol A in wastewaters. Water samples were collected upstream and downstream of wastewater treatment facilities in the vicinity of Galindo, Guernica, and Bakio. The samples were extracted by SPE with OAIS-HLB sorbents, and the analytes were eluted using ethyl acetate and n-hexane, respectively. Samples were concentrated and injected in LVI-PTV injection mode, separated on the HP-5MS stationary phase, and detected with MS in scan and selective ion monitoring (SIM) mode with characteristic mass-to-charge ratios (m/z). LOD values for PAH compounds ranged from 1 to 0.028 μg/L. The highest LOD was a B(b)F and was above the MAC prescribed according to the WFD. The PAH concentration in the samples ranged up to 1.427 μg/L [65]. Grmasha and his collaborators determined the concentration of PAH compounds from the WFD list of priority substances in surface water and sediments from the Euphrates river, which flows through Turkey and Syria, eventually merging into the Tigris river in Iraq, forming the Shatt al-Arab waterway that empties into the Persian Gulf [66]. The LOD for water samples ranged from 0.04 to 0.28 × 10⁻³ μg/L, while for sediment samples, it varied from 0.09 × 10⁻³ to 0.65 × 10⁻³ μg/g on a dry weight basis, with recoveries of 90.4 to 100% for water samples and 81.4 to 97.6% for sediment samples. To meet the WFD requirements, 6 L of water underwent SPE, followed by double elution with a dichloromethane/ethyl acetate mixture (1:1, v/v). After combining and evaporating the extracts, they were reconstituted in n-hexane. On the other hand, sediment samples were sieved through a 100-mesh sieve and extracted using an acetone/n-hexane mixture (1:1, v/v) with a microwave digestion system ramping up to 120 °C. Extracts were also evaporated after removing water and reconstituted in n-hexane. Compliance with the WFD was achieved using a GC-MS system in SIM mode with a DB-5MS column, nitrogen as the carrier gas, and a final oven temperature of 300 °C. Their study presents a comprehensive temporal and spatial analysis of PAH compounds in water and sediment samples, revealing higher concentrations of PAH compounds downstream. The benzo(a)anthracene/(benzo(a)anthracene + chrysene) and fluorene/(fluorene + pyrene) ratios indicate the use of petroleum in vehicles and electrical generators, as the measured PAHs are primarily in the 5–6 ring structure group. Creating an effective method for separating closely related PAH compounds based on their structures is challenging since the GC tandem mass spectrometry (GC-MS/MS) method is not very helpful. Barco-Bonila et al. developed a technique to quantify 24 PAH compounds using SPE and SBSE with GC-MS/MS analysis. Traditional columns like HP-5 and HP-5MS are unsuitable due to the coelution of similar PAH compounds, and some PAHs can adhere to these columns, affecting results. A slightly more polar column improves separation. Optimization of SPE extraction found that using a C18 sorbent with added acetonitrile enhances extraction by preventing PAH compounds from sticking to glass surfaces, with dichloromethane as the best elution solvent. SBSE is effective for extracting dissolved PAH and PAH in solid particles using ethyl acetate and cyclohexane. Extraction times increase with PAH size. SBSE works for analyzing all WFD PAH compounds except B(g,h,i)P, with LOD and LOQ values of 0.01 and 0.10 μg/L [28]. In an investigation of the Siverskyi Donets river basin in Ukraine, impacted by military conflicts, various environmental compartments were examined, including groundwater, surface water, sediment, and freshwater fish. Surface water samples were concentrated up to 4000 times using the Atlantic HLB-M Disk and Horizon SPE-DEX 4790 with a 47 mm disk holder. The GC-MS analysis failed to meet the WFD criteria for LOD and LOQ values when it came to pesticides, but it successfully met the standards for PAH. Legacy contaminants were below acceptable levels in surface waters, and no PAH compounds were found. Sediments had a 100% occurrence of PAH, with B(a)P being the dominant one, but pollutant levels were lower due to reduced anthropogenic activity during the conflict compared to neighboring rivers [67].

Santos and colleagues employed the single-drop microextraction (SDME) technique to analyze the levels of polycyclic aromatic hydrocarbons (PAHs) and their derivatives, including nitro-PAHs and quinones, in surface water, seawater, and groundwater samples from the Paraguaçu River in Cachoeira City, Bahia, Brazil. Additionally, seawater samples were gathered from eight locations in Todos os Santos Bay, Brazil, while groundwater samples were obtained from artesian wells at six fuel stations in Salvador City, Bahia, Brazil. The SDME method involves the use of a microsyringe with a needle top with a suspended drop of solvent immiscible with water in contact with the sample. After a certain time, a drop of solvent is withdrawn and redirected to the analysis. Furthermore, it is essential to conduct optimization for SDME, taking into account parameters such as the type of the solvent and volume, extraction duration, stirring duration, and the addition of salt. It was found that the addition of salt can lead to a decrease in extraction efficiency since it slows down the transfer of the analyte from the sample to the solvent. It is assumed that SDME does not depend on the sample volume but only on the concentration of the analyte [68]. In this method, a drop of 1.0 μL, a 10 mL sample, and stirring at 200 rpm were used. Toluene was selected as the solvent, and the extraction was carried out over a 30 min duration. Instrumental analysis was performed using GC-MS with an Rtx-5MS capillary column. LOD ranged from 0.01 μg/L to 0.03 μg/L, and LOQ was 0.04 μg/L to 0.09 for PAH. Although LOD and LOQ are insufficient to determine the concentration of test compounds according to the WFD, an increased concentration of Nap in particular was observed in all water samples. Out of the 52 compounds investigated, 16 were recognized and measured in the examined samples, reaching maximum concentrations of 0.70 μg/L for PAHs, 5.61 μg/L for nitro-PAHs, and 1605 μg/L for quinones [69].

The prep and load (PAL) solution SPME Arrow extraction represents a relatively recent approach that combines elements from both traditional SPME and SBSE methods. Its primary objective is to address the constraints associated with the automation of SBSE and the limited injection volumes of standard SPME. The PAL SPME Arrow comprises a slender metallic rod, which has a polydimethylsiloxane sorbent (PDMS) coating at one end. This coated end features a pointed tip that facilitates efficient penetration of the septum. What sets this method apart from SPME is that, when the rod is pulled into the injector, the tip also functions to seal the injector opening, preventing the introduction of contaminants from the surrounding air. Kremser and co-workers evaluated PAL SPME Arrow in the analysis of 16 PAH compounds using GC-MS with an Rxi-PAH column. The extraction time was set to 70 min. While the PAL SPME Arrow demonstrates strong affinity for PAH possessing a greater number of rings, research has indicated that achieving an LOQ in compliance with EPA and WFD standards is only feasible for naphthalene and acetophenone. The LOQ values for other PAH ranged between 0.1 and 0.8 μg/L [38]. The SPME method has proven its effectiveness in quantifying low pollutant concentrations in water samples, particularly when coupled with GC-HRMS, surpassing traditional GC-ECD and GC-FID methods in terms of resolution and sensitivity. In 2017, Dominguez and colleagues optimized and applied this method for the simultaneous determination of PAH, PBDE, and PCB. Using a PDMS sorbent with the direct immersion injection method, they achieved the best results for extracting a wide range of compounds. Extraction was conducted at 75 °C for approximately 40 min, followed by desorption at 280 °C for 3 min. The samples were separated on a TR-Dioxin-5MS stationary phase and detected with an HRMS detector. Most compounds exhibited well-defined peaks with minimal coelution. The LOQ for individual PAHs ranged from 0.1 × 10⁻³ to 15 × 10⁻³ μg/L, with the highest LOQ of 15 × 10⁻³ μg/L recorded for B(a)P, all of which were compliant with the WFD. Recoveries for all target compounds fell within the range of 84 to 118% [22]. Dobardaran and collaborators determined PAH compound levels in deionized water, tap water, and the Rur river (Germany), stemming from cigarette butt leachate processes. The experiment, conducted in a lab using university cafeteria cigarette butts, employed a GC-MS device with SPME (PDMS) headspace for analysis with a moderately polar stationary phase. LOD fit the WFD requirements in the 0.1 to 0.8 µg/L range, except for Nap, with an LOD and LOQ of 0.4 and 1.2 µg/L, respectively. Naphthalene and Flu exhibited the highest concentrations across all samples. PAH compound relative concentration levels across different water types ranked as follows: Flu > Nap > Ace > Flt > Pyr > B(a)A > Ace > Ant > Phen > B(b)F > B(a)P > B(k)F > Chr. Furthermore, an increase in high-molecular-weight PAH concentrations was noted in later samplings, possibly due to temperature effects, organic substances, and bacterial biosorption. While globally the contribution of cigarette butts to PAH presence in water samples is negligible, at a local level, this effect becomes pronounced [21]. The same group analyzed seawater samples, revealing PAH concentrations ranging from 0.11 to 3.64 µg/L, following a similar trend [23].

Water samples from the vicinity of Langkawi Island (Malaysia) were prepared according to the EPA method using ultrasonic agitation with a dichloromethane solvent. The volume of the extract was reduced with a rotary evaporator and afterward separated with pentane/dichloromethane (2:3, v/v) by column chromatography using a silica gel sorbent. Samples were analyzed using the GC-FID method with fused silica TR-5MS capillary column. The PAH concentration ranged from 6.1 ± 0.43 μg/L to 46 ± 0.42 μg/L. According to the EPA and WFD, all samples showed concentrations higher than allowed. Individual PAH concentrations in waters were in the range of 2.0 × 10⁻² to 2.4 μg/L, with a PAH concentration of 10 μg/L considered highly polluted. B(a)P was detected in all samples, the presence of which undoubtedly indicates an anthropogenic source of PAH compounds. Through correlation factors and diagnostic links of low- and high-molecular-weight PAH, it was found that various sources, including pyrogenic and petrogenic origins, are involved [30]. Laurencik and his collaborators conducted an in-depth analysis of samples of Unio crassus collected from the rivers and streams of Slovakia. They employed a cutting-edge technique known as focused ultrasound solid–liquid extraction (FUSLE), coupled with dispersive SPE for purification. Silica emerged as the optimal sorbent for SPE. The solvent of choice was a carefully selected mixture of n-hexane/toluene (65:35, v/v). Utilizing the GC-MS analysis method with a PAH selective column, SIM analysis was conducted for quantification. This method achieved an LOD ranging from 0.11 × 10⁻³ to 0.34 × 10⁻³ μg/g and an LOQ ranging from 0.036 × 10⁻³ to 0.11 × 10⁻³ μg/g. These values meet the requirements for the analysis of Flu and B(a)P [70]. Previously, a similar methodology was applied to analyze the PAH content in crustacean gammarids from three small lowland streams in Slovakia. The FUSLE process involved the use of a Branson Digital Sonifier SFX 550 with a microtip connected directly to the converter via a coupler. The freeze-dried gammarid samples were spiked, mixed with an extractant solvent comprising n-hexane/ethyl acetate (4:1, v/v), subjected to pulsed sonication, cleaned using SPE (florisil as a sorbent), centrifuged, concentrated, and ultimately prepared for analysis. The method was successfully validated on crustacean gammarid samples. In an alternative method, a 0.2 g freeze-dried gammarid sample was mixed with 2 mL of Milli-Q water and acetonitrile and combined with a salt mixture. After centrifugation, the upper phase underwent triple quadrupole MS/MS (QqQ-MS/MS) analysis, resulting in an LOD of 0.036 × 10⁻³ to 0.17 × 10⁻³ μg/g wet weight and LOQ of 0.12 × 10⁻³ to 0.56 × 10⁻³ μg/g wet weight, meeting WFD requirements. PAH concentrations ranged from 0.12 × 10⁻³ to 0.56 × 10⁻³ μg/g wet weight [71].

3.2. Innovative Techniques for Sample Preparation and PAH Detection: LC, GC, Spectroscopy, and Next-Generation Lab-on-a-Chip Technology

Cyclodextrin silica gels represent a significant technological advancement in adsorbent materials within the domain of analytical chemistry, specifically designed for the extraction of PAH from intricate sample matrices. These gels are functionalized with cyclodextrin, notably β-cyclodextrin, resulting in the introduction of mesoporosity to their surface. This mesoporosity significantly enhances their affinity and selectivity toward hydrophobic PAH molecules. Furthermore, the versatility of mesoporous organic silica extends beyond cyclodextrin modification, encompassing the covalent linkage of diverse organic functional groups to siloxane domains. This molecular engineering approach allows for tailored and precise applications in analytical separations and environmental monitoring. Results with an LOD of 0.0012–0.038 μg/L for PAH compounds in water samples with this type of stationary phases were recorded [112]. The goal of the functionalization of silica gels to certain groups is to improve the extraction efficiency and achieve the lowest possible LOD and LOQ. Silica gels functionalized with macrocyclic tetraazacalix[2]arena[2]triazine provided an LOD in the range of 0.4 × 10⁻³–3.1 × 10⁻³ μg/L for PAH compounds in water samples. Apart from silica gels, nanoparticles made of carbon, metals, graphene, and metal oxides have emerged as highly efficient techniques for extracting PAH from water samples in contemporary methods. The extraction process relies on π–π bonds formed between PAH compounds and the available π electrons on the surface of graphene. Furthermore, the magnetization of graphene nanoparticles has proven highly advantageous as it facilitates the easy retrieval of the sorbent from the solvent. Several factors can impede the extraction efficiency of PAH compounds, including the presence of humic acid, surfactants, and aggregation. Some of these challenges have been mitigated through the introduction of polar functional groups. Nanoparticles based on metals and metal oxides have shown similar efficiency [72,112]. One of the most effective sorbents from this group are titanium oxide, titanium, gold, and silver [72]. A wide array of organic compounds can be employed to coat metal nanoparticles with maximizing recovery rates. Among the commonly utilized options are aliphatic long-chain compounds, alkaloids, sugars, silanes, and aromatic-ring-containing compounds [112].

Currently, molecular imprinted polymers (MIPs) are employed for their remarkable selectivity toward specific groups of PAH compounds. These polymers are created through the polymerization of monomers in the presence of template molecules, subsequently removing the templates, resulting in a structure suited for the separation of molecules resembling the templates. The recognition of molecules occurs based on their size, shape, and π–π interactions with the sorbent. MIP-based SPME and SPE methods are frequently utilized for detecting extremely low concentrations of PAH compounds in water samples. Several adsorbents have been developed that yield satisfactory results in compliance with WFD standards. SPME comprises various adsorbents and adsorbent carriers, with many combinations proving highly effective and in accordance with WFD requirements. Additionally, graphene, graphene oxide, ionic liquids, and divinylbenzene, when combined with fused silica or PDMS, yield comparable results [112]. The optimization of MIPs is mainly based on the combination of certain template molecules, crosslinkers, and monomers, in different ratios [113]. Innovative approaches for developing nanostructured polymers involve enhancing sorbents through the introduction of polymeric materials to create numerous active sites for interaction with specific polymer types, such as hydroxyl, amino, or carboxyl groups [114,115]. A pivotal challenge for these emerging techniques is the transition from experimental stages to commercial application. Notably, magnetic poly (β-cyclodextrin-ionic liquid) nanocomposites have demonstrated remarkable selectivity in the analysis of PAH compounds across diverse environmental samples [116]. Additionally, Arias and collaborators harnessed novel imprinted polymer-based sorbents for SPE, coupled with LC ultraviolet (LC-UV) analysis, revealing exceptional selectivity for organophosphoric compounds such as diazinon. This adsorbent has also found utility in other extraction methods such as SPME and SBSE [25].

Lake Maggiore’s surface water samples were analyzed to detect PAH and assess the Ticino river’s impact on the lake water quality. PAHs were extracted using XAD-2 resins and ASE with n-hexane as the solvent, then analyzed via GC-MS with an SGE 25QC3/HT8 column. Notable concentrations of dissolved PAH in surface waters included Ace (1.1 × 10⁻³ ± 0.89 × 10⁻³ μg/L), Flt (0.79 × 10⁻³ ± 0.85 × 10⁻³ μg/L), and Phen (0.78 × 10⁻³ ± 0.44 × 10⁻³ μg/L), while rainwater showed Flu (0.067 μg/L) and Pyr (0.075 μg/L). Key solid-phase PAHs in surface waters were Flu (0.12 × 10⁻³ ± 0.11 × 10⁻³ μg/L), Phen (0.075 × 10⁻³ ± 0.081 × 10⁻³ μg/L), and I(1,2,3-c,d)P (0.078 × 10⁻³ ± 0.14 × 10⁻³ μg/L), whereas rainwater featured Pyr (7.45 × 10⁻³ ± 6.3 × 10⁻³ μg/L), B(a)Ant (18.4 × 10⁻³ ± 4.4 × 10⁻³ μg/L), and Flt (17.2 × 10⁻³ ± 2.5 × 10⁻³ μg/L). Total PAH concentrations in surface water were 2.924 ± 0.312 μg/L (solution phase) and 0.584 ± 0.033 × 10⁻³ μg/L (particulate phase), while rainwater showed 0.0754 ± 0.009 μg/L (solution phase) and 0.0275 ± 0.002 μg/L (particulate phase) [73]. Seasonal analyses showed an increase in the concentration of PAH in surface water as well as in rainwater during the winter period (especially November and December). The reason is most likely increased emissions due to intensive consumption of heating media in households. This study was performed before the adoption of the WFD in 2013; however, the concentrations of PAH compounds from the WFD are above the allowable value. Montuori and co-workers investigated the concentration of 16 PAH compounds according to the EPA standard in sediment and Sarno river (Italy) samples. Sampling was conducted across all seasons in both the lower and upper river basins. To isolate the suspended matter, water samples underwent a filtration process. PAHs from the suspended matter were extracted by ultrasonication with methanol/dichloromethane (1:1, v/v) and separated into three fractions by column chromatography. The first three fractions were eluted with n-hexane, n-hexane/ethyl lactate (1:1, v/v), and ethyl lactate, respectively. The filtrates were concentrated using the SPE method with Strata XTM sorbent, and the second fraction (PAH) was eluted with the solvent n-hexane/ethyl lactate (1:1, v/v). PAHs were extracted from sediments using column chromatography, employing the same methodology as for suspended matter. Gas chromatographic analysis with an MS detector was conducted using an SPB 20 chromatographic column. LOD values ranged from 0.01 to 0.1 × 10⁻³ μg/L, and the LOQ ranged from 0.03 × 10⁻³ to 0.2 × 10⁻³ μg/L. PAH concentrations in water samples ranged from 12.4 × 10⁻³ to 2321 × 10⁻³ μg/L, with the highest levels observed in the filtered portion of water samples, while sediment samples exhibited the lowest concentrations. Water samples were predominantly composed of PAH with two and three rings, while sediments contained predominantly high-ring PAH. Suspended matter contained both low-ring and high-ring PAHs. The sampling locations significantly influenced PAH content, with higher concentrations in lower river flows and potential variations attributed to varying weather conditions and seasons. The source of PAH compounds in the Sarno river was determined using specific PAH ratios, such as the penantrene/atracene ratio, indicating incomplete combustion and the presence of numerous polluting factories in the vicinity [74]. Increased concentrations of PAH have been reported during the winter, especially in areas where wood and coal are used as heating agents in households [73,117,118]. One possible reason for this occurrence might be the more efficient formation of PAH particles in colder conditions, which then bind to sediment or enter water systems and dissolve more readily during summer rains. Meng and co-workers conducted an extensive study of the composition and concentration of 16 EPA PAH compounds in 14 lake water and sediment samples across China. The concentration of individual PAHs in water samples ranged from 0.316 to 6.525 μg/L, while the maximum determined concentration of total PAH was 12.971 μg/L. The concentrations of Nap, Ant, Phen, and Flu exceeded the limits set by the EPA and WFD. In general, accelerated urbanization, a dense population, and intensive use of wood and coal as heating agents have led to increased PAH content in the analyzed samples [117]. PAH compounds may find their way into water systems during rainfall. When it rains, residues from vehicle emissions, lubricants, petroleum products from gas stations, and road coatings are carried off and subsequently either directly enter rivers or become mixed with wastewater. Aragon and co-workers used the through oven transfer adsorption–desorption (TOTAD) interface for the LC-GC analysis of PAH in water samples. This method allows a reverse and normal phase system for the analysis of large sample volumes including direct water sampling. In combination with the Quadrex column, LOD values were below the MAC, and most PAH compounds in the WFD were achieved [26].

In addition to the GC method, the most commonly used methods for PAH analysis in water samples are LC methods. Kumar et al. developed a high-performance LC (HPLC) technique with a SupelcosilTM column and an Eclipse XDB-C8 guard column for PAH analysis. They used dichloromethane for LLE sample extraction and a dichloromethane column with silica gel and sodium sulfate. The LOD ranged from 0.01 to 0.51 μg/L, with the LOQ from 0.03 to 1.72 μg/L. This method did not meet 2013 WFD specificity, except for Nap with the LOD and LOQ at 0.12 μg/L and 0.41 μg/L, respectively. Recoveries for PAH ranged from 78% (±2.23%) to 100% (±12.27%) [41]. Siemers et al. developed a method for the analysis of PAH, alkyl PAH, heterocyclic PAH (NSO-PAH), and phenol. Their research was divided into three distinct phases. Firstly, they conducted an analysis of water samples collected from Lower Saxony, specifically the Elba, Ems, and Weser rivers. Secondly, they gathered additional samples from the Elbe and Weser rivers. Lastly, the researchers collected seven samples from the North Sea. Phenol determination was carried out using the LLE method with derivatization, and SPE was performed using Lichorult EN cartridges. The analysis involved the use of GC-MS and LC-MS/MS techniques, resulting in LOQ values ranging from 0.4 × 10⁻³ to 2.4 × 10⁻³ μg/L, demonstrating the method’s applicability for PAH determination in accordance with the WFD standard. In the initial sampling round, total PAH compound concentrations ranged from 13 × 10⁻³ to 140 × 10⁻³ μg/L, while in the second round, they ranged from 10 × 10⁻³ to 40 × 10⁻³ μg/L. For the North Sea samples, PAH compound concentrations were below the LOQ and estimated to be in the range of several ng/L. Notably, the researchers observed concentrations exceeding the permitted limit for I(1,2,3-c,d)P in six samples during the first part of their study [29]. Sarria-Villa and co-workers examined the content of PAH compounds in sediments and water samples from the Cauca river, in Colombia, which stretches for 1350 km. The extraction of PAH from water samples involved utilizing the SPE method with BakerBond cartridges, followed by elution using a dichloromethane solvent in accordance with the USEPA 550.1 protocol. For sediment samples, PAH compounds were extracted using the USEPA 3550 B method, employing a solution composed of acetone, dichloromethane, and n-hexane in a 1:1:1 volume ratio. Analysis was carried out through HPLC with a C18 column paired with both fluorescent and UV detectors, as well as a GC method utilizing XTI-5 with a stationary phase coupled to an MS detector. The LOD ranged from 0.0016 to 0.1018 μg/L for water and 0.0004 to 0.0265 μg/g for sediment. Higher concentrations of PAH compounds were detected in the lower reaches of the river, likely attributable to precipitation and flooding events. Fluorene and Acy exhibited the highest concentrations in aqueous samples, while B(a)P was not detected in any water sample. Unlike sediment samples, aqueous systems showed increased concentrations of lower PAH compounds, facilitated by the favorable tropical conditions that enhance their solubility in water. It was found that increased concentrations of PAH compounds in individual samples, particularly downstream samples, were of both pyrogenic and petrogenic origin [75].

Fluorescence spectroscopy plays an important role in the analysis of PAH and pesticides, capitalizing on their unique fluorescent properties within the UV and blue spectral regions, as implemented in systems like AQUAMS©, CONTROS©, and HACH©. In a study conducted by Ferretto et al., PAH compounds were efficiently extracted from marine samples using the SPE method, employing C18 cartridges and dichloromethane as the solvent. Their analytical method encompassed a spectrophotometer and GC-MS for the identification of individual PAH. The research aimed to subject both methods to scrutiny using the parallel factor analysis (PARAFAC) algorithm, revealing a strong linear correlation between them. However, it is important to note that fluorescence-based PAH identification relies primarily on the presence of alkyl substituents, rendering it suitable for screening analysis rather than a standard method. Fluorescence spectroscopy achieved higher LODs for PAH compounds compared to the maximum allowable concentration stipulated by the 2013 WFD, with Nap as the exception, which displayed a low LOD of 1.29 μg/L compared to its relatively higher allowable concentration [76].

Foan and co-workers used an off-line extraction method with a microfluidic device based on lab-on-a-chip technology. The operation of this device is based on a preconcentration procedure consisting of the extraction of the analyte from the matrix and then desorption by a solvent. The microfluidic chip was functionalized with PDMS in combination with an OV-1 sorbent for PAH extraction. Extraction and desorption were achieved using a syringe pump. The outcomes obtained through this approach were juxtaposed with the results derived from the optimized SBSE technique employing a PDMS sorbent and the HPLC with fluorescence detection (HPLC-FLD) method employing an LC-PAH C18 column. It was observed that acetonitrile emerged as the most suitable solvent for microfluidic extraction. When considering the recovery of low-molecular-weight PAH, SBSE demonstrated superiority, primarily attributed to the increased thickness of its stationary phase, resulting in recoveries exceeding 90%. However, both methods exhibited reduced extraction efficiency in the presence of dissolved organic matter. Worth noting is that the LOQ achieved by HPLC-FLD for B(a)P, B(k)F, Ant, and Nap stood at 0.03 × 10⁻³ μg/L, 0.01 × 10⁻³ μg/L, 0.04 × 10⁻³ μg/L, and 1.0 × 10⁻³ μg/L, respectively, all of which are below their MAC as stipulated in the WFD. On the other hand, microfluidic extraction showed better recovery and reproducibility for higher PAH; however, the LOQs were in the range of 0.4 × 10⁻³ u to 29 × 10⁻³ μg/L for B(k)F and I(1,2,3-c,d)P [42]. In addition to the PDMS adsorbent, the group developed a microfluidic device method that uses nanoporous organosilicate (SiOCH) to extract PAH compounds from water samples. Recoveries of lower PAH compounds were improved with the introduction of a SiOCH sorbent. The aim of this research was to produce devices that will be integrated into systems for direct measurement of organic pollutants [42,77].

Emerging electroanalytical techniques are increasingly employed to analyze trace concentrations of priority substances [78,79,80,81]. These methods primarily enable the real-time measurement of various pollutants on-site, minimizing potential errors associated with sample collection, transportation, and preparation. Electroactive receptor-based systems have been innovatively devised to scrutinize specific PAH compounds, capitalizing on the redox activity exhibited by PAH via non-covalent and π–π bonds. To detect tricyclic aromatic hydrocarbons, specific molecules like thionine can be incorporated into graphene structures. Conversely, certain molecular configurations can be electrodeposited onto electrode surfaces to render them tailored for a particular PAH type. In addition to these advancements, biological redox systems have been harnessed, involving processes such as DNA electrodeposition on electrode surfaces for the assessment of PAH compounds. Electrochemical immunosensors have emerged as a highly effective means for detecting PAH such as B(a)P, offering exceptional selectivity and sensitivity. However, it is crucial to note that these immunosensors possess limitations related to their operational pH range, temperature, and time stability, as well as incubation duration [119]. In 2021, Comnea-Stancu and their research team conducted an extensive study, examining electrochemical devices suitable for monitoring the entire range of priority pollutants outlined in the WFD, such as PAH and their derivatives [120]. Additionally, novel sample preparation techniques utilizing electric fields, particularly electromembrane extraction, have gained attention for expediting the mass transfer of ionic analytes during extraction. Introduced in 2006, electromembrane extraction applies an electric field across two compartments separated by a membrane, enabling analyte migration and offering advantages such as reduced acceptor solution volume, matrix compound cleanup, and selectivity tuning. Electromigrative separation techniques, exemplified by capillary electrophoresis, provide an alternative to ion chromatography for analyzing ionic micropollutants such as pesticides. However, challenges persist in using capillary electrophoresis for very polar ionic compounds in environmental samples, necessitating improvements in sample preparation methods such as electromembrane extraction. While optimized capillary electrophoresis with diode array detection achieves medium ng/L detection limits, MS remains the preferred method for non-target screenings, especially in research laboratories, owing to fewer restrictions compared to ion chromatography-mass spectrometry hyphenation and its environmentally friendly attributes with low solvent consumption. In discussions and tables, references may be numbered sequentially [121].

4. Assessing Pesticide Levels and Techniques for Their Detection

Approximately 2.5 million tons of pesticides are consumed globally each year, with their production and usage increasing significantly [3,82]. Similar to PAH, the fate of pesticides in water depends on the physicochemical characteristics of the environment and their solubility in water. The pollution of water sources with pesticides is affected by various factors such as farming methods, application rates and frequencies, seasonal variations, trophic conditions, product quantities, soil characteristics, hydrogeology, and climatic conditions. The byproducts formed during their degradation or transformation may exert a greater influence on water quality than the original pesticides [2]. Pesticides are applied to the soil to eliminate unwanted organisms and protect agricultural produce [122]. Although OCPs exhibit high environmental toxicity, they typically exist in concentrations below the LOD for specific analytical techniques. The identification of these low concentrations has spurred the creation of highly sensitive and precise methods designed for discerning specific substances or categories of diverse pesticides [123,124,125,126]. In the quest for a reliable pesticide detection method, there is an exploration of new approaches. One such method involves the utilization of short single-stranded oligonucleotides (RNA and DNA) known as aptamers, which selectively bind to the target molecule [127]. In contrast to biosensors, aptamers offer a more cost-effective and straightforward alternative due to their lower molecular weight and increased stability. Various types of aptasensors, including those for electrochemical, fluorescent, colorimetric, and chemiluminescent detection and aptasensors based on Raman spectroscopy, have been created. Each of these approaches comes with its own set of advantages and drawbacks. Overcoming interfering effects is a key challenge in implementing new analytical methods [83,127]. Silica nanoparticles offer a contemporary and innovative approach for both qualitative and quantitative pesticide analysis, as elaborated in the section concerning PAH pollutants. Silica nanoparticles can be effectively paired with metals such as gold, silver, and iron (in the form of Fe₃O₄) to customize their affinity for specific types of pesticides. In such cases, the magnetic properties of these particles have proven valuable for isolating the analyte from the sorbent material. Conversely, silica nanoparticles linked to silanes, amines, and ionic liquids have demonstrated high effectiveness in simultaneously extracting organophosphorus pesticides, OCPs, and pyrethroids. Molecularly imprinted polymers (MIPs) represent another category of functional components integrated with silica nanoparticles, proving valuable in identifying specific pollutants due to their specialized surface areas tailored to the precise structure of the analyte. In the realm of QuEChERS methods involving nanoparticles, graphene nanostructures are employed both independently and in conjunction with the previously mentioned functional groups. Additionally, pesticides can be detected through photochemical means by using nanoparticles modified with molecules capable of generating chemiluminescent or fluorescent signals upon interaction. Modifiers utilized for these purposes include Zn(II), Eu(III), and CdTe quantum dots (for pyrethroid determination), achieving a limit of detection (LOD) as low as 164.582 μg/L [84]. So far, a limited number of experiments have been performed, and most methods have proved satisfactory for the determination of different types of pesticides. The sensitivity of the method and the LOD are usually around a few μg/L or less than 1 μg/L, which makes these methods useful for determining most persistent pollutants in accordance with EPA and WFD standards [128]. An interesting review of current trends in the detection of persistent pollutants based on MIPs with ionic liquids was performed by Liu and co-workers [129]. Hanedar and colleagues conducted an examination of pollutant contents in the Küçük Menderes Basin, which serves as the source of water flowing into the Aegean Sea through the Küçük Menderes river [85]. Sampling was conducted at 17 water bodies to ascertain the presence of all 45 primary substances, evaluate the chemical status of these water bodies, and identify pollution sources. Additionally, three samples were collected from lakes. The extracts obtained were subjected to analysis using the GC-MS method in accordance with the 8270 D and EPA 8270 methods. Based on the results obtained, it can be concluded that the method satisfies all necessary requirements for analyzing compounds with very low MAC values according to the WFD. The analyzed pesticides, along with Nap, Flu, Ant, and B(a)P, were detected in concentrations exceeding the permitted limits. Specifically, the concentrations were 0.004 μg/L for cypermethrin, 0.126–1.186 μg/L for dichlorvos, 2.482–58.456 μg/L for Nap, 0.190–12.522 μg/L for Ant, 0.046–9.367 μg/L for Flt, and 0.102–10.409 μg/L for B(a)P. A comparable study was undertaken by monitoring 45 priority substances outlined in the WFD along with 250 pollutants specified in the Regulation on Surface Water Quality. A total of 346 samples were collected from 53 locations along the Yesilirmak River in Turkey over a span of two years. The analysis of PAH and the four pesticides under investigation was carried out using the GC-MS/MS method, employing a DB-5MS non-polar column [86]. Dichlorvos was found in 56 samples, exceeding the 36 × 10⁻⁴ μg/L annual average environmental quality standards (AA-EQS) limit. Cypermethrin concentrations ranged from 0.031 to 0.375 μg/L, detected in 32 samples. Two years prior, a study on Sakarya River water samples in Turkey’s Marmara province, utilizing 100 mL water for extraction, determined LOD values for various substances, revealing analytical sensitivity. The study indicated that PAHs were the primary pollutants, hinting at potential industrial activities, while pesticides and phenol compounds likely originated from agriculture and soil runoff [87].

GC×GC techniques have gained increasing popularity in analyzing complex samples [88,130]. In contrast to conventional GC, GC×GC offers improved peak capacities, yields a more distinct and comprehensive chromatogram, enhances resolution, and improves the capability for group-type analysis. This method involves interconnecting two columns with distinct separation mechanisms through a modulator. Techniques for assessing pesticides using GC×GC analysis have already been developed and implemented. Variations in the choice of columns, modulators, and detectors enable efficient analysis of different sample types and the determination of numerous compounds. GC×GC has demonstrated effectiveness in both target and non-target analysis [2]. When combined with high-resolution time-of-flight mass spectrometry (HRTOF-MS), it becomes an exceptionally potent technique since this detector possesses the capability to precisely determine the masses of resulting ionic fragments. Recent advancements regarding the HRMS are centered on an orbitrap analyzer, which enhances quantification performance, provides higher resolution capabilities, and extends the linear range [2,88,130]. Table 3 provides a summary of analytical techniques frequently utilized to detect pesticides and other contaminants in water samples and various types of specimens. Moreover, the table displays the LOD and LOQ values attained through the mentioned methods.

4.1. Cypermethrin

The reason why pyroetroids have found great application in the last few decades (in households, public health institutions, food industries, agriculture, and horticulture) lies in the fact that they are significantly less toxic compared to other commercial pesticides [18,122]. Pyrethroids share a structural similarity with naturally occurring pyrethrins. Cypermethrin is a commercial insecticide belonging to the group of synthetic pyrethroids with a cyano group in the molecule [89]. It is often used in water purification systems in order to remove any insect larvae that may have formed. Cypermethrin occurs in α and β form, each of which has four enantiomers, while only two have insecticidal activity (R-cis-αS and 1R-trans-αS). The different cypermethrin stereoisomers are classified into different subgroups: α-cypermethrin (two cis isomers: 1R-cis-αS and 1S-cis-αR), β-cypermethrin (two cis isomers 1R-cis-αS and 1S-cis-αR and two trans isomers 1R-trans-αR and 1R-trans-αS), θ-cypermethrin (two trans isomers: 1R-trans-αS and 1S-trans-αR), and ζ-cypermethrin (four αS isomers: 1R-cis-αS, 1R-trans-αS, 1S-cis-αS, and 1S-trans-αS). The WFD refers to cypermethrin as a mixture of all isomers. The maximum permissible concentration of cypermethrin in inland surface waters and other waters is 6 × 10⁻⁴ and 6 × 10⁻⁵ μg/L, respectively. Given the potential for cypermethrin accumulation in aquatic organisms, such as shellfish, it is feasible that specific concentrations could induce oxidative stress. This can be assessed by analyzing biological markers in samples from these organisms [131]. Various methods exist for detecting such low concentrations of pollutants, with the most prevalent approach being the utilization of passive samplers or the extraction of substantial sample volumes [132]. The concentration of cypermethrin in water is influenced by factors such as electrical conductivity, pH, and nitrite concentration. It has been demonstrated that cypermethrin exhibits greater resistance to photooxidation processes in aquatic environments compared to soil environments [82]. Additionally, its degradation is directly proportional to the heavy metal content in the soil. Cypermethrin tends to aggregate in aqueous media and settle within sediment layers.

Cypermethrin: Chromatographic Methods

Rosch et al. developed a method for determining OCPs and pyrethroids that satisfies low MAC requirements. They used the GC-MS/MS method with atmospheric-pressure chemical ionization with an Rtx-5MS capillary column to analyze extracts of water samples from a small source in the Swiss Plateau. The LLE method used involved stirring a 200 mL water sample with 5 mL of n-hexane, followed by evaporating the extract for further analysis. Water was used as a modifier for chemical ionization, and it was found to produce the highest peak intensities, unaffected by air humidity. In this analysis, the intra-day recovery and absolute recovery were determined, yielding values of 93 ± 13% and 89 ± 1%, respectively. The method also established both the instrumental LOQ and the method LOQ, which were measured at 0.25 × 10⁻⁸ μg/inj and 0.125 × 10⁻⁴ μg/L, respectively. It is important to note that the method LOQ represents the concentration of a standard that produces a peak with a signal-to-noise ratio (S/N) of at least 10 for the quantifier ion and 3 for the qualifier ion. In environmental analysis, cypermethrin, along with λ-cyhalothrin and permethrin, was detected in Swiss river samples. These pyrethroids were found in 19 out of 58 samples, and in 10 of them, their concentrations exceeded the MAC according to the WFD. The average concentration of cypermethrin in these cases was 0.980 × 10⁻³ μg/L [39]. A comparative study to assess the performance of two analytical techniques, specifically, GC with an ECD detector and HPLC with a diode-array detector (DAD), was conducted for quantifying pyrethroid pesticides in water samples. The results indicated that the HPLC method outperformed GC in terms of speed, achieving a significantly faster analysis time, and it also exhibited a lower LOD for detecting all pyrethroid compounds. However, it is worth noting that having a low LOD alone does not necessarily make the method compliant with the 2013 WFD standards [90]. In recent years, iron-based magnetic composite materials have been used in the preliminary extraction of pyrethroids (including cypermethrin) from certain samples. They showed high efficiency and good recovery in the range of 93.0–104%. This method is relatively fast and efficient and can be used for large sample volumes. Furthermore, the introduction of salt has been demonstrated to adversely impact the effectiveness of the extraction process [133].

Furthermore, the utilization of passive samplers offers a more accurate assessment of the influence of organic matter on aquatic organisms [134]. Bhattacharjee and colleagues employed rice bran as a means to eliminate cypermethrin from water [91]. In another study, the content of cypermethrin in Lisungwa, Neno, Southern Malawi, was determined. Although these measurements were conducted following the 2011 World Health Organization standards, where the MAC is higher than that stipulated by the WFD, it was observed that the concentration of cypermethrin exceeded the permissible limits. The analysis was carried out using HPLC, and an increase in concentration was also noted during rainy seasons [82].

4.2. Heptachlor and Heptachlor Epoxide

Even though OCPs were banned many years ago, their lasting impact is still evident today due to their persistence primarily due to their high stability and lipophilic character. Furthermore, the continued presence of OCPs like heptachlor and its isomers, lindane, aldrin, endosulfan, and others suggests that these chemicals are still being used despite strict regulations against them. The concept of “shock concentrations” during the summer months, especially in agricultural and livestock industries, supports this observation. Conversely, the utilization of certain pesticides, such as lindane in sugar production, results in elevated concentrations during the winter season. Besides agricultural activities, OCPs make their way into water systems through processes like sediment irrigation during rainy seasons and deliberate discharge from storage facilities. These compounds are known for their ability to accumulate in the tissues of aquatic organisms due to their specific properties [18]. Heptachlor exhibits an environmental half-life of approximately 14 years. During its decomposition, the primary degradation product is heptachlor epoxide, existing in both exo- and endoisomers, with the endoisomer demonstrating greater stability. Given the increased stability of heptachlor epoxide in comparison to heptachlor itself, it becomes essential to consider the concentrations of both compounds in the environment. The chromatographic separation of both isomers presents a considerable challenge. Nevertheless, it is possible to distinguish them by utilizing a combination of various quantification ions (353 for exo and 253 for endo heptachlor epoxide), characteristic ions (81, 353, 355, 351, and 357 for exoisomer and 81, 253, and 185 for endoisomer), and fragmentation ions (237 + 81 for exoisomer and 183 + 217 + 135 for endoisomer). McManus et al. used the Zebron ZB-5 column to separate heptachlor and heptachlor epoxide isomers [92]. While there is insufficient evidence to definitively establish the harmful impact of heptachlor and heptachlor epoxide, it is worth noting that these substances are included in the list of priority compounds outlined in the WFD [19]. Additionally, their ability to penetrate deep into the soil can facilitate their movement into groundwater. When they enter aquatic organisms or humans, they can cause endocrine system disorders [93]. The most significant migration of pesticides occurs when they evaporate from the water surface or sediments and are transported through the atmosphere. This process results in the re-suspension of pesticides, allowing them to reach water bodies and sediments, creating a cyclic pattern [94]. Several techniques have been suggested for the elimination of OCPs from the environment, with a primary focus on oxidizing them using metal oxides such as TiO₂ nanoparticles, Fe₂O₃, and Fe₃O₄ (magnetite) [19]. The practical utility of such techniques is enabled by the magnetic properties of the nanoparticles. Furthermore, promising research has been conducted on the oxidation of pesticides using elemental iron. Alongside metals and metal oxides, a prevalent approach for oxidizing and eliminating OCPs involves utilizing combinations of methods such as H₂O₂, UV, and O₃ treatment. It is important to note that degradation products resulting from these processes can pose significant risks to aquatic organisms, particularly due to the presence of highly reactive OH radicals [19].

Heptachlor and Heptachlor Epoxide Analysis: Gas Chromatography Insights

The dispersive LLE (DLLE) method is a scaled-down adaptation of LLE, characterized by reduced solvent volume and a shorter extraction duration. This method requires the use of solvents that are partially soluble in water or solvents with high melting points. Carvalho and colleagues were able to achieve an LOD of 0.003 μg/L for aqueous samples by employing the GC analysis with the DLLE method. This DLLE variant is more cost-effective and environmentally friendly as it reduces solvent volume compared to LLE [95]. The technique involves freezing and collecting the solvent post-extraction, followed by melting it before analysis. Separation and detection were carried out using GC-MS with a TR-5MS capillary column. Tsai and Huang, on the other hand, attained LOD and LOQ values of 0.001 and 0.0004 μg/L, respectively, for heptachlor with GC-MS analysis in combination with DLLE [96]. Tankiewicz and Biziuk optimized the DLLE method for the determination of 34 pesticides from various categories [97]. In addition, liquid-phase microextraction (LPME) has been developed for OCPs, which has gained special attention since it requires reduced exposure to toxic solvents. The most well-known form of LPME extraction is homogeneous liquid–liquid microextraction (HLLME), which is environmentally and economically much more efficient than the previous ones [32]. Dispersive micro-SPE (DMSPE), a green-chemistry-compatible method for extracting pollutants from water, utilizes a small sorbent amount dispersed in the sample. After extraction, the sorbent is removed through filtration, centrifugation, or a magnetic field and then extracted with a suitable solvent. Although Nascimento et al. developed a DMSPE technique for 39 pesticides, the analysis did not meet WFD requirements due to high LOD values for heptachlor and heptachlor epoxide [98]. In recent years, there has been notable interest in the use of carbon nanotubes for SPE extraction of OCPs and organophosphate pesticides (OPPs) [93,99]. Chowdhury et al. conducted research across 25 districts in Bangladesh, assessing OCP content, including heptachlor, in water samples achieving WFD MAC criteria. The findings revealed that heptachlor exceeded the established LOD of 0.3 μg/L at four locations [100].

Dias et al. used the SPME method of extracting the pesticides from water samples and compared the efficiency of the DVB/CAR/PDMS sorbent, with a sorbent made of crushed cork used to close wine bottles. The cork sorbent exhibited its highest extraction efficiency when operated at 75 °C for 60 min with the addition of NaCl, whereas the DVB/CAR/PDMS sorbent demonstrated its optimal efficiency under the same 60 min extraction time but at a lower temperature of 60 °C without the need for NaCl. Utilizing the GC-ECD analytical method with a Zebron ZB-5MS chromatographic column, consistent results were obtained for both sorbents in the extraction of heptachlor and heptachlor epoxide. The LOD and LOQ for heptachlor were 0.0008 and 0.0025 μg/L, respectively, while the LOD and LOQ for heptachlor epoxide were 0.0003 and 0.001 μg/L, respectively [24].

A study explored the effects of sorbent volume (24 μL and 47 μL) on extraction efficiency using the SBSE [43]. In the SBSE process, a PDMSE sorbent optimized at 1500 rpm for three cycles with a 50 mL water sample demonstrated enhanced extraction efficiency, aligning with the principle that increased solvent and sorbent volumes boost SBSE performance. GC×GC-HRTOF-MS analysis employed DB-5 and BPX-50 capillary columns in primary and secondary dimensions, linked to a cryogenic thermal modulator. Analyte desorption utilized a temperature program from 40 °C (held for 0.5 min) to 280 °C (held for 5 min) at a flow rate of 50 mL/min. GC-MS and GC-HRMS analyses used DB-5 and DB-17HT capillary columns. Salt addition, while generally improving extraction for low or medium K_OW compounds, reduced efficiency for hydrophobic ones, leading to its omission. Cis- and trans-heptachlor epoxide target ions in water at 1 × 10⁻³ μg/L had insufficient selectivity, but with LODs of 23 × 10⁻⁶ and 22 × 10⁻⁶ μg/L, respectively, the method met WFD standards for water characterization. Ayase river samples successfully detected seven OCPs, including cis-heptachlor epoxide, without interference. The GC×GC-HRTOF-MS method determined cis-heptachlor epoxide concentration at 81 × 10⁻⁶ μg/L, achieving effective separation in the second dimension despite some coelution. GC-HRMS analysis, with larger sample volumes (20 L) and injection volumes, identified 20 OCPs. Non-target analysis using GC×GC revealed an additional 20 compounds from diverse groups. The GC-HRMS method determined cis-heptachlor epoxide concentration at 180 × 10⁻⁶ μg/L.

The Sarno River, acknowledged as Europe’s most polluted river, was analyzed for PCB and OCP concentrations in both water and sediment samples in Italy. The river, covering approximately 715 km², is surrounded by fertile land influenced by nearby volcanic activity, leading to heightened pesticide use. Water samples underwent GF/F glass fiber filter filtration, followed by the extraction of suspended solids using (dichloromethane/methanol). Column chromatography with aluminum separated the fractions, with subsequent analysis using the GC-ECD method. The LOD values for individual PCBs and OCPs ranged from 0.005 × 10⁻³ to 0.044 × 10⁻³ μg/L in the particulate phase, 0.004 × 10⁻³ to 0.110 × 10⁻³ μg/L in the dissolved phase, and from 0.10 × 10⁻³ to 4.267 × 10⁻³ μg/L in suspended solid matter. The concentration of heptachlor and heptachlor epoxide ranged from 0.14 × 10⁻³ to 1.49 × 10⁻³ μg/L. The study revealed higher OCP concentrations in sediments compared to the river, suggesting an ongoing sedimentation process and no fresh pesticide intake. Higher pesticide concentrations in the lower reaches indicated leaching during passage through agricultural areas. Alarming statistics included a daily intake of 326 g OCPs and 948 g PCBs in the Tyrrhenian Sea [36].

4.3. Dichlorvos

Dichlorvos (2,2-dichlorovinyl dimethyl phosphate or DDVP) is a chlorinated OPP used as an insecticide in agriculture, in the workplace, in households, and in the treatment of livestock and domestic animals. It is toxic to humans since acute exposure leads to respiratory problems, coma, or death [122]. Various aquatic organisms exhibit different levels of susceptibility to specific concentrations of dichlorvos in water systems [101]. Long-term exposure to dichlorvos is associated with the development of several diseases, including diabetes and cancer, which is why it is categorized as a priority substance. Based on available data, approximately 2 million people lose their lives each year due to OPP poisoning [102]. Cruz-Alcalde and Esplugas elucidated the degradation pathway of dichlorvos when it undergoes removal via ozonation or the application of hydroxyl radicals [20]. Dichlorvos exposure primarily occurs through water consumption but can also happen via contaminated air. In Bangladesh and India, where OPPs are extensively used, there has been a notable rise in cases of various health issues, including cancer, liver diseases, nervous system disorders, kidney problems, infertility, and more [135]. Similar to all OPPs, dichlorvos exhibits a lipophilic nature and has a tendency to accumulate in tissues such as the liver, brain, kidneys, and spleen. In vitro studies suggest an average blood half-life of approximately 10 min. The largest percentage of accumulated dichlorvos is metabolized in the liver by hydrolysis and demethylation and excreted in the urine or exhalation from the body [135]. Moreover, it is believed to play a role in inhibiting the enzyme acetylcholine esterase, which is crucial for transmitting impulses in the body and can lead to respiratory issues. It is also suspected to contribute to heart problems, although this has not been definitively established. There have been numerous reports of both antagonistic and synergistic effects between pesticides and metals or metalloids (such as mercury, lead, and arsenic) and their respective compounds [135]. Methods have been developed for assessing OPPs, including dichlorvos, yet a considerable portion of these methods fails to meet the WFD requirements regarding the LOD and LOQ [103]. Organophosphorus pesticides exhibit higher lipophilicity in comparison to OCPs, which are characterized by greater stability and resilience. Given the significantly lower MAC for dichlorvos compared to other OPPs, the establishment of a detection method with an LOD lower than the MAC proves to be quite challenging [37,104]. An option for detecting organothiophosphorus pesticides involves utilizing a chemiluminescent detector in conjunction with HPLC [105].

4.3.1. Dichlorvos Analysis: Biosensor and Nanotech Approaches

The OPPs and OCPs are most often determined in tandem. Ionic liquid-based solvents may serve as alternatives to traditional volatile solvents in pesticide extraction, owing to their favorable attributes. These solvents are characterized by reduced toxicity, lower volatility, enhanced dissolution capabilities for both organic and inorganic compounds, and a broader temperature range in which they remain in a liquid state [83]. Apart from chromatogrpahic techniques for analyzing and detecting dichlorvos, there have been advancements in spectrophotometric methods that rely on specific organic reagents reacting with the hydrolysis products of dichlorvos. Developing gold nanoparticles for colorimetric testing poses a significant challenge in the detection of OPPs, with recovery rates in this method ranging from 90.0 to 101.2% [106]. Immunoassay-based detection of OPPs has also proven to be a reliable approach, primarily utilizing enzymes such as acetylcholine esterase and organophosphate hydrolase, with acetylcholine esterase being a crucial enzyme involved in nerve impulse transmission. Acetylcholine chloride can undergo hydrolysis when it reacts with the enzyme acetylcholine esterase, resulting in the formation of choline. Choline can then be oxidized in the presence of hexcyanoferrate. However, when dichlorvos or another OPP is present, this hydrolysis reaction is hindered, as observed through amperometry. The method described above has an LOD of 4 μg/L for dichlorvos, which is insufficient for its application in the context of WFD [107]. Despite their sensitivity, these methods are predominantly time-consuming and costly and require a complex experimental setup. The initial exploration of biosensors for dichlorvos detection began in the 1990s, but at that time, the lowest LOD was approximately 0.0155 μg/L. Mishra and co-workers wrote an extensive review covering three stages in the development of biosensors for dichlorvos detection [136]. The initial stage marks the inception of development and the commencement of the first experimental investigation. In the subsequent phase, nanomaterials, salt gel techniques, and genetic engineering are introduced, resulting in a notable reduction in the LOD for certain detection methods. The ultimate stage represents a pivotal moment in comprehending the dichlorvos detection process, with the introduction of nanotechnology, and it also witnesses advancements in biosensor detection. Throughout these phases, only a limited number of biosensor detection methods have been created that align with the requirements of the WFD. A substantial impact in this field was made by a research team, notably through their use of a biosensor based on acetylcholinesterase, which achieved an impressive limit of detection (LOD) of 2.21 × 10⁻⁹ μg/L for dichlorvos. This enzyme is embedded within activated conductive carbon, forming a connection to an electrochemical biosensor. These remarkably low LODs open up promising avenues for broader applications of enzymatic detection methods for OPPs [108]. In 2005, Law and Higson created an acetylcholine esterase-based biosensor that achieved a similar LOD to the previous research group [109]. They utilized screen-printed carbon electrodes containing cobalt phthalocyanine for detecting specific OPPs. Despite the membrane surface being coated with just the initial acetylcholine esterase layer, this setup demonstrated the potential to detect extremely low concentrations. Such low LODs imply that the detection of a particular pesticide occurs at the level of several thousand molecules. Exceptional stability, sensitivity, and selectivity suggest that biosensor methods for OPP detection could replace current methods in the future. Some biosensor stabilizers function by replacing hydrogen in the enzyme, preventing enzyme denaturation. Additionally, biosensors based on butyrcholine esterase, cholinesterase, choline oxidase, and tyrosinase are also employed for dichlorvos detection in water samples [136]. A similar approach using a combination of graphene and CdSe/ZnS hybrid systems on indium-doped tin oxide-coated glass electrode surfaces for the detection of dichlorvos through photochemical biosensor techniques was explored. The stability of the photochemical sensor was improved by the inclusion of acetylcholinesterase in the method [110]. Another type of biosensor utilizes carbon nanotubes in conjunction with ionic liquids. To address the sensitivity issue often associated with ionic liquids and carbon nanotubes, metal complexes such as cobalt phthalocyanine tetrasulfonate are introduced to amplify the current signal transmission. Similar to previous devices, thiocholine is measured amperometrically due to acetylcholine esterase inhibition. These systems have yielded an LOD of 1.55 × 10⁻⁴ μg/L and demonstrated a linear response within the range of 4.4194 × 10⁻⁴ to 0.1768 × 10⁻⁴ μg/L [111]. Despite achieving satisfactory results, it is worth noting that these biosensors are predominantly non-reversible [137].

4.3.2. Dichlorvos Analysis: Gas Chromatography

Khademi et al. developed and optimized the arrow SPME method with DVB/CAR/PDMS sorbent for the determination of OPPs. While this sorbent exhibited high affinity and resulted in reasonable GC-MS chromatogram peaks for OPPs, the LOD and LOQ for dichlorvos were 0.7 × 10⁻³ and 2.3 × 10⁻³ μg/L, respectively. The LOD is at the MAC limit for inland water but about ten times higher for other water bodies. Using a central composite design method, optimal extraction conditions were determined as 60 °C, 70 min, and 17% salt addition, with pH having a minimal effect. This method was applied to river water samples in Germany, finding no traces of OPPs. Arrow SPME proved to be more precise and sensitive than classic SPME, offering smoother penetration through septa, accommodating more sorbent volume, and better sorbent protection against mechanical stress, extending its lifespan [37].

5. Passive Samplers: Contemporary Insights

Using traditional water sampling methods at specific locations can often result in reduced sensitivity and non-representative sampling and analysis outcomes [138,139,140]. To address these monitoring limitations, time-integrative sampling techniques offer a solution, with passive samplers emerging as a prominent choice. Passive samplers are in situ devices that allow integrative sampling over a long period of time, several days, weeks, or months. These samplers capture the biologically available fraction of contaminants by continuously accumulating aqueous micropollutants over periods ranging from days to weeks [141,142]. Passive samplers, which are the primary alternative to classical sampling approaches, are designed to replicate bioconcentration kinetics observed in living organisms. They transition between various sampling modes, from integrative to equilibrium, making them a versatile and efficient tool for monitoring waterborne contaminants in diverse environmental settings. Furthermore, passive sampling devices have been developed to minimize the impact of various factors on the representativeness of samples during both collection and transfer to the laboratory. These diffuse devices contain a sorbent that serves as a medium for analyte molecules to accumulate due to differences in chemical potential. One of the important ways of applying passive samplers is in the detection of compounds that are released into the environment during a short time interval. Some of the most important advantages of passive samplers compared to classical spot sampling methods are relatively low cost and better efficiency in achieving much lower LOD and LOQ values [118,143,144]. An ideal passive sampler design should be cost-effective, straightforward to assemble, and uncomplicated to set up, recover, and assess [118]. Passive samplers should also possess the ability to detect a broad spectrum of substances with both precision and sensitivity [118]. Another important advantage is that they measure only the free concentration of the solute in water, eliminating the need for adjustments related to suspended solids. Consequently, the analysis results can be directly compared with findings from around the world. The foundation of passive samplers can be traced back to the observation that certain organisms have the capacity to accumulate specific compounds, depending on their polarity. This characteristic of passive samplers allows for the assessment of time-weighted average (TWA) pollutant concentrations in diverse environmental realms such as water, soil, and air, offering a substantial advantage over discrete spot samplers and thereby furnishing a more extensive and quantitative dataset [141,142,143,144,145,146].

Passive samplers are commonly categorized based on whether they sample organic or inorganic substances. The most frequently used passive samplers for organic substances include the Ceramic dosimeter (ceramic dosimeter and toximeter), Chemcatcher (universal passive sampler using Empore disk), Dosimeter (activated carbon receiving phase in a perforated acrylic housing), Ecoscope (sampler based on solvent-filled dialysis membranes and chelating sorbent discs), Gaiasafe (paper or fabric strips impregnated with a solution of a binding agent), Gore-Sorber (various sorbent materials filled in a carrier hose made of Gore-Tex), LDPE and silicone strips (low-density polyethylene or silicone strips), MESCO (membrane-enclosed sorptive coating), nd-SPME (negligible depletion-solid phase microextraction), Passive sampler according to Lee and Hardy (silicone polycarbonate permeation membrane and an adsorbent receiving phase), PDB (passive diffusion bag samplers), PISCES (passive in situ concentration-extraction sampler), POCIS (polar organic chemical integrative sampler), Porous Sampler according to De Jonge and Rothenberg (water-permeable porous sampler), Sampler according to Kot-Wasik (stainless steel housing), Solvent-filled dialysis membranes (non-polar solvent immiscible with water filled in a cellulose dialysis membrane), SPATT (solid-phase adsorption toxin tracking), SPMD (semi-permeable membrane device), TLC plate (thin-layer chromatography plate), TRIMPS (trimethylpentane-containing passive sampler), and TWA-SPME (solid-phase microextraction applied for the determination of TWA concentrations). Each of these samplers employs unique technologies and materials to effectively collect and analyze the target substances in various environmental contexts. Commonly employed passive samplers for organic substances encompass Chemcatcher, DGT (diffusion gradients in thin films), PIMS (passive integrative mercury sampler), PLM (permeation liquid membrane), SLM (supported liquid membrane), and SLMD (stabilized liquid membrane device) [118,147].

In this section, we will elaborate on various analytical techniques and specific passive sampling methods that have demonstrated compliance with the requirements of numerous legislations in a wide range of cases. The selection of the sampling site is of the most importance due to the particular distribution of pollutants [141,142]. One of the primary parameters in passive sampling, known as Rs, signifies the theoretical value of the water volume passing through the passive sampler per unit of time. Rs is unique to each analyte and device and constitutes an essential factor in quantitative analysis. Beyond the sampler type and exposure duration, the sampling Rs is affected by factors such as temperature, pH levels, ionic strength, water flow rate, solute concentration, and the type and conditions of the water systems [141,142,148,149]. So far, a large number of passive samplers have been developed depending on the priority of the analysis, of which POCIS (for hydrophilic) substances and SPMDs (mainly for hydrophobic compounds) are the two most common [148,150,151,152,153]. Passive samplers are frequently employed alongside spot sampling to harmonize and contrast outcomes, facilitating the assessment of the combined contribution of both approaches. Each variety of passive sampler comes with a predefined maximum exposure duration. Conversely, samplers must remain in the aquatic system for a sufficient duration to ensure adequate qualitative and quantitative results. However, it is worth noting that passive samplers like POCIS have not yet been officially integrated into WFD regulatory protocols, and their appropriateness for this purpose still requires verification. Nevertheless, in the guidance for minor surface waters, passive samplers are identified as a highly dependable method for sampling when analyzing extremely low concentrations.

Commonly utilized sorbent materials in passive samplers include Oasis MAX, Strata XAW, Oasis HLB, Oasis MAX, and C18 with suitable binding gels. Additionally, MIPs and ionic liquids find application in this context. However, there is a need for further sorbent development, particularly those with enhanced affinity for polar and ionic compounds. SPMD samplers are mainly employed for substances that predominantly exist as ions in the environment, typically those with a log of K_OW exceeding 3. A recent addition to the sampler repertoire is the hydrophilic-lipophilic balance sorbent-embedded cellulose acetate membrane (HECAM), which has demonstrated exceptional efficiency in analyzing OPPs and other emerging organic contaminants, whether they are hydrophilic or hydrophobic, in water. Valenzuela and co-workers wrote an extensive review on the different types of samplers, their practical application, and mathematical models that describe the parameters important for their operation. The goal behind developing new passive samplers is moving toward such methods becoming QuEChERS as well as developing new nanomaterial-based sorbents [141,142]. In addition to the aforementioned samplers, others have been developed such as Empore disk-Chemcatcher, MESCO, membranes assisted passive sampler (MAPS), in situ sealed or snap sampler (ISS), DGT developed for anionic pesticide sampling and Solid Ceramic Dosimeters for time-integrative passive sampling of volatile organic compounds in waters. Several mathematical models have been developed that describe the operation and sampling process using different passive samplers [141,142]. In the case of less hydrophobic and hydrophilic compounds like OCPs and PAHs, equilibrium is attained more rapidly. Nonetheless, it has been demonstrated that achieving equilibrium in most passive samplers is challenging due to the multitude of fluctuating parameters affecting the sampling process. A month duration is sufficient to collect samples of two- and three-ring PAHs because they quickly reach equilibrium on a silicone rubber (SR) passive sampler [154].

A novel method for detecting various pesticides using a passive sampler that incorporates hollow fiber liquid-phase microextraction, in conjunction with the GC-MS technique, has been introduced. This approach enables the analysis of OPPs, which are highly unstable in aqueous environments and rapidly degrade [155]. One of the initial studies in Europe comparing results obtained with various passive samplers was conducted by Miege and colleagues. However, the SPMD passive sampler did not exhibit adequate specificity for PAH compounds, making it unsuitable for assessing low-MAC pollutants in accordance with WFD guidelines [156]. A comparison between two refined passive sampling techniques was conducted, utilizing GC-MS with a 5% phenyl polysiloxane capillary column. The results indicated that the SPMD method outperforms POCIS in terms of achieving higher recovery rates for PAH determination [157]. Vrana and colleagues conducted an analysis of PAH compound content and concentration in samples from the Danube river near Vienna and Bratislava cities. They employed passive SPMD samplers, with three placed in Vienna and one in Bratislava. The sampling occurred in two phases: the first phase involved sampling at a single Vienna location during both summer and winter seasons. The SPMDs comprise an LDPE membrane filled with 1 mL of 95% pure triolein. With nominal dimensions of 2.54 × 91.4 cm and an exposure surface area of 460 cm², these samplers have a wall thickness of 75–90 μm. Each sampler contains 2 μg of individual performance reference compounds and is stored at −20 °C in gas-tight metal containers before use. The sampler volume (triolein + membrane) is 4.95 mL. To clean the SPMD samplers from debris and mud, analytes are extracted twice for 24 h by dialysis to hexane. Dialysates undergo further purification through gel permeation chromatography and silica gel or sulfuric acid-modified silica gel for PAH and PCB analysis, respectively. In the second phase, sampling spanned four seasons, each lasting 14 days, and the collected samples were subsequently analyzed using the GC-MS method. PAH separation was achieved using an HP-5MS column, while PCB separation utilized a DB5-MS column. The LOQ values ranged from 0.01 × 10⁻³ to 0.29 × 10⁻³ μg/L, satisfying the requirements of the WFD. During the initial phase, PAH concentrations fell within the range of 5 × 10⁻³ to 39 × 10⁻³ μg/L, and during the second phase, they ranged from 13 × 10⁻³ to 72 × 10⁻³ μg/L, all of which were within the WFD concentration limits. PCB concentrations were generally close to the LOQ, ranging from 5 × 10⁻³ to 16 × 10⁻³ μg/L, posing minimal environmental risk. In nearly all samples, except for sporadic deviations, elevated PAH concentrations were observed during the winter season, likely attributed to household combustion, as previously mentioned. The pollution source was identified as diffuse, with minor variations [118,158]. POCIS samplers designed for pharmaceutical sampling were used to assess the presence of PAH, pesticides (including cypermethrin and dichlorvos), phenols, pharmaceuticals, and hormones in surface water samples from the Baïse river, located in the Garonne basin in the southwestern region of France [40]. Two sets of triplicates were employed: one for chemical analysis and the other for bioassay, with the samplers submerged in water for a duration of one month. Subsequent to the analysis, the passive sampler discs were cleansed, and SPE was carried out using an Oasis HLB sorbent and a 30 mL mixture of dichloromethane and methanol (1:1, v/v). The resulting extracts were then concentrated using a vacuum evaporator and nitrogen stream. PAH recovery ranged from 70 to 102 ± 7%, while cypermethrin and dichlorvos exhibited recoveries of 73 ± 30% and 74 ± 30%, respectively. Out of a total of 132 samples analyzed, only 9 displayed recoveries below 70%. It is worth noting that a meaningful correlation between spot and passive sampling can be established only when the compound concentration remains constant over a period of time, as spot sampling captures only a single moment in the passive sampling process [40]. Low-density polyethylene (LDPE) samplers were utilized to assess the concentration and interchange of OCPs between the atmosphere and water within the Great Lakes Basin, with a specific focus on Lake Ontario and Lake Erie [159]. Trained volunteers deployed LDPE sheets in air and water at various locations along the southern coast of Lake Ontario. In Lake Erie, samplers were placed at nine U.S. coast locations and six Canadian locations with ship-based access by Environment Canada. Water and air samplers were colocated when possible. Atmospheric samplers, protected by inverted bowls, were situated 1–2 m above ground or water. Water samplers, tethered to anchored ropes, were suspended about 1 m below the water surface. Sampling campaigns, lasting approximately two months each, occurred between April and November 2011. Lake Erie on-lake samplers were exclusively deployed during the final campaign (August to October 2011). Following the sampling periods, the collected extracts were subjected to analysis via GC-MS employing a DB-5 capillary column. For the LDPE passive water samplers, extraction involved the use of dichloromethane, followed by elution through an aluminum-packed chromatographic column. Elution was accomplished with a mixture of n-hexane and dichloromethane (3:2, v/v), followed by solvent evaporation and replacement with n-hexane. The LOD for heptachlor and heptachlor epoxide were determined as 43 × 10⁻⁶ and 41 × 10⁻⁶ µg/PE, respectively, with recovery rates of 96 ± 8% and 101 ± 7%, respectively, for these compounds. The highest recorded heptachlor concentration in Lake Erie was 8.4 × 10⁻⁶ µg/L, affirming the efficacy of LDPE samplers in the comprehensive sampling of both air and water for the detection of various OCPs. Apart from quantifying concentrations, passive samplers find utility in toxicological assessments, non-target analysis, and biomonitoring [40,160]. In their study, Ahrens and colleagues investigated the effectiveness of five distinct passive samplers, namely SR, POCIS-A, POCIS-B, Chemcatcher^®, styrenedivinylbenzene-reverse phase sulfonate (SDB-RPS), and Chemcatcher^® C18, for the comprehensive analysis of 124 pesticides, including cypermethrin and dichlorvos, in the waters of river Fyris located in Uppsala, Sweden [160]. The study involved the retrieval of samplers on various days, specifically on days 5, 11, 10, and 26. Subsequent to the passive sampling period, compounds were extracted using different methods depending on the type of the passive sampler. The SR sampler underwent a drying process with a nitrogen stream and was afterward subjected to extraction using the Soxhlet method. For GC-MS analysis, samples were extracted through the SPE method using a 300 mL petroleum ether/acetone mixture (1:1, v/v), followed by evaporation under a nitrogen stream. LC-MS/MS analysis involved extraction with 300 mL of methanol, with subsequent nitrogen-assisted evaporation. POCIS samplers, after drying under vacuum, were subjected to SPE extraction, with analyte elution performed using 5 mL EA for GC-MS analysis and 1.5 mL methanol followed by 8 mL dichloromethane/methanol (4:1, v/v) for LC-MS/MS analysis. In both cases, extracts underwent nitrogen-assisted evaporation, with solvent replacement using cyclohexane/acetone (9:1, v/v) for GC-MS analysis and acetonitrile for LC-MS/MS analysis. For Chemcatcher^® SDB-RPS and Chemcatcher^® C18 samplers (Chemcatcher is produced in T.E Laboratories Ltd., Loughmartin Business Park, Tullow, Co. Carlow, Ireland), two rounds of sonication with ethyl acetate were conducted, first with 5 mL and then with 3 mL. The resulting extracts were combined, evaporated under a nitrogen stream, and the solvent was substituted with cyclohexane/acetone (9:1, v/v). GC-MS analysis utilized an HP-5MS UI capillary column, while online LC-MS/MS was performed using an SPE column Strata C18-E followed by a Strata X, and the subsequent sample separation took place on an analytical column Zorbax Eclipse Plus C18. The experiment demonstrated the effectiveness of K_OW values in predicting the affinity of specific compounds to particular passive samplers, with hydrophobic compounds exhibiting greater affinity for SR and polar compounds displaying increased affinity for POCIS A, POCIS B, and Chemcatcher^® SDB-RPS passive samplers [149,160]. On the other hand, POCIS A and POCIS B were shown to have the highest capacity and could be used over a longer time interval of analysis, in contrast to the SR passive sampler which showed the lowest capacity. The LOD and LOQ for dichlorvos were 10 × 10⁻³ µg/L and 33 × 10⁻³ µg/L, respectively, while the lowest recovery of 44% was in the case of SR samplers, and the highest recovery of 100% and 110% was in the case of POCIS A and POCIS B samplers, respectively [160]. Nguyen and co-authors conducted a study to evaluate the effectiveness of passive sampling employing three distinct passive samplers: POCIS, Speedisk, and SorbiCell [144]. Water samples were collected from areas managed by the Waternet and Rijnland water boards, situated near Amsterdam, known for contamination with polar substances from urban, agricultural, and wastewater treatment sources. Notably, the polyethersulfone membranes of POCIS were excluded from the extraction process. Subsequently, vacuum extraction, centrifugation, and nitrogen flow were used to dry the columns. Speedisks underwent vacuum extraction prior to elution, while both dry SPE sorbents and Speedisks were eluted three times with 3 mL of acetone. SorbiCell was eluted with 3 × 3 mL of methanol/acetone (1:1, v/v). Portions of acetone extracts from POCIS and Speedisk, as well as methanol/acetone fractions from SorbiCell, were transformed into methanol (1 mL) by evaporating the aliquot under nitrogen flow and dissolving the residues in methanol. Ultra-performance LC (UPLC-TOF) analysis was then performed on the methanol extracts. Additionally, another part of the acetone extracts from POCIS and Speedisk, along with methanol/acetone fractions from SorbiCell, was converted into a 40% methanol in HPLC water solution (2 mL) using the same method, for LC–MS/MS analysis of 16 compounds with negative ionization. The study reported average recovery rates of 72 ± 5% for POCIS, 73 ± 5% for Speedisk, and 43 ± 5% for SorbiCell. The LODs for the passive sampler devices were determined by dividing the LOD of the extracts by the concentration factors of water. None of the four pesticides of interest were detected in the water samples. However, LOD values for dichlorvos were established at 0.02 × 10⁻³, 0.05 × 10⁻³, and 0.4 × 10⁻³ µg/L for POCIS, Speedisk, and SorbiCell, respectively, while for cypermethrin, LOD values were 0.03 × 10⁻³, 0.1 × 10⁻³, and 0.4 × 10⁻³ µg/L for POCIS, Speedisk, and SorbiCell, respectively. POCIS demonstrated superior results, accumulating the broadest range of compounds in terms of log K_OW and the highest number of polar compounds, the primary targets of these samplers. The study concluded that the SorbiCell system is the least suitable for sampling a wide range of polar organic compounds in surface waters due to lower compound detection, issues with clogging, and inconvenient deployment. An investigation was conducted at three monitoring sites along rivers in France (the Clain, the Gier, and the Jalle) using passive POCIS samplers. Two methods for measuring contaminant levels were compared: the commonly used spot sampling method (162 samples) and passive sampling using POCIS (156 duplicates). Comparable repeatability was observed between POCIS and spot sampling replicates during the year-long study, which included 26 consecutive campaign series. Duplicate POCIS devices were exposed for 14 days each, and the results were averaged for the same campaign. The POCIS device comprised 200 mg of the Oasis^® HLB (Oasis is produced by Waters Corporation, 34 Maple Street, Milford, MA, USA) phase enclosed by polyether sulfone membranes (diameter 900 mm, pore size 0.1 μm) and metal washers with a 45 cm² open surface. Each POCIS was shielded with aluminum foil before and after in situ deployment, and the exposure system included a stainless steel canister from Exposmeter. Spot water samples (1 L) were collected every 14 days from the top 50 cm of sub-surface water during POCIS operations and were promptly transported to the laboratory, alongside POCIS devices, in a coolbox (5 °C ± 3 °C) within 24 h, following ISO 5667 protocols. The analysis of the extracts occurred in three separate laboratories, establishing an LOQ of 2.9 × 10⁻³ µg/L for dichlorvos; however, the WFD criteria were not met [147]. The study identified occasional instances of contamination for dichlorvos and metazachlor in September, as well as for chlortoluron in September/October when using POCIS, but not in single-spot samples. This highlighted POCIS ability to facilitate comparisons of AA-EQS with threshold values, especially for contaminants with low thresholds or significant daily variations. An interesting combined approach involving POCIS, the Oasis HLB passive sampler, and GF/F, coupled with a rapid 48 h methanol extraction process, yielded significant results. Polyethersulfone (PES) and PTFE membranes underwent methanol and Milli-Q water preconditioning, followed by air drying. Oasis HLB sorbent was treated with three rounds of methanol shaking and drying. POCIS-PES and POCIS-PTFE were assembled with a 180 cm² membrame/g Oasis HLB ratio; POCIS had a 45.8 cm² exposure area and 218.2 mg Oasis HLB. To mitigate pore size effects, consistent pore sizes were maintained for PES and PTFE, with different thicknesses (30 μm for PTFE, 132 μm for PES). Both POCISs were covered with solvent-rinsed aluminum foil until deployment. LC-MS with thermospray ionization (LC-MS/TS) using the Zorbax Eclipse Plus C18 column was employed, resulting in an LOD of 0.48 × 10⁻³ µg/L for dichlorvos [161]. The feasibility of utilizing cork-based passive samplers for water analysis was explored, and the findings from this research demonstrate the efficacy of employing cork-based passive samplers for the detection and measurement of triclosan in natural water sources [162]. Both cork passive samplers and miniaturized versions effectively identified and quantified triclosan, with cork passive samplers consistently revealing higher concentrations due to their enhanced sensitivity. These results underscore the promise of cork-based passive samplers as valuable tools for the surveillance and evaluation of emerging pollutants in natural water bodies. Furthermore, these samplers present a more environmentally friendly alternative to traditional techniques and have the potential to sensitively monitor emerging contaminants.

6. Conclusions

This review described analytical methods and approaches investigated in order to reach the requirement laid on analytical methods to be used in agreement with the WFD, focusing mostly on problematic substances like benzo(a)pyrene from the PAH group and pesticides (dichlorvos, cypermethrin, heptachlor, and heptachlor epoxide). Various sampling and analytical techniques meet WFD standards, but no single method covers all 45 substances. Multiresidue methods involve passive sampling, microextraction, and high-sensitivity detectors like HRMS. Passive sampling devices offer cost-effective solutions for extended water monitoring, addressing sensitivity and representativeness challenges. They provide a defined medium for calculating time-weighted average concentrations, essential for assessing pollutant bioavailability and fate in aquatic environments. Despite providing integrated measurements, they lack real-time monitoring crucial for studies on rapid environmental changes. Factors like temperature and water flow affect contaminant uptake, leading to problems with quantitative analysis. These methods successfully analyze diverse water sources, including wastewater, river water, seawater, tap water, and rainwater. Advancements in chromatographic methods, such as novel stationary phases and detection systems, aim to enhance precision and trace concentration monitoring. Additionally, sensor-based and electrochemical approaches are emerging, though their reproducibility poses challenges in routine analyses. Chromatographic methods with various types of mass spectrometers presented in this review achieve detection limits on the order of 10⁻⁶ μg/L, while modern analytical methods reach detection levels as low as 10⁻¹³ μg/L. However, electroanalytical methods often lack selectivity. Thus, it is recommended to use them for screening approaches, and when a compound is detected, it must be confirmed by analysis using a standardized method. The ongoing evolution of these methodologies reflects a collective effort to enhance our ability to monitor and safeguard the health of aquatic ecosystems. The Water Framework Directive, EPA, and similar regulations push the boundaries of analytical methods, particularly in terms of the LOD and LOQ, ensuring the quality of parameters for the status of water bodies.