# Polycyclic aromatic hydrocarbons in water environments: Impact, legislation, depollution processes and challenges, and magnetic iron oxide/graphene-based nanocomposites as promising adsorbent solutions

Joana Vaz-Ramos, Stéphane Le Calvé, Sylvie Begin

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Joana Vaz-Ramos, Stéphane Le Calvé, Sylvie Begin. Polycyclic aromatic hydrocarbons in water environments: Impact, legislation, depollution processes and challenges, and magnetic iron oxide/graphene-based nanocomposites as promising adsorbent solutions. Journal of Hazardous Materials, 2025, 490, pp.137726. 10.1016/j.jhazmat.2025.137726. hal-05388656

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Review

# Polycyclic aromatic hydrocarbons in water environments: Impact, legislation, depollution processes and challenges, and magnetic iron oxide/graphene-based nanocomposites as promising adsorbent solutions

Joana Vaz-Ramos<sup>a,b</sup>, Stéphane Le Calvé<sup>a,*</sup>, Sylvie Begin<sup>a,b,**</sup>

## HIGHLIGHTS

* PAHs in the environment, their health risks and current PAH depollution techniques.

* Highlight the great need of more efficient and more specific adsorbents for PAHs.

* State of the art and potential of graphene/ iron oxide materials for PAH removal.

* Understanding the interactions between PAHs and graphene/iron oxide adsorbents.

* The existing experimental challenges in comparing PAH adsorbent performances.

## GRAPHICAL ABSTRACT

<figure><img src="image_2.png" /></figure>

## ARTICLE INFO

**Keywords:**

PAH

Persistent organic pollutants

Pollutant removal

Nanomaterials

Adsorption

Composites

## ABSTRACT

Environmental pollution is a big challenge of today's world, as population continues to grow, and industrialisation and urbanisation increase. Out of the different micropollutants in the atmosphere and aquatic environments, polycyclic aromatic hydrocarbons are of particular importance because they have known severe associated health risks to human life and they have high stability, leading to their persistence in the environment. They are generally present in the environment in low concentrations, but, even at these levels, they pose threats. This review thus focuses on this family of pollutants, on their occurrence and consequences, as well as the current methodologies employed to remove them from water environments and the challenges that remain. This work then focuses on the potential of magnetic iron oxide/graphene nanocomposites for the adsorption of PAHs, extensively discussing past and undergoing works, as well as the interactions between these adsorbents and PAHs.

<footnote><sup>a</sup> Institut de Chimie et Procédés pour l'Energie, l'Environnement et la Santé (ICPEES), UMR-7515 CNRS-Université de Strasbourg, 25 rue Becquerel, Strasbourg 67087, France</footnote>

<footnote><sup>b</sup> Institut de Physique et Chimie des Matériaux de Strasbourg (IPCMS), UMR-7504 CNRS-Université de Strasbourg, 23 rue du Lœss, Strasbourg Cedex 2 67034, France</footnote>

<footnote>* Corresponding author.</footnote>

<footnote>** Corresponding author at: Institut de Chimie et Procédés pour l'Energie, l'Environnement et la Santé (ICPEES), UMR-7515 CNRS-Université de Strasbourg, 25 rue Becquerel, Strasbourg 67087, France.</footnote>

<footnote>E-mail addresses: slecalve@unistra.fr (S. Le Calvé), sylvie.Begin@unistra.fr (S. Begin).</footnote>
# 1. Introduction

As world population continues to grow and industrialisation processes increase, water pollution is a major world concern. As a consequence of human activities, such as agriculture, industries and domestic practices, many contaminants are released into water media [1]. In fact, according to the United Nations World Water Development Report 2023, more than 80 % wastewater worldwide is still released to the environment without treatment [2]. There is a wide range of pollutants present in water that can be harmful to human health, animals and the ecosystems and there is a growing concern for emerging pollutants, which are usually present in trace quantities (from less than ng/L up to µg/L) [1,3]. The effects to humans and the environment of some of these contaminants are still unknown [1], while others are well-known to have dangerous and severe effects [4]. The presence of the pollutants in water can cause perturbations to aquatic life forms, leading to decrease in biodiversity and loss of habitats, and they can also enter and accumulate in the food chain [5]. Humans can then be exposed to these pollutants by consumption of contaminated seafood and fish, or simply by drinking water [5]. Additionally, some of these pollutants are highly stable and resist to degradation and can also bioaccumulate [1,5] and thus are capable of staying in the environment for many years. These persistent pollutants include pesticides, pharmaceuticals, polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs), per- and polyfluoroalkylated substances (PFAS), other industrial chemicals, personal care products, etc. [1,6]. Due to their negative effects to humans and the environment, coupled with their persistence in the environment, it becomes of the utmost importance to manage their presence.

This review focuses on PAHs in the environment due to their long-lasting and widespread character in the environment, associated health risks (even at low concentrations), as well as due to the currently existing challenges related to their analysis and removal. This review will start with a description of PAHs, followed by their sources and behaviour in the environment, their effects on the human health and the environment, and finally, the existing legislation of PAHs will be briefly discussed. Following this, the general conventional techniques used nowadays in the depollution of PAHs in wastewaters will be described, before focusing more deeply on adsorption processes, given their widespread use for pollutant removal but also as this adsorption step is crucial for accurate analyte extraction in detection and quantification analyses. Adsorption concepts will be briefly described, before focusing on the current state of the art of the use on graphene/iron oxide nano-materials as adsorbent for this specific application.

# 2. Polycyclic aromatic hydrocarbons (PAHs) in air and water environments

## 2.1. General description

Polycyclic aromatic hydrocarbons (PAHs) are a large family of organic molecules with two or more fused aromatic rings in their structure (Table 1) [7,8]. They are considered as persistent pollutants because they are ubiquitous in the environment, being widespread in water, air, sediments and soil [9]. Furthermore, they have been found in foods, such as olive oil [10,11], and in consumer products, such as toys [10]. PAHs are highly hydrophobic compounds and are generally insoluble in water (or have a very low solubility); they have high melting and boiling points; and they have low volatility [7,8]. They are known to have dangerous effects on human health and the ecosystems [9], which is why they have gathered global attention in recent years. They are often found in low concentrations and so they are also referred to as micropollutants, which are compounds found in trace or ultra-trace levels in the environment (usually from ng/L up to a few µg/L in water, air and soils [3,12]).

Currently, more than 100 PAHs are known [10]; however, 16 of them

have been considered as priority PAHs by the US EPA (Environmental Protection Agency of the United States of America) because of their health effects [13]. In Europe, health agencies of certain countries are now extending the number of PAH monitored to include benzo(e)pyrene and benzo(j)fluoranthene [10] and they are also included in the Registration, Evaluation, Authorisation and Restriction of Chemicals (REACH) regulation of the EU [14]. In Table 1, the 16 US EPA Priority PAHs are presented along with their molecular structure and relevant properties.

PAHs can be categorised according to their number of rings and molecular weight into: low molecular weight PAHs (LMW-PAHs) for 2- and 3-ring PAHs; medium molecular weight PAHs (MMW-PAHs) for 4- and 5-ring PAHs; and high molecular weight PAHs (HMW-PAHs) for PAHs with 5 or more rings [19]. Alternatively, other authors consider only LMW-PAHs (two or three rings) and HMW-PAHs (four rings or more) [15]. In this work, the distinction between only LMW-PAHs and HMW-PAHs was used.

PAHs can also be categorised depending on the type of aromatic rings: alternant PAHs only have benzene rings, while non-alternant PAHs have benzene rings and other rings with less than 6 carbons [20].

In general, the physical properties of PAHs are dependent on their molecular weight and PAHs with the same molecular weight have similar properties, such as vapour pressure. As the molecular weight of PAHs increases, their solubility in water and vapour pressure decrease, while their melting and boiling point increase [7,8,15]. HMW-PAHs are more hydrophobic and more stable and, thus, less susceptible to degradation and so they can persist in the environment for many years [7].

The sources of PAHs in the environment can be both natural and anthropogenic. Natural sources include forest fires, volcanic eruptions, decay of organic matter or erosion of rocks containing petroleum hydrocarbons [8,21,22]. However, their main sources are anthropogenic. They are generated in the incomplete combustion of organic matter, such as fossil fuels and biomass [8,19,20,22]. Thus, they can occur in many different combustion processes, such as those involved in residential heating and cooking, vehicle emissions and air traffic, industrial production, agricultural fires, smoking, oil spills, etc. [19,22]. The introduction of PAHs into the environment occurs mostly by their emission into the atmosphere and usually as a mixture containing multiple of these compounds [21]. Airborne PAHs can be found in two phases depending on their vapour pressure: gaseous phase and particulate bound. In the atmosphere, they can travel long distances and they can also easily enter other environmental media, such as water bodies, soils and vegetation, by precipitation phenomena and/or by deposition/settling [6,19,21].

The main anthropogenic sources of PAHs as well as their transport in the environment are summarised in Fig. 1. For more details on the sources and occurrence of PAHs in the environment, references [19–23] can be checked.

After PAHs enter the environment, they can undergo transformations through multiple processes such as oxidation, photolysis, etc., and generate different PAH derivatives, which are often more toxic than the parent PAHs [9].

## 2.2. Effects of PAHs to human health

Humans can be exposed to PAHs and their derivatives through multiple routes [7]: inhalation of the PAHs present in air, ingestion of PAHs in water, as well as by skin contact with PAHs present in the different environments. Furthermore, another possible route of exposure is by ingestion, because, PAHs can easily enter the food chain and bio-accumulate, and food can also get contaminated with PAHs during cooking at high temperatures [20,24,25].

The severity of the effects of PAH exposure to human life greatly depends on the time and route of exposure, as well as on the concentration and type of PAH, and on the health condition of the exposed
<table><caption>**Table 1**

List of the US EPA 16 priority PAHs, their chemical formula and structure, molecular weight and water solubility [9,15–18].</caption><thead><tr><th>PAH</th><th>Chemical formula</th><th>Chemical structure</th><th>Molecular Weight (g/mol)</th><th>Vapor pressure (Pa)<sup>a</sup></th><th>Water solubility (mg/L)<sup>b</sup></th></tr></thead><tbody><tr><td>Naphthalene (NAP)</td><td>C<sub>10</sub>H<sub>8</sub></td><td></td><td>128.1</td><td>7.2</td><td>31.7</td></tr><tr><td>Acenaphthylene (ACY)</td><td>C<sub>12</sub>H<sub>8</sub></td><td></td><td>152.1</td><td>-</td><td>3.96</td></tr><tr><td>Acenaphthene (ACE)</td><td>C<sub>12</sub>H<sub>10</sub></td><td></td><td>154.2</td><td>&lt; 0.5</td><td>3.90</td></tr><tr><td>Fluorene (FLU)</td><td>C<sub>13</sub>H<sub>10</sub></td><td></td><td>166.2</td><td>0.04</td><td>1.69</td></tr><tr><td>Phenanthrene (PHE)</td><td>C<sub>14</sub>H<sub>10</sub></td><td></td><td>174.2</td><td>0.091</td><td>1.10</td></tr><tr><td>Anthracene (ANT)</td><td>C<sub>14</sub>H<sub>10</sub></td><td></td><td>178.2</td><td>3.6 &times; 10<sup>-4</sup> - 0.11<sup>c</sup></td><td>0.0432</td></tr><tr><td>Fluoranthene (FLE)</td><td>C<sub>16</sub>H<sub>10</sub></td><td></td><td>202.2</td><td>8 &times; 10<sup>-4</sup></td><td>0.20 - 0.26</td></tr><tr><td>Pyrene (PYR)</td><td>C<sub>16</sub>H<sub>10</sub></td><td></td><td>202.2</td><td>0.012 - 0.6<sup>c</sup></td><td>0.135</td></tr><tr><td>Benzo(a)anthracene (BaA)</td><td>C<sub>18</sub>H<sub>12</sub></td><td></td><td>228.3</td><td>-</td><td>9.4 &times; 10<sup>-3</sup></td></tr><tr><td>Chrysene (CHY)</td><td>C<sub>18</sub>H<sub>12</sub></td><td></td><td>228.3</td><td>8.4 &times; 10<sup>-5</sup></td><td>2 &times; 10<sup>-3</sup></td></tr><tr><td>Benzo(b)fluoranthene (BbF)</td><td>C<sub>20</sub>H<sub>12</sub></td><td></td><td>252.3</td><td>6.7 &times; 10<sup>-5</sup></td><td>1.5 &times; 10<sup>-3</sup></td></tr><tr><td>Benzo(k)fluoranthene (BkF)</td><td>C<sub>20</sub>H<sub>12</sub></td><td></td><td>252.3</td><td>1.3 &times; 10<sup>-8</sup> - 0.7 &times; 10<sup>-4</sup><sup>c</sup></td><td>0.8 &times; 10<sup>-3</sup></td></tr><tr><td>Benzo(a)pyrene (BaP)</td><td>C<sub>20</sub>H<sub>12</sub></td><td></td><td>252.3</td><td>7.3 &times; 10<sup>-7</sup></td><td>1.62 &times; 10<sup>-3</sup></td></tr><tr><td>Benzo(ghi)perylene (BghiP)</td><td>C<sub>22</sub>H<sub>12</sub></td><td></td><td>276.3</td><td>1.4 &times; 10<sup>-8</sup></td><td>2.6 &times; 10<sup>-4</sup></td></tr><tr><td>Indeno(1,2,3-cd)pyrene (IcdP)</td><td>C<sub>22</sub>H<sub>12</sub></td><td></td><td>276.3</td><td>1.3 &times; 10<sup>-8</sup></td><td>0.062</td></tr><tr><td>Dibenzo(ah)anthracene (DahA)</td><td>C<sub>22</sub>H<sub>14</sub></td><td></td><td>278.4</td><td>1.3 &times; 10<sup>-8</sup></td><td>2.49 &times; 10<sup>-3</sup></td></tr></tbody></table>

<footnote>a vapor pressure values at 20°C, except for ANT, CHY and BaP at 25°C.</footnote>
b water solubility values at 25°C for ACY, ACE, FLU, PHE, BaA, CHY, BkF, BaP, DahA, and BghiP; at 24°C for ANT, and PYR; at 20°C for IcdP; room temperature for NAP; and not specified for FLE and BbF.

c for PAHs where ranges are presented, additional values of vapour pressure at 25°C according to pubchem.ncbi.nlm.nih.gov are reported: ANT – 0.08 Pa, PYR – 6.0 × 10⁻⁴ Pa, and BkF – 1.3 × 10⁻⁷ Pa.

<figure><img src="image_3.png" /><figcaption>**Fig. 1.** Main anthropogenic sources of PAHs in the atmosphere and water bodies (rivers, lakes, oceans, etc). This figure has been designed using images from Flaticon.com.</figcaption></figure>

individuals [20]. Nevertheless, PAHs are known to have dangerous long-term effects on human health, which include carcinogenicity, genotoxicity, mutagenicity and immunotoxicity [9], with some of them being known endocrine disruptors [26]. The toxicity of PAHs is also usually dependent on their size, with HMW-PAHs being more carcinogenic than LMW-PAHs [20]. Nevertheless, there are exceptions, such as naphthalene (2 rings) that has higher carcinogenicity than certain HMW-PAHs, such as benzo(ghi)perylene (6 rings) [20].

Given their different carcinogenicity, multiple agencies worldwide have classified PAHs into groups according to this criteria, like the International Agency for Research on Cancer (IARC) [20], US EPA [7] or the European Union (EU) [10]. Out of the priority PAHs, benzo(a)pyrene (BaP) has gathered the attention of the scientific community, due to its higher carcinogenicity. It is classified as carcinogenic by the IARC (Group 1), probably carcinogenic by the US EPA (Group B2) and presumed carcinogenic by the EU (Group 1B). Additionally, the EU also classified it as mutagenic (Group 1B) and presumed reproductive toxicant (Group 1B) [27]. Therefore, BaP is generally used as a marker for PAH carcinogenicity and considered to represent PAHs [20].

PAHs can have other long term effects caused by chronic exposure to them, which include cataracts, damage of the kidneys and liver, breathing problems, and decrease of the function of the immunity system [21]. It has also been reported that PAHs can cause short-term effects in cases of acute exposure of high concentration of PAHs. These side effects include irritation and inflammation [21]. Furthermore, certain PAHs, namely anthracene, benzo(a)pyrene and naphthalene, are known to be skin irritants, or even skin sensitisers, in the case of ANT and BaP [21,28].

## 2.3. Legislation on the presence of PAHs in water environment

Given the health risks of PAHs exposure, many international

regulatory agencies have established guidelines and target values for the concentration of these pollutants in water environments (Table 2), most of them being in terms of BaP concentration.

PAHs are heavily regulated in aquatic environments (Table 2). In the European Union, PAHs in water bodies have been monitored and legislated for many years. In 2000, the EU established the Water Framework Directive (WFD) (Directive 2000/60/EC, last consolidated in 2014) [29] and PAHs were included in the list of priority pollutants of concern in surface water [5] (i.e., water that is on top of land, such as rivers and lakes, as well as coastal and transitional waters [29]). Following the WFD, the EU established the Environmental Quality Standards (EQS) that can be found in Directive 2008/105/EC (last consolidated in 2013 with Directive 2013/39/EU) [30]. Two different types of standards were set to protect against both long-term and short-term effects: the annual average (AA) and maximum allowable concentrations (MAC), respectively [5,30]. While MAC values also take into account short-term effects to aquatic life, the AA values have protection of human health as the main goal. In this directive, the family of PAHs, including BaP, BbF, BkF, BghiP and IcdP, is mentioned. However, for AA-EQS, a value is only presented for BaP, corresponding to 1.7 × 10⁻⁴ µg/L, and it is indicated that BaP can be taken as a tracer for the PAH family, due to its high carcinogenicity. Additionally, EQS values are also established for naphthalene, anthracene and fluoranthene. For PAHs (in terms of BaP) and for fluoranthene, an EQS for biota (in this case, crustaceans and molluscs) is also established.

The EU also established the Drinking Water Directive (Directive 98/83/EC last amended in 2020 with Directive (EU) 2020/2184 [31]) where standard values for many different pollutants including PAHs and other emerging pollutants can be found. In these directives, it is established that a concentration of 0.10 µg/L must be met in drinking water for a sum of PAHs that includes BbF, BkF, BghiP and IcdP. For BaP alone, the established concentration is lower (0.01 µg/L).
<table><caption>**Table 2**
Guidelines and regulation on PAHs in the environment by different regulatory agencies.</caption><thead><tr><th>Agency</th><th>Environmental media</th><th>Established Standards</th><th>Compound</th><th>Guideline value (µg/L)</th><th>MAC-EQS</th><th>EQS</th></tr></thead><tbody><tr><td>WHO</td><td>Drinking water</td><td>Compound</td><td>BaP</td><td>0.7</td><td></td><td></td></tr><tr><td>EU</td><td>Surface waters<br>(Directive 2008/105/EC,<br>2013/39/EU)</td><td>Compound</td><td></td><td>AA-EQS<br>(µg/L)</td><td>0.27 or<br>0.027<sup>a</sup></td><td>Biota<br>(µg/<br>kg)</td></tr><tr><td></td><td></td><td>Compound</td><td>PAHs</td><td>1.7 × 10<sup>-4</sup></td><td></td><td>5</td></tr><tr><td></td><td></td><td>Compound</td><td>- BaP</td><td>-</td><td>0.027<sup>a</sup></td><td>-</td></tr><tr><td></td><td></td><td>Compound</td><td>- BbF</td><td>-</td><td>0.017</td><td>-</td></tr><tr><td></td><td></td><td>Compound</td><td>- BkF</td><td>-</td><td>0.017</td><td>-</td></tr><tr><td></td><td></td><td>Compound</td><td>- BghiP</td><td>-</td><td>8.2 × 10<sup>-4</sup></td><td>-</td></tr><tr><td></td><td></td><td>Compound</td><td>- IcdP</td><td>-</td><td></td><td></td></tr><tr><td></td><td></td><td>Compound</td><td>NAP</td><td>2</td><td>130</td><td>-</td></tr><tr><td></td><td></td><td>Compound</td><td>ANT</td><td>0.0063</td><td>0.12</td><td>30</td></tr><tr><td></td><td></td><td>Compound</td><td>FLE</td><td>0.1</td><td>0.1</td><td>-</td></tr><tr><td>EU</td><td>Drinking water<br>(Directive 98/<br>83/EC, 2020/<br>2184)</td><td>Compound</td><td>BaP</td><td>0.010</td><td></td><td></td></tr><tr><td></td><td></td><td>Compound</td><td>PAHs</td><td>0.10 (sum of BbF, BkF, BghiP and<br>IcdP)</td><td></td><td></td></tr><tr><td>US EPA</td><td>Ambient water</td><td>Compound</td><td></td><td>Recommended quality criteria<br>(µg/L)</td><td></td><td></td></tr><tr><td></td><td></td><td>Compound</td><td>BaA</td><td>0.0012</td><td></td><td></td></tr><tr><td></td><td></td><td>Compound</td><td>BaP</td><td>0.00012</td><td></td><td></td></tr><tr><td></td><td></td><td>Compound</td><td>BbF</td><td>0.0012</td><td></td><td></td></tr><tr><td></td><td></td><td>Compound</td><td>BkF</td><td>0.012</td><td></td><td></td></tr><tr><td></td><td></td><td>Compound</td><td>DahA</td><td>0.00012</td><td></td><td></td></tr><tr><td></td><td></td><td>Compound</td><td>IcdP</td><td>0.0012</td><td></td><td></td></tr><tr><td>US EPA</td><td>Drinking water</td><td>Compound</td><td>BaP</td><td>0.2</td><td></td><td></td></tr><tr><td>France</td><td>Discharges into<br>the<br>environment</td><td>Compound</td><td>ANT</td><td>25</td><td></td><td></td></tr><tr><td></td><td></td><td>Compound</td><td>FLE</td><td>25</td><td></td><td></td></tr><tr><td></td><td></td><td>Compound</td><td>NAP</td><td>130</td><td></td><td></td></tr><tr><td></td><td></td><td>Compound</td><td>PAHs</td><td>25 (sum of BaP, BbF, BkF, BghiP and<br>IcdP)</td><td></td><td></td></tr></tbody></table>

Similarly, the World Health Organization (WHO) has established Guidelines for Drinking-water Quality, where a recommended value of 0.7 µg/L for BaP is established [32].

The US EPA has also defined quality criteria for ambient water [33, 34], where all mentioned PAHs should meet recommended maximum concentrations of 0.00012–0.012 µg/L, depending on the PAH. PAHs in drinking water are also monitored in terms of BaP, which should be in concentrations lower than 0.2 µg/L [35].

In France, the ministerial order of 2 February 1998 (amended multiple times until 2023) imposes limit values for the industrial discharges [36]. For example, in coke fuel production plants, the effluents must meet a maximum PAH concentration of 0.1 mg/L and 0.03 g of PAH per g of produced coke [36]. In this ministerial order, they also established limit values for multiple PAHs (ANT, FLE, NAP, BaP, BbF, BkF, BghiP and IcdP) for all effluents discharged into natural water bodies [36] (Table 2).

As seen here, the established limit values are not consensual for the different regulatory entities. For drinking water, while the US EPA and WHO established higher concentrations of 0.2 and 0.7 µg/L, respectively, the EU is more cautious and established a concentration of only 0.01 µg/L for BaP. This value established by the EU is 100 times lower than BaP solubility in water, and thus it is possible for BaP to be frequently present in such range of concentrations. Therefore, the lower concentration established by the EU should be taken as the guide value, in order to better protect human health. Furthermore, if we look at the limit concentrations of discharges into water bodies, we see that PAHs can be released in concentrations around 25 µg/L (130 µg/L for NAP), which are significant concentrations if we compare them with their

solubility.

### 3. Current conventional techniques for depollution of PAHs from polluted water

PAHs have been detected in wastewater treatment plants (WWTPs) that collect domestic or industrial effluents in various regions of the world [15]. PAHs have been found in WWTPs influent streams in significant concentrations [15], ranging from 190 ng/L [37] up to 312 µg/L [38]. During the treatment processes in WWTPS, the PAH concentration decreases but residual PAHs are still found at the end [15] and, thus, released into the environment. The different methods that can be used to remove PAHs from polluted water in WWTPs are given in Fig. 2 and detailed below.

#### 3.1. Physical treatment

In membrane processes, polluted water is passed through a porous filter media that retains pollutants, while letting water pass through [8]. The materials used in the filters can correspond to activated carbon, sand or different polymers (for example, polyamides) or composites [8]. In these processes, while particle-bound PAHs are removed due to the size exclusion of the particles by the membrane, the soluble PAH fraction gets bound to the filter materials. Different membrane processes can be used to remove PAHs from polluted waters, such as ultrafiltration, microfiltration, nanofiltration and reverse osmosis [8]. Nanofiltration has been reported the most efficient in the removal of PAHs from water [8,9], with size exclusion being an important trapping mechanism in this type of membrane process, considering the high molecular weight of PAHs. Membrane technologies can be used alone, but, preferentially, they should be combined with other conventional water treatment methods [9] as their removal efficiency is not total.

Flotation is a simple process to operate, where gas is bubbled through polluted water to separate pollutants, in particular oils, greases, and suspended solids. Its effectiveness depends on the pollutants to be removed, temperature and other parameters [8]. Dissolved air flotation is known to remove large particles [8] and therefore this technique can be used to remove particulate-bound PAHs in water. Furthermore, the removal of LMW-PAHs from aqueous water has been shown in recent studies [39], as well as the removal of total PAHs from produced water by inducing air flotation in the presence of the surfactant Tween 80 (polyoxyethylene (80) sorbitan monooleate) [40].

Adsorption processes are one of the main techniques for the treatment of wastewater in general [41] and are also a well-known method for the removal of PAH from water [9], where PAHs are removed onto the surface of adsorbent materials. They can be divided into batch

<figure><img src="image_4.png" /><figcaption>**Fig. 2.** Main methods for the removal of PAHs from polluted waters (adapted from [8]).</figcaption></figure>

<footnote>a higher value established for inland surface waters and lower value for other surface waters.</footnote>
processes, more used in laboratory scale, and continuous processes, more interesting for large-scale WWTP applications. Numerous materials have been reported to adsorb PAHs from water, such as sand, activated carbons, biochar, resins or fibrous membranes, and more recently, carbon nanomaterials, such as carbon nanotubes or graphene [9,41]. Activated carbon is the most common adsorbent for the capture of PAHs and other toxic pollutants in wastewaters [41], because of its high efficiency related to high specific surface area and homogeneous microporosity [41]. This material is produced from carbonisation and subsequent activation of carbonaceous sources (such as petroleum coke, sewage sludge, or plant-based residues) [41]. Its adsorption efficiency depends on the contact time with polluted water, concentration of both activated carbons and the PAHs, on the other components present in the water sample, as well as on the activation parameters during its synthesis. According to Pathak et al., from 2010 to 2021, around 50 % of the works reporting on the adsorption of PAHs corresponded to this type of adsorbents, followed by biochar [41].

Adsorption processes have advantages over other methods of PAH removal, such as being flexible and simple to operate, and low-cost [8,9, 41]. However, regeneration of the adsorbents is frequently needed, which can be expensive. It also poses risks of secondary pollution by the adsorbed PAHs that need to be recovered and treated, in order to recycle the adsorbents [8,9]. In addition, in the case of highly particle-laden water, an upstream filter must be installed to avoid clogging the adsorbent.

Furthermore, adsorption also occurs in other water treatment processes, such as filtration, where PAHs can be adsorbed onto the filter materials, or in biological treatment, with PAHs being adsorbed onto sludge particles [9].

## 3.2. Chemical treatment

Coagulation and precipitation are common treatment methods for polluted water and are often used together for higher efficiency results. In coagulation, it is possible to remove colloidal and suspended matter from water [8,9], which can allow the removal of particulate bound PAHs, as well as to decrease the organic content, including PAHs, in solution [9]. As an example, aluminium sulphate is a coagulant that is commercially available and that has been used in the treatment of industrial wastewater to remove PAHs [9]. These processes are easy to implement and operate, however they lead to the formation of sludge and secondary waste [8].

Electrochemical technologies include different processes such as electro-coagulation and electro-oxidation. They have shown potential in removing PAHs from waste [8], but the difficulty to scale-up and the necessity of skilled labour [8] limit their use in wastewater treatment.

In chemical oxidation, PAHs or other pollutants are degraded by oxidation. The efficiency of this technique depends on the chosen oxidant and the dose, contact time, pH, etc. [9]. Sodium persulfate and potassium permanganate have been used for the removal of PAHs from water, as well as conventional Fenton oxidation where OH radical are produced in water by H₂O₂ and a source of Fe²⁺ [8]. Chemical oxidation processes are generally quick, however, for PAHs they tend to not lead to complete oxidation to water and CO₂, but instead to the formation of intermediates that are often more resistant to oxidation [9]. Furthermore, it is possible that these formed compounds are more dangerous than the parent PAHs.

In recent years, advanced oxidation processes have become promising to treat water pollutants that are not removable by conventional techniques (either because they have high chemical stability or low biodegradability) [9]. The oxidants used in these processes include ozone, hydrogen peroxide and Fenton-based processes (for example, ultraviolet/hydrogen peroxide/Fe³⁺) [8]. Ozone has been shown to lead to high efficient degradation of PAHs, both in laboratory scale and in WWTPs [9,42], however the problem of possible toxicity of oxidation products remains. In general, in advanced oxidation processes, complete

oxidation of PAHs is typically an expensive process involving several steps and combining potentially several oxidation techniques.

## 3.3. Biological treatment

Biological processes are used in large scale in WWTPs [15] and PAHs can be biodegraded either aerobically or anaerobically [8], in different processes that include activated sludge, bioreactors, sequencing batch reactors, membrane bioreactors, bioremediation, phytoremediation, etc. [8].

In activated sludge treatment, which is widely employed in municipal WWTPs, microorganisms and pre-treated wastewater are mixed, promoting the removal of remaining nutrients and degradation of organic matter present in water [1]. In this treatment, PAHs can be removed by three mechanisms: biodegradation by the microorganisms, adsorption onto the sludge, and volatilisation during aeration [15]. However, it has been reported that the contribution of volatilisation is less than 2 % [15].

In bioremediation and phytoremediation, microorganisms, enzymes and green plants are used on polluted substrates to remove the present pollutants [8]. Bioremediation using microorganisms has been reported for PAH removal [8,43] and multiple plants and substrates (e.g., sudan grass and *Arundo donax*) have been shown to be efficient on the removal on PAHs from soil or in WWTPs [8,44,45]. However, these processes tend to need lengthy degradation times, as well as long times to find the optimal working conditions [8]. Additionally, the removal of PAHs in wetlands has also been reported, depending on the microfauna and flora present [42].

## 3.4. Combined treatment

Combined treatment makes the most of different chemical, physical and biological treatments, in order to have more efficient elimination processes for WWTPs. These integrated processes have also shown promising results for PAH removal, such as being able to extensively eliminate HMW-PAHs present in water [8]. For example, biodegradation processes coupled with Fenton oxidation have been used to remove PAHs like NAP and PHE [8]. Biofiltration, which combines physical and biological treatments, has also been studied for PAH removal [42].

Of all treatments used in WWTPs, it seems that currently there is not a process implemented specifically to target the presence of PAHs in polluted water, but, instead, there is the evaluation of the already established processes also in terms of PAH removal efficiency. It seems that adsorption processes (particularly with activated carbon), and activated sludge treatments (where sorption into sludge and biodegradation occur) [41,42] are the main processes for removal of PAHs in WWTPs. Of these two, adsorption processes are particularly promising since by designing the adsorbent materials, there is potential to make these processes more efficient for PAH capture.

## 4. Sorption processes

Sorption are physical-chemical processes that have been known for decades, but it was in the last four decades that there was a big development of adsorption processes, leading to it now being considered as a major separation technique even industrially [46]. In the case of PAHs, adsorption is an important mechanism to consider in depollution applications. Furthermore, adsorption processes can also be important for the monitoring and quantification of PAHs in the environment, since multiple analytical sampling/ preparation methods, involve adsorption of PAHs into a solid phase prior to analysis.

In fact, the quantification of PAHs in the environment usually involves different steps: sample collection; sample preparation; and PAH detection and analysis. In the case of water environments, the sampling can simply correspond to direct collection of a sample of water into a container for analysis of liquid phase PAHs, but it can also be done by
collection into solid sorbents, followed by further sample preparation [47]. Afterwards, environmental samples can go through different steps of clean-up, pre-concentration and so on, until the sample is finally ready for PAH detection and analysis usually by chromatographic methods (such as HPLC-FLD/UV [16,20,47,48] or GC-MS [20,47,48]). Extraction processes are commonly used in these preparation steps as they allow both pre-concentration and isolation of the PAHs from the matrix depending on the specificity of the extraction process. Different types of extraction can be performed, ranging from sorptive extraction processes, with solid adsorbents, to solvent extraction [47]. Sorptive extraction processes include solid-phase extraction (SPE), solid-phase microextraction (SPME), stir bar dispersive extraction (SBDE), magnetic solid phase extraction (MSPE), etc. [7,47].

Thus, we see the importance of adsorption processes for both the removal and monitoring of PAHs in the environment. Therefore, the theoretical concepts behind sorption processes will be here briefly covered.

## 4.1. Adsorption of target compounds

In adsorption, there are two phases: the liquid (or gas phase), where the target molecules (called adsorbate) are, and the solid phase (called adsorbent), which will "capture" the target molecules on its surface [46, 49]. The adsorbate molecules in the liquid/gas, when brought in contact with the adsorbent, diffuse and get attached to its surface [49]. Generally, since it is a surface phenomenon, the higher the specific surface area (SSA) of the adsorbent, the higher the adsorption of the adsorbate [49]. Adsorption processes have found application in many different areas, such as catalysis, hydrocarbon fractionation, water treatment, enzyme purification and protein isolation, odour/taste removal, etc. [49]. However, there are still limitations to a large scale use of adsorptive processes, namely related to the cost of the adsorbent and the necessity to regenerate it [50].

Adsorption processes can occur in batch (discontinuous processes) or in a dynamic reactor with a constant flow of the adsorbate through an adsorbent or adsorbent bed (continuous processes). Adsorption can also be divided into physisorption and chemisorption, depending on the nature of the adsorption process [49].

Adsorption is a complex surface process. It depends highly on the nature of adsorbent and adsorbate, as well as in other parameters, such as the adsorbent SSA and porosity, temperature, pressure (for gases) and pH (for solutions), as well as the presence of other compounds in the liquid or gas [49]. When studying adsorption processes, it is common to define the adsorption capacity q, which corresponds to the amount of compound adsorbed (usually in mg) per g of adsorbent. For a typical batch liquid-solid adsorption process, the adsorption capacity is given by:

$$q = \frac{(C_0 - C)V}{m_{ads}} \qquad (1)$$

where q is the adsorption capacity, C₀ is the initial adsorbate concentration in the liquid, C is the remaining adsorbate concentration in the liquid after adsorption, V is the solution volume, and mads is the adsorbent mass.

It is also common to define the removal efficiency of the adsorbent (RE), which is given by:

$$RE = \frac{C_0 - C}{C_0} \times 100 \qquad (2)$$

Several types of models are used to better understand the adsorption process between a certain system of adsorbate/adsorbent. These can be divided into three categories: kinetic models, isotherm models and thermodynamic equations [49].

### 4.1.1. Adsorption kinetics

Adsorption kinetics correspond to the rate at which the adsorbate molecules move from the liquid/gas phase and are adsorbed to the adsorbent surface, in certain experimental conditions [51]. The study of adsorption kinetics allows to understand when the adsorption equilibrium is reached (i.e., when the molecules adsorb at the surface of the absorbent at the same rate as they desorb from the adsorbent [46]). These studies are performed by quantifying the analyte concentration after different adsorption times.

There are multiple models that have been developed to describe adsorption kinetics and they have been largely described in the literature [51,52]. The most commonly used are the pseudo-first order and the pseudo-second order models [51,52].

The pseudo-first order model was described by Lagergren in 1898 [53,54] and generally this model is valid when there are no significant changes in adsorbate concentration [51].

The pseudo-second order model was first presented by Ho and McKay [53,55] and it initially assumed the existence of a chemisorption rate-limiting step [51] and it can represent processes where concentration of adsorbate is low [51].

Both pseudo-first order and pseudo-second order models have been widely used to describe adsorption of organic pollutants in water [52, 53], with pseudo-second order being the most used for adsorption of a wide range of organic compounds by nanoadsorbents [52]. Even though pseudo-second order was first developed to describe chemical adsorption, it has been widely used in situations where adsorption was physical [53].

Other less common models are sometimes used for kinetics studies of the adsorption of organic pollutants, such as the Weber-Morris intraparticle model [51,52].

### 4.1.2. Adsorption isotherms

Adsorption isotherms are used to represent the equilibrium relationship between the adsorbed target molecules in the surface of the adsorbent and the target molecules still in the surrounding liquid (or gas) phase at constant temperature [56]. Adsorption isotherm studies are done by performing experiments with different initial concentrations of analyte, in order to establish a relationship between concentration and adsorption capacity at equilibrium. Isotherm studies give an indication about how much adsorbate can be adsorbed by the chosen adsorbent [52], in the chosen experimental conditions (and at a constant temperature), which is essential to assess the potential of an adsorbent and to optimise adsorption technologies [56]. They also provide insights into how the adsorbate and adsorbent interact [56].

A wide range of isotherm models can be found in the literature, with different assumptions and for different situations. The Freundlich and Langmuir isotherm models are two of the most used isotherm models in the adsorption of organic pollutants [52].

Freundlich isotherm describes reversible and non-ideal adsorption [50] and it can apply not only to monolayer, but also multilayer adsorption in heterogeneous surfaces [50,56]. It is an empirical model first proposed in 1909 by Herbert Freundlich for gas-solid adsorption [56], but that is also widely applied to liquid-solid adsorption.

The Langmuir isotherm was also originally established to describe gas-solid adsorption [50], but is extensively used to describe adsorption of organic molecules in liquid media. This model assumes that there is only monolayer adsorption, and that the adsorbent surface and the distribution of adsorption sites on it are homogeneous [56]. Moreover, it considers that the adsorption only occurs in a finite number of fixed active sites, which are identical to each other, and that adsorbate molecules on adjacent sites have no interaction with each other [50,56].

Freundlich and Langmuir models are the most used ones for adsorption of organic compounds in polluted water but apart from these, the Dubinin-Radushkevich isotherm model, or the more generalised version Dubinin-Astakhov isotherm, have also been applied to fit adsorption data of organic pollutants on nanoadsorbents but to a much
smaller extent [52]. Other models have also been used in certain cases to describe the adsorption of organic pollutants by nanomaterials [50,52], such as the Temkin isotherm, the Sips isotherm or the Toth isotherm.

## 4.2. Desorption for chemical analysis and/or sorbent regeneration

Desorption is the process by which the adsorbed molecules are released from the surface of the adsorbent [49] back to a liquid/gas phase. After adsorption occurs, this process is essential to ensure the regeneration of the adsorbents, so that they can be reused again [52]. The reuse of adsorbents is important for economic reasons, but it is also valuable from an environmental point of view.

Pollutants can be desorbed from the adsorbent surface by different techniques and this process also needs to be optimised. Desorption using organic solvents is a common method to desorb organic pollutants. For this, the appropriate solvent must be chosen to efficiently elute the pollutant molecules from the adsorbent and, thus, this solvent should have high affinity with the targeted pollutants, in order to have an efficient desorption [57]. In the case of PAHs, for example, typical desorption solvents include dichloromethane, acetone, chloroform, methanol, hexane and acetonitrile, or mixtures of these solvents [58–60]. Multiple cycles of desorption can be performed to increase the desorbed amount. However, the elution solvent should not damage the adsorbent materials [57], so that it does not reduce their adsorption capacity, and it should be the least costly and the most environmentally-friendly possible [57]. In large scales, the generated volume of solvents can be significative, which can be a disadvantage of this desorption process.

Alternatively, thermal desorption or regeneration can also be used, where the temperature is increased to promote the desorption of the adsorbate molecules from the adsorbent surface. Thermal desorption is a commonly used technique in the analysis of volatile and semi-volatile compounds coupled with GC analysis [61]. Alternatively, thermal desorption can also be coupled with the use of an organic solvent, with temperature helping to extract the adsorbed molecules into the organic solvent. While this regeneration method is viable for analytical applications, it can become quite expensive due to its energy consumption on larger scale and it is also time consuming [62]. Furthermore, high temperatures can lead to a degradation of the pores depending on the adsorbent type [62], which can compromise the reuse of the adsorbent or at least degrade its sorption performance.

Other methods exist to desorb the pollutants and regenerate the adsorbents, such as acid treatments [57], desorption by the use of ultrasounds [63] or supercritical fluid extraction [62], but they are not used as frequently. Thermal desorption and solvent desorption are thus the two most commonly used methods. At lab-scale, a lot of studies employ organic solvents to regenerate or extract the adsorbed organic pollutants from the adsorbents and under optimal conditions total or almost total desorption is possible. However, at a larger scale, such as in wastewater treatment plants (WWTPs), adsorbent regeneration remains one of the big challenges, especially as the environmental impact and costs of such procedures always need to be considered.

## 4.3. Main conclusions and research opportunities of the previous sections

In the previous sections, the need to develop tools to remove PAHs was put into evidence. However, to the best of our knowledge, these still correspond to small-scale laboratory studies, not yet implemented in real WWTPs.

On a larger scale, Arvia Technology, a UK-born company specialised on the tertiary treatment of industrial water, has reported on the use of their Nyex™-a system for the removal of benzo(a)pyrene and fluoranthrene in a wastewater stream of a chemical plant [64]. This system is a combined process of adsorption and electrochemical oxidation, [64] and they reported reaching removals above 68.3 % for fluoranthrene [64]. However, this system was developed for active pharmaceutical

ingredients (APIs), herbicides, pesticides and residual organic chemicals and not specifically for PAHs [64].

More generally, in the case of polluted water treatment in WWTPs, it has been reported that adsorption with activated carbon [41] and activated sludge treatments [15,42] are the main PAH removal techniques. Nevertheless, it also seems that the removal of PAHs relies heavily on the methods already employed in WWTPs for depollution practices in general, even if they are not optimised for PAH depollution. Moreover, even if big reductions of PAH load are achieved in WWTPs, according to reported values, there are still important trace levels of PAHs in the outlet of WWTPs [15]. For example, according to Gateuille et al., on the Seine River basin there is a discharge of 21 kg/year from WWTP into the river based on PAH mass balances [65], but this value should be taken with caution due to approximations and limited data availability [65]. Berardi et al. reported influent total PAH concentrations of 5.6–66 µg/L and effluent concentrations of 0.075–0.22 µg/L in the Baciacavallo WWTP in Prato, Italy in 2011 [66]. Similarly, Mezzanotte et al. reported total PAH concentrations (dissolved + particulate) of 2.41 and 5.34 µg/L in the secondary treatment influents and 0.15–0.19 and 0.15–0.26 µg/L in the outlets of the Nosedo and Alto Seveso WWTPs, respectively, in Italy in 2013–2014 [67].

So, PAHs are still released into the environment at the exit of WWTPs, even if in much lower levels than the untreated wastewater. This could be related to the lack of methods in WWTPs that target specifically the depollution of such type of nefarious pollutants. Thus, it is important to develop more efficient processes for the capture of PAHs in water. As previously discussed, adsorption processes have gathered a lot of attention and activated carbon is the main adsorbent used for PAH capture in large scale. Actually, activated carbon is a well-known good adsorbent for many types of pollutants ranging from heavy metals to pharmaceuticals, pesticides, dyes and many other organic pollutants (which include PAHs) [68,69]. This wide range of applicability and consequent competition between pollutants for adsorption sites could possibly or partially explain the incomplete removal of PAHs in complex environmental waters in WWTPs.

Thus, there is an opening to design adsorbent materials that are more efficient and more specific for PAHs to be implemented in depollution practices. The development of highly selective adsorbents would also bring improvements to the detection and chemical analysis of PAHs in the environment, as it would allow the specific and more sensitive extraction and analysis of PAHs in complex environmental samples. This is particularly important because it is not enough to remove PAHs from the environment, but also to accurately quantify them, in order to know the risks posed to human health.

In the last decades, nanotechnology has caught the interest of the scientific community for a wide range of applications, including adsorption. Nanomaterials are materials where at least one dimension is in the nanometric range (1–100 nm). They have unique properties because of their small size, such as quantum size effects, high specific surface area, high reactivity, strong sorption, fast dissolution, tuneable surface chemistry, and others [8]. Their high specific surface area means that they have a high number of sites available per mass unit for the adsorption of pollutants on their surface. Therefore, nano-based adsorption processes have the potential to be integrated with conventional water treatment technologies, in order to increase the efficiency of depollution practices by adsorption. This means that there is thus the potential to further improve the performance of carbon-based adsorbents by using nanoscale carbon materials.

A lot of research has been performed with the focus on developing nanomaterials (not limited to carbon) as efficient adsorbents for the removal of different pollutants, including PAHs. A wide range of nanoadsorbents have been investigated in recent years [6], leading to even more attention being paid to adsorption processes. The nanomaterials that have been studied for PAH adsorption include carbon-based materials (e.g., carbon nanotubes, graphene, graphene oxide) [70], metal-based materials (such as silver, titanium dioxide, iron oxide and
zero-valent iron nanoparticles, etc.) [6], polymers [6], zeolites [6] and composite materials [6].

Regarding carbon nanomaterials, graphene can be particularly interesting because its separated layers can ensure a lot of surface for adsorption. It has already been tested for PAH removal and appears promising for such applications [41]. To easily recover graphene from water after adsorption processes, it is possible to add a magnetic component to the adsorbent, such as the well-know, green and cheap iron oxide, creating an (ideally) efficient and easily separable magnetic graphene composite. Studies have already been done in the literature regarding the applicability of such composites for PAH adsorption [59, 60, 71–78], as well as the graphene [17, 79–85] and magnetic iron oxide [86–93] counterparts. Thus, the next sections will describe such type of adsorbents, before discussing their performance for PAH removal in the works found in the literature.

## 5. Magnetic iron oxide/graphene nanocomposites: properties, synthesis and their advantages in PAH depollution practices

Out of the different nanomaterials that have potential for adsorption applications [6], carbon-based nanomaterials such as graphene, magnetic nanomaterials, and their composites appear as some of the most promising nanomaterials that can be used for removal of PAHs or other organic pollutants. On one hand, among carbon based materials, graphene and its derivatives are well-known to be able to efficiently adsorb many organic pollutants [41], as will be further described in this section. On the other hand, magnetic nanomaterials, in particular iron oxide, are promising because they can be functionalized for capturing PAHs but above all because of their easy separation after adsorption under application of an external magnetic field.

In this section, graphene, iron oxide nanomaterials and their composites will thus be described, in terms of their properties and synthesis routes.

### 5.1. Graphene and its derivatives

Graphene is a 2D material that consists of a single layer of sp² carbon atoms arranged in a hexagonal honeycomb structure [52,94,95]. Graphene is the building block of graphite and all graphitic forms [95] and it has distinct properties from other carbons nanomaterials (like carbon nanotubes), such as its exceptional strength [95,96]. Furthermore, apart from pristine graphene, it is possible to have graphene oxide (GO), which is prepared by the oxidation of graphite [52] and has oxygenated groups (such as carboxylic and hydroxyl groups) in the edges or in other regions of graphene layers, and reduced graphene oxide (RGO), prepared from the reduction of GO [52].

In particular, graphene has strong affinity with many different types of organic and inorganic pollutants, such as pharmaceuticals, pesticides, dyes, aromatic compounds (including PAHs) and heavy metals [97,98]. This strong affinity with organic pollutants has been explained by π-π

stacking interactions between the aromatic rings of graphene and the organic pollutant molecules exhibiting benzyl groups or alkene functions [6,7]. Coupled with its high theoretical specific surface area (~2630 m²/g [99,100]), high hydrophobicity and porosity, this makes it a promising material for adsorption applications. However, the main disadvantage of graphene, GO and RGO in depollution practices is that in water they tend to agglomerate, thus reducing the surface available for adsorption, and it is also difficult to separate and recover them from water after the adsorption process [6,7].

The structure, preparation methods and key characteristics of graphene are summarised in Fig. 3. One of the most used methods to prepare graphene corresponds to oxidation-reduction processes [101], but many other methods exist. For a more detailed description on synthesis processes, see references [95,96,98,101,102].

### 5.2. Iron oxide nanoparticles

In the last decades, a lot of attention has been given to magnetic iron oxide nanoparticles [103]. Iron oxides are compounds made of Fe and O, and are widely abundant, differing in their composition, crystal structure and iron valence [104]. They are even more interesting as nanomaterials, since their size reduction to the nanoscale leads to different properties from the bulk iron oxides and especially to a superparamagnetic behaviour. Spinel iron oxide nanoparticles consist mostly of magnetite (Fe₃O₄) and maghemite (γ-Fe₂O₃) phases [104,105]. The magnetite phase is the most magnetic phase of all transition metal oxides, with a saturation magnetisation of 92 emu/g at 300 K [104]. However at the nanoscale, magnetite nanoparticles are very sensitive to oxidation and thus most of time, they are oxidized at their surface with an oxidation degree depending on the size of nanoparticles [106,107]. The maghemite phase is the full oxidised form of magnetite and is less magnetic than magnetite, with a saturation magnetisation of 74 emu/g at 300 K [104].

The structure of magnetite, preparation methods and key characteristics of this type of nanoparticles are summarised Fig. 4.

The properties of iron oxide nanoparticles (such as shape, size, and hydrophilicity/hydrophobicity) are dependent on their synthesis route [108]. The main chemical synthesis routes used nowadays are the coprecipitation method (without or with additional hydrothermal treatment), thermal decomposition and the polyol method [109]. These nanoparticles have been largely described in the literature (see for example, references [103,105,109,110]).

Considering adsorption applications, one important aspect to discuss is the surface of iron oxide NPs. The surface of bare iron oxide is covered with hydroxyl groups, as is the general case for metal oxides, due to water chemisorbed at the surface, thus forming Fe-OH groups [104, 112]. In aqueous environment, a part of these surface hydroxyl groups are amphoteric and can lead to positive or negative surface charges depending on the pH, due to protonation or deprotonation of some hydroxyl surface groups, and thus making them amphoteric solids [104,

<figure><img src="image_5.png" /><figcaption>Fig. 3. The structure, preparation methods and key characteristics of graphene [95,96,98,101,102].</figcaption></figure>
<figure><img src="image_6.png" /><figcaption>Fig. 4. The structure, preparation methods and key characteristics of iron oxide nanoparticles (Magnetite crystal structure obtained with software Vesta [111]).</figcaption></figure>

[112] able to act both as a base and as an acid.

In order to know when iron oxides are positively or negatively charged, it is important to consider their isoelectric point, which corresponds to the pH when the surface has as many negative as positive charges. For iron oxide, the isoelectric point is at around 7 [113], and, below this pH, iron oxide NPs are positively charged, while, above this pH, they are negatively charged [109,112] as illustrated in Fig. 5.

Other molecules with certain functional groups (such as carboxylate or ammonium) can thus interact with the charged surface of iron oxides through electrostatic interactions [109]. Moreover, these hydroxyl groups are chemically reactive species that can react with other molecules such as phosphonates and form strong chemical bonds [104,109, 114,115]. The chemical surface groups of iron oxide are thus very important because they allow their functionalisation, as well as they can allow iron oxide NPs to adsorb certain pollutants through electrostatic interactions or chemisorption.

Iron oxide nanoparticles are thus interesting for water depollution applications, firstly because their magnetic behaviour makes it possible to recover the adsorbents very easily after adsorption by magnetic separation [52] in the presence of an external magnetic field. This easy and efficient recovery of the adsorbents is important to ensure that no residual nanoadsorbents are released into the environment. Because of their superparamagnetic behaviour, they can form stable suspensions in liquids, with no magnetic-induced agglomeration. However, they might still agglomerate at water pH if their surface is not functionalised, since their isoelectric point is around 7 [113].

Furthermore, they can be easily functionalised with suitable surface groups for targeted adsorption of organic pollutants. For the particular case of PAHs, they can be functionalized, for example, with organic moieties such as benzene or alkene functions, which can interact with PAHs through π-π stacking.

Finally, their small size and thus high specific surface area means that they have a lot of surface available for potential adsorption. In fact, some literature works have reported on the potential of bare iron oxide for the adsorption of PAHs [89,92,93] (as will be discussed later),

though the adsorption mechanism remains unclear.

### 5.3. Magnetic iron oxide/graphene nanocomposites

Graphene (or its derivatives) can be combined with iron oxide nanoparticles to have magnetic iron oxide/graphene composites nanomaterials for the removal of PAHs or depollution practices in general. This allows to have a composite nanomaterial that couples the properties of graphene and iron oxide nanoparticles (IONPs) and enables to overcome the limitations of each of these two nanomaterials. Like this, graphene is present in the nanocomposite for its highly efficient removal of PAHs, while IONPs are present to ensure the easy magnetic separation of the adsorbents once the adsorption process, i.e., the remediation, is finished. Furthermore, the presence of iron oxide nanoparticles around graphene prevents the graphene layers from stacking. Thus, the two main drawbacks of graphene adsorbents, which are the hard recovery and easy agglomeration, are overcome in the composite. Similarly, while IONPs alone might be found lacking in the efficiency of adsorption, the presence of graphene ensures it. However, the main disadvantage is that IONPs can occupy active sites on the surface of graphene, which might decrease slightly its adsorption capacity.

Composite nanomaterials can generally be produced by two types of methods: ex-situ and in-situ [116]. In the ex-situ methods, the already manufactured nanomaterials are merged to form a composite by processes such as ball milling, thermal curing, etc. [116]. In-situ methods correspond to the chemical synthesis of the magnetic nanostructures in a media where the other material is already dispersed [116]. In the case of iron oxide/graphene nanocomposites, these in-situ methods are more commonly used, in particular the hydrothermal or solvothermal routes for the synthesis of iron oxide NPs in a liquid medium where graphene or its derivatives are suspended [116]. For example, the polyol method has been used numerous times to prepare graphene/iron oxide composites for different applications [117,118].

<figure><img src="image_7.png" /><figcaption>Fig. 5. Surface charge of iron oxides depending on solution pH.</figcaption></figure>
# 6. Removal of PAHs from water by magnetic iron oxide/graphene nanocomposites

Magnetic iron oxide/graphene composites nanomaterials have been widely reported in the literature for their use in the removal and extraction of PAHs, as well as their individual components graphene (and its derivatives) and iron oxide NPs. These works will be presented and discussed in this section.

## 6.1. Performance of graphene and its derivatives for PAH removal

A considerable number of works has been published on the use of graphene for the adsorption of PAHs from water solutions, either in controlled laboratory conditions (Table 3) [17,71,79–85,119–122] or with real water samples spiked with pollutants (Table 4) [123–126]. Some of the more relevant results in terms of removal efficiency or adsorption capacity are presented in Fig. 6.

These works used graphene in a varied range of morphologies, including graphene layers [71,124], graphene wool [83,84,119,120] and graphene quantum dots [122]. Graphene derivatives have also been studied for this purpose, both GO [17,80,81,85,125] and RGO [81,85]. The use of these derivatives has been justified mostly by their highest hydrophilicity and thus easier dispersion in water [80,125]. Additionally, some works have also been carried out studying their functionalisation and doping with specific moieties, such as sulphur [79,122] and nitrogen [85,122], organic moieties [80,82] or even inorganic materials such as manganese oxide [121] or silver nanoparticles [120]. At this stage, the incorporation of iron oxides is not discussed because it will be later detailed in length. On the other hand, the incorporation of graphene in solid phase extraction procedures has also been reported [71, 123–126] to extract the analytes in a large volume of water and then elute them in a small volume of the appropriate solvent, enabling the analytes to be concentrated before analysis.

In terms of experimental conditions, these were very different from work to work. The adsorption of multiple PAHs by graphene nanomaterials has been tested, usually one or more of the US EPA 16 priority PAHs (either in single-solute experiments or in mixtures), but other times, PAH derivatives have also been studied (e.g. 1-naphthol [79], 9-anthracenemethanol [80], and 2-methylanthraquinone [85]). Moreover, a wide range of initial PAHs concentrations has been used in the different published works (from 0.1 µg/L [125] up to 350 mg/L [121]). Typically, lower concentrations were used when HMW-PAHs were present since they were practically insoluble in water, while higher concentrations were used when the PAH studied was more soluble in water, such as NAP [79] or FLE [80]. In order to overcome the limitations of low solubility, multiple authors have reported the use of a co-solvent, such as acetonitrile [17] or methanol [85], present in less than 1 % v/v in order to have no significant effects on the adsorption. The adsorbent concentration also varied in a large range of values from work to work, with the lowest being 4.2 mg/L [17] and the highest being 20 g/L [82]. These facts are important to remark because adsorption processes depend both on the quantity of pollutant (adsorbate) and quantity of adsorbent present in the aqueous media. Furthermore, these two concentrations are used to calculate the adsorption capacity (q) of a material and caution should be used when comparing q values obtained under different experimental conditions.

For single PAH experiments, works studying the adsorption of both LMW and HMW-PAHs have been published, but a bigger number of works was found on PAHs with lower molecular size since the lower solubility of HMW-PAHs implies the use of a co-solvent or the use of very sensitive analytical techniques to quantify PAH concentration before and after adsorption.

Zhao et al. studied the adsorption of naphthalene (2 rings) and its derivative 1-naphthol (NAP-OH) using sulfonated graphene in batch experiments at room temperature and neutral pH [79]. They reported that by increasing the graphene concentration from 50 mg/L up to

300 mg/L, the removal efficiency increases from ~40 to ~99 %, for NAP (see Fig. 6(a) – yellow square symbol), and is slightly lower, from ~38 to ~96 %, for NAP-OH [79]. In terms of maximum adsorption capacity, the values reported were converted to mg/g for comparison with other works and the qmax was higher for NAP-OH than for NAP (347 vs 298 mg/g, respectively). Nevertheless, the value obtained for NAP was still one of the highest reported for PAHs with graphene-based adsorbents with considerably low amount of adsorbent (see Fig. 6(b)). Interestingly, when looking at the adsorption capacity in mmol/g (Table 3), it is visible that the adsorbent had similar adsorption capacity for the two pollutants (2.407 mmol/g for NAP-OH and 2.326 mmol/g for NAP) [79], suggesting that the presence of the hydroxyl group had no significant impact (positive or negative) in their adsorption at pH 7.

Sun et al. also studied the adsorption of NAP in water using GO and RGO and comparing it with graphite as well [81]. At neutral pH, they obtained a removal efficiency of ~25, ~60 and ~95 % for graphite, GO, RGO, respectively (see Fig. 6(a) – yellow circle symbol) [81]. The much lower RE obtained with graphite compared with GO and RGO shows clearly the advantage of graphene-based nanomaterials (GO and RGO) in comparison with their bulk counterpart (graphite), possibly due to their higher specific surface area. Furthermore, RGO showed higher adsorption of NAP than GO [81] and the authors reported that this was due to different structure and surface characteristics of the two adsorbents [81]. The reduction of GO to RGO leads to the loss of oxygen functions in the graphene layers and decrease of defects in the sp² carbon network, and thus it is possible that there is more availability for π-π interactions between the aromatic rings of graphene and those of NAP, which led to the increase of the adsorption of this compound. The authors also determined the qmax for NAP on RGO and obtained ~767 mg/g [81], which was 2.6 times higher than that obtained by Zhao et al. for NAP (298 mg/g) [79] (however the adsorbent and NAP concentrations were different) and it was the highest value reported, to the best of our knowledge, for PAH adsorption with graphene materials (see Fig. 6(b). Sun et al. also studied the adsorption of other PAHs, namely PYR (4 rings), and again observed that RGO was a better adsorbent than GO and graphite [81]. Additionally, the authors also studied the adsorption isotherms with RGO for a third PAH which was ANT (3 rings), in order to see the influence of the aromatic ring number in the adsorption performance (see Fig. 6(b) - yellow, light blue and red circles). The adsorption capacity with the best adsorbent (RGO) was much higher for NAP (~767 mg/g) than for PYR (~198 mg/g) and ANT (~91 mg/g) [81]. The authors attribute this to the surface properties of RGO and to the properties of the studied PAHs but it is important to note that the initial PAH concentration used was much higher with NAP. Nonetheless, the lowest adsorption with RGO occurs for ANT (3 rings) and not for PYR (4 rings), and the authors attributed this to the lowest water solubility of ANT [81] (0.0432 mg/L vs 0.135 mg/L for PYR – see Table 1). However, we could also expect the opposite (better adsorption of ANT than PYR) since lower solubility would lead the hydrophobic molecules to bound preferentially to the adsorbent in suspension instead of staying in solution. In fact, this could also be related to the surface properties of RGO and the different molecular conformation of PAHs. The authors report that porosity and aggregation of RGO affect the chemical diffusion of big organic molecules [81] so this could explain why the adsorption of PYR (with 2 × 2 rings) was better than ANT (3 linear fused rings).

Similarly, the published work of Wang et al. reported the adsorption of different PAHs (NAP, PHE and PYR exhibiting 2, 3 and 4 aromatic rings, respectively) with two different adsorbents (graphene nanosheets (GNS) and GO) [17] (see Fig. 6(b) – up pointing triangle). They observed that GNS was a much better adsorbent for the three studied PAHs, even though it was used in much lower concentrations (qmax=126–442 mg/g for GNS vs. 3.67–7.81 mg/g for GO) [17]. The authors explained the higher adsorption capacity of GNS by the presence of more π-π interactions between GNS and the PAHs, as well as a “sieving effect” caused by the presence of wrinkles in the graphene layers [17].
<caption>**Table 3**</caption>

Published works on the adsorption of PAHs from artificial water (under controlled laboratory conditions) by graphene and its derivatives (GO, RGO, etc.), presented from oldest to newest. The nomenclature is defined at the bottom of the table<sup>a</sup>.

<table><thead><tr><th>Adsorbent<sup>a</sup></th><th>Ads. SSA (m<sup>2</sup>/g)</th><th>Ads. conc. (mg/L)<sup>b</sup></th><th>PAH<sup>a</sup></th><th>PAH conc. (µg/L)</th><th>Adsorption capacity (q<sub>max</sub>, mg/g)<sup>c</sup>/ Removal efficiency (RE, %)/ Recovery (%)<sup>d</sup></th><th>Recyclability</th><th>Process (Ads. models<sup>a</sup>)</th><th>Analytical method<sup>e</sup></th><th>Ref.</th></tr></thead><tbody><tr><td rowspan="3">Sulfonated graphene</td><td rowspan="3">520</td><td>50</td><td rowspan="3">NAP</td><td>26 × 10<sup>3</sup> (0.2 mmol/L)</td><td>RE: ~40<br>RE: ~99</td><td rowspan="3">-</td><td rowspan="3">Batch (L)</td><td rowspan="3">UV (220 nm)</td><td rowspan="3">[79]</td></tr><tr><td>300</td><td>(13–50)× 10<sup>3</sup></td><td>q<sub>max</sub>: ~298 (2.326 mmol/g)</td></tr><tr><td>40</td><td>29 × 10<sup>3</sup> (0.2 mmol/L)</td><td>RE: ~38<br>RE: ~96</td><td>UV (290 nm)</td></tr><tr><td rowspan="3">Graphene</td><td rowspan="3">-</td><td>50</td><td rowspan="3">NAP-OH</td><td>29 × 10<sup>3</sup> (0.2 mmol/L)</td><td>q<sub>max</sub>: ~347 (2.407 mmol/g)</td><td rowspan="3">-</td><td rowspan="3">Batch</td><td rowspan="3">GC-MS</td><td rowspan="3">[71]</td></tr><tr><td>300</td><td>(25–100)× 10<sup>3</sup></td><td>Max. Recovery: ~80<br>Max. Recovery: ~87<br>Max. Recovery: ~70</td></tr><tr><td>80</td><td>1</td><td>Max. Recovery: ~80</td><td></td></tr><tr><td rowspan="3">GO</td><td rowspan="3">~110</td><td>24</td><td rowspan="3">FLE<br>ANT-MeOH</td><td>10 × 10<sup>3</sup></td><td>RE: 62</td><td rowspan="3">-</td><td rowspan="3">Batch</td><td rowspan="3">UV-Vis Spectrophotometer</td><td rowspan="3">[80]</td></tr><tr><td>10 × 10<sup>3</sup></td><td>RE: 61.3</td></tr><tr><td>10 × 10<sup>3</sup></td><td>RE: 93.2 / q: ~447<br>RE: 72 / q: ~349</td></tr><tr><td rowspan="3">BBGO</td><td rowspan="3">~140</td><td>25</td><td rowspan="3">FLE<br>ANT-MeOH</td><td>10 × 10<sup>3</sup></td><td>RE: ~25</td><td rowspan="3">-</td><td rowspan="3">Batch (DA)</td><td rowspan="3">HPLC-UV</td><td rowspan="3">[81]</td></tr><tr><td>12.8 × 10<sup>3</sup></td><td>q<sub>max</sub>: ~12 (0.058 mmol/g)</td></tr><tr><td>1.8–44.6</td><td>RE: ~60<br>q<sub>max</sub>: ~96 (0.476 mmol/g)</td></tr><tr><td rowspan="3">Graphite</td><td rowspan="3">-</td><td>400</td><td rowspan="3">NAP<br>PYR</td><td>12.8 × 10<sup>3</sup></td><td>RE: ~95</td><td rowspan="3">-</td><td rowspan="3">Batch</td><td rowspan="3">HPLC-UV</td><td rowspan="3">[81]</td></tr><tr><td>1.8–44.6</td><td>q<sub>max</sub>: ~767 (5.981 mmol/g)</td></tr><tr><td>1.8–44.6</td><td>RE: ~95<br>q<sub>max</sub>: ~198 (0.979 mmol/g)</td></tr><tr><td rowspan="3">GO</td><td rowspan="3">-</td><td>200</td><td rowspan="3">NAP<br>PYR</td><td>12.8 × 10<sup>3</sup></td><td>RE: ~60<br>q<sub>max</sub>: ~96 (0.476 mmol/g)</td><td rowspan="3">-</td><td rowspan="3">Batch</td><td rowspan="3">HPLC-UV</td><td rowspan="3">[81]</td></tr><tr><td>1.8–44.6</td><td>RE: ~60<br>q<sub>max</sub>: ~96 (0.476 mmol/g)</td></tr><tr><td>200</td><td>12.8 × 10<sup>3</sup> (0.026–30) × 10<sup>3</sup></td><td>RE: ~95<br>q<sub>max</sub>: ~767 (5.981 mmol/g)</td></tr><tr><td rowspan="3">RGO</td><td rowspan="3">-</td><td>200</td><td rowspan="3">NAP<br>PYR<br>ANT</td><td>12.8 × 10<sup>3</sup></td><td>RE: ~95<br>q<sub>max</sub>: ~767 (5.981 mmol/g)</td><td rowspan="3">-</td><td rowspan="3">Batch</td><td rowspan="3">HPLC-UV</td><td rowspan="3">[81]</td></tr><tr><td>1.8–44.6</td><td>RE: ~95<br>q<sub>max</sub>: ~198 (0.979 mmol/g)</td></tr><tr><td>2.0–129.4</td><td>RE: ~95<br>q<sub>max</sub>: ~198 (0.979 mmol/g)</td></tr><tr><td rowspan="3">GNS</td><td rowspan="3">392</td><td>25</td><td rowspan="3">NAP<br>PHE<br>PYR</td><td>n.m.</td><td>q<sub>max</sub>: 126<br>q<sub>max</sub>: 142</td><td rowspan="3">-</td><td rowspan="3">Batch (DA, P2O)</td><td rowspan="3">HPLC-FLD</td><td rowspan="3">[17]</td></tr><tr><td>4.2</td><td>q<sub>max</sub>: 442</td></tr><tr><td>4.2</td><td>q<sub>max</sub>: 3.67</td></tr><tr><td rowspan="3">GO</td><td rowspan="3">236</td><td>250</td><td rowspan="3">NAP<br>PHE<br>PYR</td><td>n.m.</td><td>q<sub>max</sub>: 3.67<br>q<sub>max</sub>: 6.39</td><td rowspan="3">-</td><td rowspan="3">Batch</td><td rowspan="3">HPLC-FLD</td><td rowspan="3">[82]</td></tr><tr><td>45.5</td><td>q<sub>max</sub>: 7.81</td></tr><tr><td>9.1</td><td>RE: 69.7–89.7</td></tr><tr><td>GO/<br>polyHIPEs</td><td>-</td><td>20 × 10<sup>3</sup></td><td>NAP, FLU, PHE, ANT, FLE, PYR</td><td>20–200</td><td>RE: 77.8–97.1<br>q<sub>max</sub>: 47.5 (Total PAHs)</td><td>10 times</td><td></td><td></td><td></td></tr><tr><td rowspan="3">RGO/<br>polyHIPEs</td><td rowspan="3">10.68</td><td>20 × 10<sup>3</sup></td><td>BaA, CHY, BbF, BaP, BkF, DahA, BghiP</td><td>20–200<br>40–400</td><td>RE: 77.8–97.1<br>q<sub>max</sub>: 47.5 (Total PAHs)</td><td>10 times</td><td></td><td></td><td></td></tr><tr><td>670</td><td>PHE</td><td>1000–5000<br>0.3–0.8</td><td>q<sub>max</sub>: 5 / RE: 98.5<br>REcovery: ~85–97</td><td>-</td><td>Batch (S, P2O, Endo)</td><td>Spectro-fluorometer</td><td>[83]</td></tr><tr><td>670</td><td>PYR</td><td>1000–5000<br>0.3–0.8</td><td>q<sub>max</sub>: 20 / RE: 99.9<br>REcovery: ~80–100</td><td>8 times</td><td>Batch (S, P2O, Endo)</td><td>Spectro-fluorometer</td><td>[83]</td></tr><tr><td rowspan="5">GW</td><td rowspan="5">279</td><td>1000</td><td rowspan="5">Mix 15 PAHs</td><td>400–500</td><td>RE: 52.2–72.6<br>RE: 55.1–80<br>RE: 65.3–85.1<br>RE: 80–100</td><td>-</td><td>Batch (F)</td><td>GC-MS</td><td>[84]</td></tr><tr><td>1250</td><td>q<sub>max</sub>: 9.40 × 10<sup>-4</sup> – 4.91<br>REcovery: 45–75</td><td>6 times</td><td>Batch (F)</td><td>GC-MS coupled with SPE</td><td>[84]</td></tr><tr><td>1500</td><td>q<sub>max</sub>: 4.89 × 10<sup>-2</sup>/ Recovery: ~56</td><td>-</td><td>Batch (single solute; F, Exo/ Endo)<sup>g</sup></td><td>Spectro-fluorometer</td><td>[84]</td></tr><tr><td>2500</td><td>q<sub>max</sub>: 12.2 / Recovery: ~23</td><td>-</td><td>Batch (single solute; F, Exo/ Endo)<sup>g</sup></td><td>Spectro-fluorometer</td><td>[84]</td></tr><tr><td>1000</td><td>q<sub>max</sub>: 12.2 / Recovery: ~23<br>q<sub>max</sub>: 2.35 / Recovery: ~36</td><td>-</td><td>Batch (single solute; F, Exo/ Endo)<sup>g</sup></td><td>Spectro-fluorometer</td><td>[84]</td></tr><tr><td rowspan="3">GW</td><td rowspan="3">-</td><td>500</td><td rowspan="3">Mix 15 PAHs</td><td>100–500</td><td>q<sub>max</sub>: 1.52 / Recovery: ~5.5</td><td>-</td><td>Batch</td><td>Spectro-fluorometer</td><td>[84]</td></tr><tr><td>100–500</td><td>q<sub>max</sub>: 16.0 / Recovery: ~6.5</td><td>-</td><td>Batch</td><td>Spectro-fluorometer</td><td>[84]</td></tr><tr><td>100–500</td><td>q<sub>max</sub>: 34.48 × 10<sup>-2</sup> / RE: 95.8<br>RE (with NOM)<sup>h</sup>: 16.5–75.1</td><td>-</td><td>Batch</td><td>Spectro-fluorometer</td><td>[84]</td></tr><tr><td rowspan="3">GW</td><td rowspan="3">-</td><td>1000</td><td rowspan="3">Mix 15 PAHs</td><td>100–500</td><td>q<sub>max</sub>: 5.9 × 10<sup>-2</sup> / RE: 94.8</td><td>-</td><td>Batch (S/F, Exo/ Endo)<sup>h</sup></td><td>Spectro-fluorometer</td><td>[120]</td></tr><tr><td>1000</td><td>q<sub>max</sub>: 13.7 / RE: 98.7</td><td>-</td><td>Batch (S/F, Exo/ Endo)<sup>h</sup></td><td>Spectro-fluorometer</td><td>[120]</td></tr><tr><td>1000</td><td>q<sub>max</sub>: 74.1</td><td>-</td><td>Batch (L)</td><td>SPE-GC-MS</td><td>[121]</td></tr></tbody></table>

(continued on next page)
<table><caption>Table 3 (continued)</caption><thead><tr><th>Adsorbent<sup>a</sup></th><th>Ads. SSA (m<sup>2</sup>/g)</th><th>Ads. conc. (mg/L)<sup>b</sup></th><th>PAH<sup>a</sup></th><th>PAH conc. (µg/L)</th><th>Adsorption capacity (q<sub>max</sub>, mg/g)<sup>c</sup>/ Removal efficiency (RE, %)/ Recovery (%)<sup>d</sup></th><th>Recyclability</th><th>Process (Ads. models<sup>e</sup>)</th><th>Analytical method<sup>a</sup></th><th>Ref.</th></tr></thead><tbody><tr><td>C<sub>16</sub>-GQDs</td><td>0.3364</td><td>500</td><td>PHE</td><td>50–10 × 10<sup>3</sup></td><td>q<sub>max</sub>: 103.0 (q<sub>max</sub>, exp: ~14) RE: ~47</td><td>-</td><td>Batch (S, Endo/ Exo)<sup>f</sup></td><td>Spectro-fluorometer</td><td>[122]</td></tr><tr><td>C<sub>16</sub>-NGQDs</td><td>0.4021</td><td>500</td><td></td><td>5000</td><td>q<sub>max</sub>: 874.4 (q<sub>max</sub>, exp: ~16) RE: ~44</td><td></td><td></td><td></td><td></td></tr><tr><td>C<sub>16</sub>-SNGQDs</td><td>0.9856</td><td>500</td><td></td><td>5000</td><td>q<sub>max</sub>: 1377.1 (q<sub>max</sub>, exp: ~16) RE: ~50</td><td></td><td></td><td></td><td></td></tr><tr><td>GO</td><td>31.4</td><td>100</td><td>ANT 2-MAQ</td><td>1000</td><td>q<sub>E</sub>: 2.09 q<sub>E</sub>: 4.06</td><td>-</td><td>Batch (F, P1O/ P2O)<sup>g</sup></td><td>GC-MS coupled with liquid extraction</td><td>[85]</td></tr><tr><td>RGO</td><td>282</td><td>100</td><td>ANT 2-MAQ</td><td>1000</td><td>q<sub>E</sub>: 4.39 q<sub>E</sub>: 6.60</td><td></td><td></td><td></td><td></td></tr><tr><td>NRGO</td><td>200</td><td>100</td><td>ANT</td><td>1000</td><td>q<sub>E</sub>: 5.77 q<sub>E</sub>: ~0.8–6.5; RE: ~40–95</td><td>3 times</td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td>2-MAQ</td><td>50–1500</td><td>q<sub>E</sub>: 9.29 q<sub>E</sub>: ~0.4–13.9; RE: ~91–94</td><td>3 times</td><td></td><td></td><td></td></tr></tbody></table>

Interestingly, here, the adsorption capacity of both GNS and GO increased with the increase in the number of PAH aromatic rings (and decrease in water solubility), with the highest adsorption occurring for PYR [17]. This is the opposite of what had been described in the work of Sun et al. [81]. Wang et al. explain that these results are related to PAH concentration and, while at low concentrations, the adsorption capacity of the adsorbents follows the order NAP < PHE < PYR, at high PAH concentration it follows the order PYR < PHE < NAP [17]. The authors hypothesize that not only the typical π-π interactions and hydrophobic effect take place, but also other adsorption mechanisms become important [17]. Furthermore, they explain that the wrinkles on graphene layers are high energy adsorption sites and their presence depends on the chemical environment [17]. They noted that at higher PAH concentrations, the surface of graphene becomes flat without wrinkles, and they thus attribute the two adsorption behaviours to these morphological changes in graphene [17]. This could suggest that interactions between PAH and graphene (and derivatives) are dependant not only on the type of graphene material (graphene, GO, RGO, etc.) but also on their surface properties (both functionalization and morphology).

Song et al. also compared GO and RGO, as well as nitrogen doped RGO, in the adsorption of PAHs, in this case ANT and the derivative 2-methylanthraquinone (2-MAQ) (see Fig. 6 – red left triangle) [85]. Again, they reported better performance for RGO than for GO, and even better for nitrogen-doped RGO (NRGO) with adsorption capacities for ANT of 2.09, 4.06 and 5.77 mg/g for GO, RGO and NRGO respectively

[85]. The higher adsorption of NRGO was explained by the higher presence of mesopores, even though the reported SSA was lower than for RGO [85]. Furthermore, there was higher adsorption of the derivative 2-MAQ which the authors attributed to the π-donor character of NRGO as well as the π-acceptor behaviour of the aromatic rings in 2-MAQ, which allowed stronger interaction [85].

In terms of organic functionalisation, the group of Zhang et al. studied the effect of functionalising GO with an organic molecule called brilliant blue (BB), a triarylmethane dye, on the adsorption performance [80]. They reported that the presence of this dye led to an increase of removal efficiency of both FLE and the studied derivative anthracenemethanol (ANT-MeOH) [80]. The presence of this molecule on the surface, as well as the increase in specific surface area, could be responsible for the higher adsorption performance of BBGO. In the case of GO, no significant differences were obtained for FLE and ANT-MeOH, however with BBGO the adsorption of FLE was much higher, confirming that BB played an active role in the adsorption of FLE [80].

Adeola et al. has also reported multiple works focused mostly in single PAH experiments, using graphene wool (graphene deposited in quartz wool) as adsorbent [83,84,119,120] (see Fig. 6). They studied the removal of PHE and PYR using this graphene wool and obtained higher adsorption of pyrene (q<sub>max</sub>: 5 vs 20 mg/g for PHE and PYR, respectively) (see Fig. 6(b) – diamond symbols) [83]. However, they reported that at lower concentrations, the contrary was observed, i.e., higher adsorption of PHE [83]. This is the opposite of what was reported by Wang et al. for GNS and GO adsorbents [17], and could further confirm the importance

<footnote><sup>a</sup> Nomenclature: GO – graphene oxide; BBGO – brilliant blue functionalised graphene oxide; RGO – Reduced graphene oxide; GNS – graphene nanosheets; polyHIPEs – polymeric high-internal phase emulsions; GW – graphene wool; αAgNPs – oleylamine capped silver nanoparticles; NF – nanoflowers; C<sub>16</sub>-GQDs – hexadecylamine functionalised graphene quantum dots; C<sub>16</sub>-NGQDs – nitrogen doped C<sub>16</sub>-GQD; C<sub>16</sub>-SNGQDs – sulphur/nitrogen co-doped C<sub>16</sub>-GQD; NRGO – nitrogen-doped RGO; PAH – polycyclic aromatic hydrocarbon; NAP – naphthalene; NAP-OH – 1-naphthol; EPA – Environmental Protection Agency of the USA; PCB – polychlorinated biphenyl; FLE – fluoranthene; ANT-MeOH – anthracenemethanol; PYR – pyrene; ANT – anthracene; PHE – phenanthrene; ACE – acenaphthene; FLU – fluorene; BaA – benz(a)anthracene; CHY – chrysene; BbF – benzo(b)fluoranthene; BkF – benzo(k)fluoranthene; BaP – benzo(a)pyrene; DahA – dibenz(ah) anthracene; BghiP – benzo (ghi)perylene; 2-MAQ – 2-methylanthraquinone; L – Langmuir isotherm; F – Freundlich isotherm; S – Sips isotherm; DA – Dubinin-Astakhov; P1O – Pseudo-first order model; P2O – Pseudo-second order model; endo – endothermic; exo – exothermic; SPME – solid phase microextraction; GC – gas chromatography; MS – mass spectrometry; HPLC – high performance liquid chromatography; FLD – fluorescence detector; SPE – solid phase extraction.</footnote>

<footnote><sup>b</sup> in extraction experiments, if the adsorbent concentration was not directly given, it was calculated by diving the given mass by the solution volume.</footnote>

<footnote><sup>c</sup> q<sub>max</sub> values presented as reported in the work, generally corresponding to the isotherms' values. When this value was not given by the authors, values of the adsorption capacity at equilibrium (q<sub>E</sub>) were presented (if possible).</footnote>

<footnote><sup>d</sup> recovery determined based on the recovered amount of PAH after desorption.</footnote>

<footnote><sup>e</sup> organic matter (NOM) was added to study its impact on the adsorption performance of GW.</footnote>

<footnote><sup>f</sup> elution solvents presented correspond to the optimized conditions.</footnote>

<footnote><sup>g</sup> adsorption generally endothermic with exception of NAP adsorption that was exothermic.</footnote>

<footnote><sup>h</sup> Sips isotherm and endothermic for GW and Freundlich isotherm and exothermic for GW- αAgNP.</footnote>

<footnote><sup>i</sup> endothermic for C<sub>16</sub>-GQDs and exothermic for C<sub>16</sub>-NGQDs and C<sub>16</sub>-SNGQDs.</footnote>

<footnote><sup>j</sup> pseudo-first order adsorption for ANT and pseudo-second order adsorption for 2-MAQ.</footnote>
<caption>**Table 4**</caption>

Published works on the extraction of PAHs from PAH-spiked real water samples by graphene and its derivatives (GO, RGO, etc.), presented from oldest to newest. The nomenclature is defined at the bottom of the tableª.

<table><thead><tr><th>Adsorbent<sup>a</sup></th><th>Ads. SSA (m²/g)</th><th>Ads. conc. (mg/L)<sup>b</sup></th><th>PAH<sup>a</sup></th><th>PAH conc. (µg/L)</th><th>Relevant experimental conditions</th><th>Recovery (%)<sup>c</sup></th><th>Process<sup>a</sup></th><th>Analytical method<sup>a</sup></th><th>Ref.</th></tr></thead><tbody><tr><td>G-APTES</td><td>-</td><td>100</td><td>NAP, ACY, FLU, PHE, FLE, PYR</td><td>5</td><td>Rainwater samples<br/>River water samples</td><td>84.6–104.3<br/>92.3–109.5</td><td>SPE<br/>(Elution with ACN)</td><td>HPLC-UV</td><td>[123]</td></tr><tr><td>SLG-MMM</td><td>-</td><td>n.m.</td><td>NAP, ACE, FLU, ANT, PHE, FLE, PYR, BaA</td><td>25<br/>50<br/>100</td><td>Tap water samples<br/>Pond water samples</td><td>89.3–105.4<br/>10.0–100.5<br/>99.0–101.0<br/>99.0–100.5</td><td>Extraction (Elution with CH₂Cl₂: Acetone (1:1))</td><td>HPLC-UV</td><td>[124]</td></tr><tr><td>GO</td><td>-</td><td>100</td><td>16 EPA PAHs</td><td>0.1<br/>0.5<br/>0.1<br/>0.5</td><td>River water samples<br/>Seawater samples</td><td>80–110<br/>85–107<br/>85–109<br/>80–107</td><td>Extraction (Elution with hexane)</td><td>GC-MS</td><td>[125]</td></tr><tr><td>GNC-Co</td><td>-</td><td>n.m.</td><td>NAP, ACE, ACY, FLU, ANT, PHE, FLE, PYR</td><td>50</td><td>River water samples</td><td>88.8–115.7</td><td>SPME<sup>d</sup></td><td>GC-MS</td><td>[126]</td></tr></tbody></table>

<figure><img src="image_8.png" /><figcaption>**Fig. 6.** Compilation of relevant values of (a) removal efficiency and (b) adsorption capacity as a function of adsorbent concentration for the adsorption of PAHs in water with graphene-based materials [17,79,81,83–85,119,120]. Non-exhaustive compilation focused on the most studied PAHs (more details in Tables 3 and 4). Color-scheme indicates different PAHs and shape indicates different literature works as reported in the legend.</figcaption></figure>

of the graphene surface and morphology to the adsorption process. In another work, the impact of natural organic matter on PYR adsorption by graphene wool was evaluated [119] and they noticed that the removal efficiency greatly decreased in its presence [119] which is explained by changes in the solution chemistry as well as binding of the organic matter to graphene active sites [119]. Moreover, Adeola et al. also studied the adsorption of a mix of 15 PAHs [84]. They observed that removal efficiency increased with the increase of adsorbent concentration [84]. Additionally, they saw a decrease in adsorption capacities of PAHs in the mixture when compared to single solute adsorption experiments (9.40 × 10⁻⁴ – 4.91 mg/g for the mixture of PAHs vs 0.0489 – 12.2 mg/g for single solute experiments) [84], which is explained by the competitive adsorption occurring in the mixture of PAHs [84]. In single solute adsorption, higher qmax was obtained for BghiP (16.0 mg/g; 6 rings) and ANT (12.2 mg/g; 3 rings), followed by BaA (2.35 mg/g; 4 rings), BaP (1.52 mg/g; 5 rings) and NAP (4.89 × 10⁻² mg/g; 2 rings) [84] (see Fig. 6(b) – right triangle in red, dark blue and yellow for ANT, BaP and NAP), which indicates that adsorption did not follow a trend

based on solubility or number of aromatic rings (see Table 1). Finally, they also tested the functionalisation of GW with silver NPs for BaP adsorption [120] and saw a significant increase in qmax in the functionalized GW (from 5.9 × 10⁻² mg/g to 13.670 mg/g). They explained this behaviour by the increase of active sites as well as surface hydrophobicity, due to oleylamine/oleic acid coating on the silver NPs [120].

For mixtures of PAHs, polymer functionalised GO and RGO have also been tested [82] and again functionalised RGO showed better removal of PAHs, with an adsorption capacity for total PAHs of 47.5 mg/g [82]. Additionally, other works for mixtures of PAHs have reported testing graphene-based adsorbents for extraction techniques [71,123–127]. Most of these extraction tests considered much lower PAH concentrations, usually between 0.1 and 5 µg/L (see Table 3 and Table 4), with few exceptions [124,126]. Karamani et al. reported the results on the adsorption of a mix of PAHs, as well as a mix of PCBs and phthalic acids using graphene [71]. For a PAH concentration of 1 µg/L (2 and 4 µg/L for PCBs and phthalic acids, respectively), the recovery was only around 80 % (with 87 % for PCBs and 70 % for phthalic acids) [71].

<footnote>ª Nomenclature: G – graphene; APTES – aminopropyltriethoxysilane; SLG-MMM – Single layer graphene in mixed matrix membrane; GO – graphene oxide; GNC-Co – Metal and nitrogen doped graphene; PAH – polycyclic aromatic hydrocarbon; NAP – naphthalene; ACY – Acenaphthylene; FLU – Fluorene; PHE – Phenanthrene; FLE – fluoranthene; PYR – Pyrene; EPA – Environmental Protection Agency of the USA; ACE – acenaphthene; ANT – Anthracene; SS – stainless steel; ACN – acetonitrile; SPE – solid phase extraction; SPME – solid phase microextraction; GC – gas chromatography; MS – mass spectrometry; HPLC – high performance liquid chromatography; n.m. – not mentioned.</footnote>

<footnote>ᵇ in extraction experiments, if the adsorbent concentration was not directly given, it was calculated by diving the given mass by the solution volume.</footnote>

<footnote>ᶜ recovery determined based on the recovered amount of PAH after desorption.</footnote>

<footnote>ᵈ GNC-CO fixed on a stainless steel wire, directly injected into the GC inlet after PAH extraction.</footnote>
Interestingly, some works, after process optimisation, tested extraction processes in real water samples that were spiked with mixtures of PAHs (see Table 4). However, the number of these studies using real water (spiked with PAHs) was much smaller than the number of controlled-laboratory studies using purely artificial water. Huang et al. used graphene functionalised with triethoxysilane for the extraction of 6 PAHs from spiked rain water, river water and tap water samples and obtained recoveries higher than 84.6 % for all samples [123]. Mukhtar and See used single layer graphene incorporated in mixed matrix membranes for the extraction of a mixture of 8 PAHs spiked in sewage pond water at different concentrations [124]. They obtained recoveries generally around 100 %, except for ACE spiked at 25 µg/L where its recovery was only 10 % [124]. Han et al. reported on the extraction of the 16 US EPA priority PAHs spiked in river and seawater samples at very low concentrations (0.1 and 0.5 µg/L) using GO and its aggregation in solution [125]. Extraction recoveries ranged between 80 % and 110 % [125] and no clear relation was observed between the recovery and the number of rings and water solubility. Finally, Wang et al. studied the application of a composite of graphene-based metal and nitrogen-doped carbon (GNC-Co) [126] fixed on the surface of a stainless steel wire for SPME in river water samples spiked with a mix of 8 PAHs at a concentration of 50 µg/L and obtained recoveries 88.8 % for the studied PAHs [126].

To summarise, graphene and its derivatives have been widely tested for the removal of PAHs from water. Both batch adsorption experiments (Table 3) and extraction procedures from real water doped with PAH mixtures (Table 4) have been studied. A wide range of concentrations (both for the adsorbent and PAHs) has been used and both single-component and multi-component adsorbate mixtures have been tested. It seems that most of these works using graphene based materials focused either on PAH mixtures or on the study of the adsorption of PAHs with lower molecular weight (4 aromatic rings or less), mainly naphthalene, anthracene, phenanthrene and pyrene (Fig. 6), with very few works being found on the adsorption of 5-ring and 6-ring PAHs, namely BaP [84,120], BbF [121] or BghiP [84]. This is possibly due to the higher water solubility of smaller PAHs (up to pyrene in Table 1), which allows higher concentrations to be used. However, considering that HMW-PAHs are generally more dangerous to human health and the environment, even in lower concentrations, it is crucial to understand if they are efficiently removed from water using this type of adsorbents. This is particularly important considering that it seems that there is not a consensus in whether the removal efficiency increases or decreases with the increase in number of rings (and decrease in solubility), as seen by the different trends observed in multiple works reported and discussed here [17,81,83,84].

In terms of the adsorption mechanisms, different isotherms have been found to describe the adsorption process, such as Langmuir, Freundlich, Sips, and Dubinin-Astakhov model [17,79,81,83-85, 120-122]. The Freundlich and Sips isotherms were the most common, thus suggesting multilayer heterogeneous adsorption processes. Adsorption kinetics were not studied as often, but both pseudo-first order and pseudo-second order were used depending on the work [17, 83,85]. Finally, adsorption thermodynamics have been studied in particular by Adeola et al. [83,84,120]. Generally, adsorption of PAHs on graphene wool was found to be endothermic, except in the case of naphthalene, or in the case of BaP using silver NPs modified graphene wool. Similarly, for graphene quantum dots, adsorption of phenanthrene was found to be endothermic, but it became exothermic when these graphene quantum dots were doped with sulphur or nitrogen [122], which could suggest additional interaction mechanisms besides π-π stacking in doped materials.

The regeneration of graphene-based adsorbents was also studied [82-85], showing that these materials generally have the potential to be recycled and reused at least 3 and up to 10 times [82]. In all the works that performed adsorbent regeneration, solvent-aided extraction was the method of choice used to desorb the PAHs. Adsorbent regeneration was

performed with various different solvents, such as hexane [83,84] or ethanol [85], or mixtures of solvents such as acetonitrile: dichloromethane (1:1) [82].

## 6.2. Performance of iron oxide nanoparticles for PAH removal

Numerous works have been found in the literature on the use of iron oxide nanomaterials for the adsorption of PAHs because of their magnetic properties that allow easy separation. These works are summarised in Table 5 and Table 6, and relevant results will be further discussed. These include both batch experiments or so-called magnetic SPE/ extraction experiments (which differs from batch experiments because there is elution of the adsorbed PAHs and the eluted analytes are the ones quantified) [71,72,86-93,128-137]. These magnetic properties mean that they have the potential to be used both in the removal of PAHs, by implementation of large-scale magnetic separation techniques, such as magnetic drum belt separators, as well as in their quantification in the water media by techniques such as magnetic SPE. For an easier comparison, some of the more relevant results in terms of removal efficiency or adsorption capacity can be found in Fig. 7.

In the majority of cases, iron oxide nanomaterials are functionalised with appropriate organic moieties to promote the efficient PAH adsorption, such as fatty acids [87,132], dopamine-based molecules [86, 88,129,137], porphyrins [131], tannic acid [91], other aromatic groups [92,93,135], chitosan [130], quantum dots [122], gold [90,128] and others. Adsorbent concentration varied from 10 mg/L [89] up to 2 g/L [129], while for PAHs the concentration used was between 1 ng/L [128] and 12 mg/L [92].

In fact, there are not a lot of works focusing on the adsorption of PAHs directly by naked iron oxide nanomaterials. Hassan et al. performed interesting studies on the adsorption of BaP and PYR by green synthesised, non-functionalised iron oxide NPs [89] (see Fig. 7 – down pointing triangle). In their work, they tested different adsorbent concentrations and adsorption times and observed that, by increasing the amount of NPs from 10 to 130 mg/L, the removal efficiency increased from 20 % to 71.6 % for BaP and from 30 % to 67.8 % for PYR, with 30 min of contact time [89]. Furthermore, with 90 mg/L of iron oxide NPs, they were able to remove up to 99 % of BaP and 98 % of PYR by increasing the adsorption time to 150 min [89], leading to respective $q_{max}$ values of 0.029 and 2.8 mg/g [89]. Meng et al. studied the adsorption of phenanthrene using goethite ($\alpha$-FeOOH) and obtained maximum adsorption capacities ranging from 0.531 mg/g at 38°C up to 0.636 mg/g at 18°C [133], further indicating that iron oxides/hydroxides alone can adsorb PAHs, at least until a certain extent (see Fig. 7 (b) – green crossed circle symbol). Yang et al. presented in their work [72] results on the adsorption of NAP and its derivatives with iron oxide NPs as a comparison to the GO/iron oxide composite that they developed (that will be later discussed more in depth) [72]. Even though, iron oxide alone is not the focus of their work, they reported $q_{max}$ values between 73 and 89 mg/g for NAP and its derivatives [72] (see Fig. 7(b) – yellow hexagon symbol), which is not an insignificant value (but much lower than those obtained for the developed composite). Finally, the works of Wei et al. [92,135] also reported on the performance of naked $Fe_3O_4$ NPs on the removal of PHE and a derivative, even though their focus was the functionalised NPs (see Fig. 7 – green left triangle, square and star symbols). In one work, the $q_{max}$ obtained with $Fe_3O_4$ was in the same range of the system functionalised with chrolopropyltriethoxysilane (25.36 vs 24.81 mg/g) and only slightly lower than the benzhydrylamine functionalised iron oxide (25.36 vs 38.11 mg/g) [92]. In another two works [93,135], Wei et al. further improved the adsorption performance for both PHE and the derivative 9-PHE-OH by functionalising the iron oxide NPs with either a Schiff base polymer ($q_{max}$ = 65.87-72.89 mg/g) [135] or with terephthaldehyde ($q_{max}$ = 84.38-87.45 mg/g) [93].

Besides, Liao et al. reported the use of $Fe_3O_4$ nanoparticles coated with different fatty acids for the adsorption of BaP [87]. The presence of
<caption>**Table 5**</caption>

Published works on the adsorption of PAHs from artificial water (under controlled laboratory conditions) by bare and functionalised iron oxide, presented from oldest to newest. The nomenclature is defined at the bottom of the table¹.

<table><thead><tr><th>Adsorbent<sup>a</sup></th><th>Ads. SSA (m<sup>2</sup>/g)</th><th>Ads. Conc. (mg/L)<sup>b</sup></th><th>PAH<sup>c</sup></th><th>PAH conc. (µg/L)</th><th>Adsorption capacity (q<sub>max</sub>, mg/g)/ Removal efficiency (RE, %)/ Recovery (%)<sup>d</sup></th><th>Recyclability</th><th>Process (Ads. models<sup>e</sup>)</th><th>Analytical method<sup>a</sup></th><th>Ref.</th></tr></thead><tbody><tr><td>Fe<sup>0</sup>/iron oxide-oxyhydroxide</td><td>-</td><td>300</td><td>16 EPA PAH<br>Mix PCB<br>Phthalic acids</td><td>1<br>2<br>4</td><td>Max. Recovery: ~10<br>Max. Recovery: ~45<br>Max. Recovery: ~65</td><td>-</td><td>MSPE (Elution with CH<sub>2</sub>Cl<sub>2</sub>: hexane (9:1))<sup>d</sup></td><td>GC-MS</td><td>[71]</td></tr><tr><td>FeO•Fe<sub>2</sub>O<sub>3</sub></td><td>-</td><td>100</td><td>NAP<br>1NAP-A<br>NAP-OH</td><td>n.m.</td><td>q<sub>max</sub>: 73 (0.57 mmol/g)<br>q<sub>max</sub>: 82 (0.57 mmol/g)<br>q<sub>max</sub>: 89 (0.62 mmol/g)</td><td>-</td><td>Batch (L)</td><td>UV-Vis Spectrophotometer</td><td>[72]</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>/PDA</td><td>84</td><td>40</td><td>FLE, PYR, BaA, BbF, BaP, BghiP</td><td>0.005–0.200</td><td>Recoveries: 76–103</td><td>-</td><td>MSPE (Elution with ACN)</td><td>HPLC-FLD</td><td>[86]</td></tr><tr><td>PSA-Fe<sub>3</sub>O<sub>4</sub></td><td>-</td><td>100</td><td>BaP</td><td>10</td><td>q<sub>E</sub>: 9.069 × 10<sup>-3</sup></td><td>-</td><td>Batch</td><td>HPLC-FLD</td><td>[128]</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>/Au</td><td>-</td><td>100</td><td>BaP</td><td>10</td><td>q<sub>E</sub>: 96.607 × 10<sup>-3</sup><br>Recovery: 96.0–99.6</td><td>5 times</td><td>(F, P2O)</td><td></td><td></td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub></td><td>-</td><td>n.m.</td><td>NAP</td><td>n.m.</td><td>q<sub>E</sub>: 16.8 × 10<sup>-3</sup></td><td>-</td><td>Batch</td><td>HPLC-FLD</td><td>[87]</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>@C6</td><td>-</td><td>n.m.</td><td>ANT</td><td>n.m.</td><td>q<sub>E</sub>: 20.3 × 10<sup>-3</sup></td><td>-</td><td>Batch</td><td>HPLC-FLD</td><td>[87]</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>@C8</td><td>-</td><td>n.m.</td><td>BaP</td><td>2–10</td><td>RE: ~36<br>RE: ~60<br>RE: ~78<br>RE: ~90<br>RE: ~88</td><td>-</td><td>Batch (F)</td><td></td><td></td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>@C10</td><td>-</td><td>200</td><td>BaP</td><td>2–10</td><td>q<sub>max</sub>: 63.90 × 10<sup>-3</sup></td><td>-</td><td>Batch (F)</td><td></td><td></td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>@C11</td><td>-</td><td>200</td><td>BaP</td><td>2–10</td><td>Recoveries: 90.6–93.3</td><td>3 times</td><td>Batch (F/D-R, P2O)<sup>e</sup></td><td>HPLD-FLD</td><td>[129]</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>@C10</td><td>-</td><td>200</td><td>BaP</td><td>2–10</td><td>q<sub>E</sub>: 0.185<br>q<sub>E</sub>: 0.367</td><td>10 times</td><td>Batch (F)</td><td>Spectro-fluorometer</td><td>[130]</td></tr><tr><td>Fe@SiO<sub>2</sub>@PDA</td><td>58.97</td><td>2000<br>1000</td><td>PHE<br>ANT</td><td>500<br>400</td><td>RE: ~48<br>RE: ~96<br>RE: ~96</td><td>-</td><td>Batch (F)</td><td></td><td></td></tr><tr><td>MC<br>MC pyrolised at 550°C</td><td>-</td><td>500</td><td>ANT</td><td>50</td><td>RE: ~2 (NAP), ~53 (ACE), ~60 (ACY), ~87–95 (FLU, ANT, PHE, FLE, PYR)</td><td></td><td>Batch</td><td>HPLC</td><td>[88]</td></tr><tr><td>MC pyrolised at 800°C</td><td>-</td><td>50</td><td>NAP, ACE, ACY, FLU, ANT, PHE, FLE, PYR</td><td>2.2–44</td><td>RE: 57.0–100 (NAP to PYR); 0.0–3.5 (BaA to BghiP)<br>RE: 56.4–100 (NAP to PYR); 34.6–66.9 (BaA to BghiP)</td><td></td><td>Batch</td><td>HPLC-PDAD</td><td>[89]</td></tr><tr><td>MC pyrolised at 800°C</td><td>-</td><td>50</td><td>NAP, ACE, FLU, ANTE, ANT, FLE, PYR, BaA, CHY, BbF, BaP, BkF, IcdP, DahA, BghiP</td><td>10</td><td>RE: 20 (30 min)<br>RE: 71.6 (30 min)<br>RE: 99–79.5 (150 min)<br>q<sub>max</sub>: 0.029<br>RE: 30 (30 min)<br>RE: 67.8 (30 min)<br>RE: 98–78.3 (150 min)<br>q<sub>max</sub>: 2.8<br>Recovery: 85.6–118.7 (cycle 2); 100.9–119.6 (cycle 12)</td><td>5 times</td><td>Batch (L, P2O, Exo)</td><td>HPLC-FLD</td><td>[131]</td></tr><tr><td>Iron oxide NPs</td><td>10<br>130<br>90</td><td></td><td>BaP</td><td>1<br>1–4</td><td>RE: 8.4<br>RE: 79.7<br>q<sub>max</sub>: 59.89 × 10<sup>-3</sup><br>RE: 14.7<br>RE: 93.9<br>q<sub>max</sub>: 76.23 × 10<sup>-3</sup></td><td>5 times</td><td>Batch (L, P2O)</td><td>HPLC-FLD</td><td>[132]</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub></td><td>-</td><td>1000</td><td>NAP, ACE, FLU, ANTE, ANT, FLE, PYR, BaA, CHY, BbF, BaP, BkF, IcdP, DahA, BghiP</td><td>n.m.</td><td>RE: 0.5–14<br>RE: 0.1–75</td><td>6 times</td><td>Batch (L, P2O)</td><td>HPLC-FLD</td><td>[90]</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>-PBI-DA</td><td>-</td><td>1000</td><td>BbF, BaP, BkF, IcdP, DahA, BghiP</td><td>100–400</td><td>RE: 0.5–14<br>n.m.</td><td>3 times</td><td>Batch (L/Lin, Exo)<sup>f</sup></td><td>HPLC-UV</td><td>[133]</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>@SiO<sub>2</sub>@TCPP</td><td>-</td><td>500</td><td>ACY, ACE, FLU, PHE, ANT, FLE, PYR, BaA, CHY, BbF, BaP, BkF, IcdP, DahA, BghiP</td><td>10</td><td>RE: 8.4<br>RE: 79.7<br>q<sub>max</sub>: 59.89 × 10<sup>-3</sup><br>RE: 14.7<br>RE: 93.9<br>q<sub>max</sub>: 76.23 × 10<sup>-3</sup></td><td>6 times</td><td>Batch (L, P2O)</td><td>HPLC-FLD</td><td>[132]</td></tr><tr><td>C<sub>11</sub>-Fe<sub>3</sub>O<sub>4</sub></td><td>163.5</td><td>30<br>180<br>180</td><td>BaP</td><td>2</td><td>RE: 0.5–14<br>RE: 0.1–75</td><td>6 times</td><td>Batch (L, P2O)</td><td>HPLC-FLD</td><td>[132]</td></tr><tr><td>CDs/C<sub>11</sub>-Fe<sub>3</sub>O<sub>4</sub></td><td>289.8</td><td>30<br>180<br>180<br>n.m.</td><td>BaP</td><td>2</td><td>RE: 0.5–14<br>n.m.</td><td>6 times</td><td>Batch (L, P2O)</td><td>HPLC-UV</td><td>[133]</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>@SiO<sub>2</sub>@MMTA-Au</td><td>-</td><td>200<br>600</td><td>NAP, PHE, BaA, BaP, IcdP</td><td>0.1–75</td><td>RE: ~50–67<br>RE: ~88–96</td><td>6 times</td><td>Batch (L/Lin, Exo)<sup>f</sup></td><td>HPLC-UV</td><td>[90]</td></tr><tr><td>Goethite (α-FeOOH)</td><td>44.52</td><td>1000</td><td>PHE</td><td>0–614.7</td><td>q<sub>max</sub>: 0.636 (18°C)<br>q<sub>max</sub>: 0.567 (28°C)<br>q<sub>max</sub>: 0.531 (38°C)</td><td>6 times</td><td>Batch (L/Lin, Exo)<sup>f</sup></td><td>HPLC-UV</td><td>[133]</td></tr><tr><td>MNPs@G3.0</td><td>-</td><td>1000</td><td>PHE, ANT, FLE, PYR, BaP</td><td>10</td><td>RE: ~9–82<br>RE: ~52–97<br>RE: 78.0–98.6</td><td>6 times</td><td>Batch (L/Lin, Exo)<sup>f</sup></td><td>HPLC-UV</td><td>[134]</td></tr><tr><td>MNPs@G3.0@4-MBA</td><td>-</td><td>1000<br>n.m.</td><td></td><td></td><td>RE: ~98<br>RE: ~54</td><td>10 times</td><td>Batch</td><td>Spectro-fluorometer</td><td>[91]</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>-SiO<sub>2</sub>-TA</td><td>-</td><td>333</td><td>BghiP</td><td>5<br>50</td><td></td><td></td><td>Batch</td><td>Spectro-fluorometer</td><td>[91]</td></tr></tbody></table>

(continued on next page)
<table><caption>Table 5 (continued)</caption><thead><tr><th>Adsorbent<sup>a</sup></th><th>Ads. SSA (m<sup>2</sup>/g)</th><th>Ads. Conc. (mg/L)<sup>b</sup></th><th>PAH<sup>c</sup></th><th>PAH conc. (µg/L)</th><th>Adsorption capacity (q<sub>max</sub>, mg/g)/ Removal efficiency (RE, %)/ Recovery (%)<sup>c</sup></th><th>Recyclability</th><th>Process (Ads. models<sup>a</sup>)</th><th>Analytical method<sup>a</sup></th><th>Ref.</th></tr></thead><tbody><tr><td>Fe<sub>3</sub>O<sub>4</sub></td><td>35.55</td><td rowspan="3">200</td><td rowspan="3">DahA</td><td>5</td><td>RE: ~96</td><td rowspan="3">-</td><td rowspan="3">Batch (L/S/ To, P2O/ WM)<sup>b</sup></td><td rowspan="3">UV-Vis Spectrophotometer</td><td rowspan="3">[92]</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>-3-CHLOR</td><td>15.47</td><td>50</td><td>RE: ~29</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>-1.5BENZY</td><td>7.717</td><td>50</td><td>RE: ~94</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub></td><td>36</td><td rowspan="3">200</td><td rowspan="3">IcdP</td><td>5</td><td>RE: ~10</td><td rowspan="3">4 times</td><td rowspan="3">Batch (L/T/S, P2O)<sup>b</sup></td><td rowspan="3">UV-Vis Spectrophotometer</td><td rowspan="3">[135]</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>-BIPH-PHEN</td><td>25</td><td>50</td><td>RE: ~10</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>-TERE</td><td>17.98</td><td>8000</td><td>q<sub>max</sub>: 25.36</td><td>-</td></tr><tr><td></td><td></td><td>8000</td><td>q<sub>max</sub>: 24.81</td><td>-</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td></td><td>8000</td><td>q<sub>max</sub>:38.11</td><td>4 times</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td></td><td>8000</td><td>q: ~13.3</td><td>-</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td></td><td>8000</td><td>q: ~29.6</td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td></td><td>8000</td><td>q: ~34.1 / RE: 85.65</td><td>4 times</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td></td><td>8000</td><td>q<sub>max</sub>:65.87</td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td></td><td>8000</td><td>q: ~39.4 / RE: 98.52</td><td>4 times</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td></td><td>8000</td><td>q<sub>max</sub>:72.89</td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td></td><td>8000</td><td>q: 32.25 / RE: ~85</td><td>4 times</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td></td><td>8000</td><td>q<sub>max</sub>: 84.38</td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td></td><td>8000</td><td>q: 39.76 / RE: ~94</td><td>4 times</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td></td><td>8000</td><td>q<sub>max</sub>: 87.45</td><td></td><td></td><td></td><td></td><td></td><td></td></tr></tbody></table>

<sup>a</sup> Nomenclature: PDA – polydopamine; PSA – primary secondary amine functionalised magnetic silica microspheres; C<sub>6</sub> – hexanoic acid; C<sub>8</sub> – n-octanoic acid; C<sub>10</sub> – decanoic acid; C<sub>11</sub> – undecanoic acid; MC – magnetic chitosan; PBI-DA – perylene bismide – dopamine ligand; NPs – nanoparticles; TCPP – Meso-Tetra (4-carboxyphenyl) porphyrin; CDs – carbon dots; MMTA – 2-mercapto-4-methyl-5-thiazoleacetic acid; MNPs@G<sub>3</sub>.0 – Fe<sub>3</sub>O<sub>4</sub>@Polyamidoamine dendrimer; 4-MBA – 4-mercaptobenzoic acid; TA – tannic acid; CHLOR – 3-chloro-propyltriethoxysilane; BENZY – benzydrylamine; BIPH – 4,4'-biphenyldicarboxaldehyde; PHEN – m-phenylenediamine; TERÉ – terephthalaldehyde; PAH – polycyclic aromatic hydrocarbon; EPA – Environmental Protection Agency of the USA; PCB – Polychlorinated biphenyl; NAP – naphthalene; 1NAP-A – 1-naphthylamine; NAP-OH – 1-naphthol; BaP – benzo(a)pyrene; ANT – anthracene; ACE – acenaphthene; ACY – acenaphthylene; FLU – fluorene; PHE – phenanthrene; FLE – fluoranthene; PYR – Pyrene; ANTE – anthrene; BaA – benz(a)]anthracene; CHY – chrysene; BbF – benzo(b)fluoranthene; BkF – benzo(k)fluoranthene; IcdP – indeno(1,2,3-cd)pyrene;; DahA – dibenz(ah) anthracene; BghiP – benzo(ghi)perylene; 9-PHE-OL – 9-phenanthrol; L – Langmuir isotherm; F – Freundlich isotherm; Lin – linear isotherm; S – Sips isotherm; To – Toth isotherm; T – Temkin isotherm; DR – Dubinin-Radushkevich; P10 – Pseudo-first order model; P20 – Pseudo-second order model; WM – Weber-Morris diffusion model; exo – exothermic; MSPE – magnetic solid phase extraction; GC – gas chromatography; MS – mass spectrometry; HPLC – high performance liquid chromatography; FLD – fluorescence detector; PDAD – photodiode array detector.

<sup>b</sup> in extraction experiments, if the adsorbent concentration was not directly given, it was calculated by dividing the given mass by the solution volume.

<sup>c</sup> recovery determined based on the recovered amount of PAH after desorption.

<sup>d</sup> elution solvents presented correspond to the optimized conditions.

<sup>e</sup> for PHE, Freundlich isotherm at 298 K and D-R isotherm at 303 and 308 K; for ANT, Freundlich isotherm.

<sup>f</sup> linear isotherm at 18 and 38°C and Langmuir isotherm at 28°C.

<sup>g</sup> Langmuir, Sips and Toth isotherm all fit well for the isotherm experiments; pseudo-second order and Weber-Morris diffusion both fit well in the kinetics experiments.

<sup>h</sup> Langmuir, Temkin and Sips isotherms fit similarly to the data.

fatty acids increased the removal efficiency from 36 % in bare iron oxide NPs up to around 90 % using decanoic acid coating [87] and this efficiency increased with increase of fatty acid chain length reaching a maximum for a chain with 10 carbons [87]. The authors suggested that the adsorption of BaP occurred by Van der Waals interactions with the alkyl chain [87] and additionally the increase of hydrophobicity with increase of chain length might have also contributed to higher adsorption yield. With decanoic acid, they did further isotherm studies, obtaining a q<sub>max</sub> of 63.90 µg/g [87] (see Fig. 7(b) – dark blue pentagon). Yang et al. also studied the performance of Fe<sub>3</sub>O<sub>4</sub> coated with undecanoic acid for BaP removal [132] (see Fig. 7 – dark blue diamond symbol) and they saw an increase on removal efficiency with the increase of adsorbent concentration [132]. They showed a further increase in RE by adding carbon quantum dots to C<sub>11</sub>-Fe<sub>3</sub>O<sub>4</sub> [132]. In both works, the q<sub>max</sub> obtained was around 60–76 µg/g [87,132], which is in agreement with the previous discussed work of Liao et al. [87].

On the other hand, Li et al. used polydopamine to functionalise zero-valent iron@silica particles for removal of PHE and ANT [129] (see Fig. 7(b) – green and red right triangles, respectively). The adsorption capacity obtained in kinetic batch experiments was 0.185 mg/g for PHE and 0.367 mg/g for ANT [129], even though the adsorbent concentration was two times higher with PHE [129]. Thus, this indicated more efficiency to remove ANT, which has lower solubility in water and tends to confirm the behaviour observed when the solubility decreases in some works with graphene-based materials, as discussed in the previous section (see, for example, ref [17]). Wang et al. also used polydopamine functionalised Fe<sub>3</sub>O<sub>4</sub> NPs for the extraction of 6 PAHs by magnetic SPE

(MSPE) [86] and obtained recoveries between 76 % and 103 % [86]. Likewise, Alfarhani and Hammza coated iron oxide@silica particles with tannic acid and used this material for the extraction by MSPE of BghiP, DahA and IcdP in different concentrations [91]. At 10 µg/L of PAH concentration, the extraction efficiency of the three PAHs from their aqueous solution was similar (~94–98 %) but at 50 µg/L it decreased significantly (~54 for BghiP, ~29 for DahA and ~10 % for IcdP) [91]. Furthermore, Gutiérrez et al. used perylene bismide-dopamine functionalised iron oxide NPs to simultaneously extract 15 PAHs [88] (see Fig. 7(a) – circle symbol) and they observed that both naked iron oxide and functionalised NPs showed removal values between 56 % and 100 % for PAHs with molecular weight lower or equal to that of pyrene, but with slightly better performance for functionalised NPs [88]. The impact of the functionalisation became significant for PAHs with higher molecular weight than pyrene, since the removal increased from 0–3.5 % in naked NPs to 34.6–66.9 % in functionalised NPs [88].

Iron oxide NPs (mostly functionalised) have also been largely tested with real water samples, spiked with different concentrations of PAHs (Table 6), both with mixtures of multiple PAHs [86,90,131,134,136,137] or for PAHs alone, such as BaP and PYR [87,89,128,132]. Generally good recoveries were obtained in all cases, with the lowest being 73.7 % [131]. With exception of the study conducted by Hassan et al. [89], all the results done with spiked water samples used functionalised iron oxide NPs. Nevertheless, in the only study that tested bare iron oxide NPs with real water samples, very good removals of BaP and PYR were obtained (>97 %) [89].
<caption>**Table 6**</caption>

Published works on the extraction of PAHs from PAH-spiked real water samples by bare and functionalised iron oxide, presented from oldest to newest. The nomenclature is defined at the bottom of the tableª.

<table><thead><tr><th>Adsorbent<sup>a</sup></th><th>Ads. SSA (m²/g)</th><th>Ads. conc. (mg/L)<sup>b</sup></th><th>PAH<sup>a</sup></th><th>PAH conc. (µg/L)</th><th>Relevant experimental conditions</th><th>Recovery (%)<sup>c</sup></th><th>Process<sup>a</sup></th><th>Analytical method<sup>a</sup></th><th>Ref.</th></tr></thead><tbody><tr><td>Fe<sub>3</sub>O<sub>4</sub>/PDA</td><td>84</td><td>40</td><td>FLE, PYR, BaA, BbF, BaP, BghiP</td><td>0.02</td><td>Tap water samples<br>River water samples<br>Influent water samples<br>Effluent water samples</td><td>80.0–103<br>88.5–100<br>83.5–108<br>89.0–104</td><td>MSPE (Elution with ACN)</td><td>HPLC-FLD</td><td>[86]</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>/Au</td><td>-</td><td>100</td><td>BaP</td><td>0.001–0.10</td><td>Tap water samples<br>Lake water samples<br>River water samples</td><td>90.0–96.4<br>83.0–89.5<br>77.1–87.9</td><td>MSPE (Elution with ACN:CH<sub>2</sub>Cl<sub>2</sub>)</td><td>HPLC-FLD</td><td>[128]</td></tr><tr><td rowspan="2">Fe<sub>3</sub>O<sub>4</sub>-TBCD</td><td rowspan="2">-</td><td rowspan="2">400</td><td rowspan="2">FLU, ANT, FLE, PYR, BaA, BbF, BkF</td><td>1.5–600</td><td>Tap water samples<br>Snow water samples<br>River water samples</td><td>85.9–104.7<br>81.1–115.5<br>82.1–111.9</td><td rowspan="2">MSPE (Elution with ACN)</td><td rowspan="2">HPLC-FLD</td><td rowspan="2">[136]</td></tr><tr><td>1.5–600</td><td>Tap water samples<br>Lake water samples<br>Wastewater samples</td><td>89.3–94.6<br>89.7–92.2<br>90.7–94.8</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>@PDA/ZIF-7</td><td>-</td><td>900</td><td>FLU, ANT, FLE, PYR, BbF, BbP</td><td>1</td><td>Rain water samples</td><td>92.8–99.4</td><td>MSPE (Elution with EtOAc)</td><td>GC-MS</td><td>[137]</td></tr><tr><td rowspan="2">Iron oxide NPs</td><td rowspan="2">-</td><td rowspan="2">90–130</td><td>BaP</td><td>1</td><td>Drinking water samples</td><td>97–99</td><td>Batch</td><td>HPLC-PDAD</td><td>[89]</td></tr><tr><td>PYR</td><td>100</td><td>Ground water samples</td><td>97.9–98.5</td><td>MSPE (Elution with methanol: ACN (1:1))</td><td>GC-MS</td><td>[131]</td></tr><tr><td rowspan="2">Fe<sub>3</sub>O<sub>4</sub>@SiO<sub>2</sub>@TCPP</td><td rowspan="2">-</td><td rowspan="2">500</td><td>ACY, ACE, FLU, PHE, ANT,<br>FLE, PYR, BaA, CHY, BbF,<br>BaP, BkF, IcdP, DahA,<br>BghiP</td><td>0.05<br>0.5<br>5<br>0.05<br>0.5<br>5</td><td>River water samples</td><td>73.7–100.6<br>71.1–104.5<br>84.6–106.0<br>77.4–102.4<br>73.7–106.7<br>96.3–107.1</td><td rowspan="2">MSPE (Elution with CAN: THF (8:2))<sup>d</sup></td><td rowspan="2">HPLC-FLD</td><td rowspan="2">[132]</td></tr><tr><td>BaP</td><td>0.1–0.5</td><td>River water samples<br>Lake water samples<br>Wastewater samples</td><td>93.6–96.4<br>90.7–93.8<br>88.3–90.3</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>@SiO<sub>2</sub>@MMTA-Au</td><td>-</td><td>600</td><td>NAP, FLU, ACE, PHE, ANT,<br>FLU, PYR, BaA, BbF, BkF,<br>BaP, DahA, IcdP, BghiP</td><td>1–750</td><td>Real water samples</td><td>80.00–120.00</td><td>MSPE (Elution with ACN)</td><td>HPLC-FLD</td><td>[90]</td></tr><tr><td>MNPs@G3.0@4-MBA</td><td>-</td><td>1333</td><td>PHE, ANT, FLE, PYR, BaP</td><td>1–10</td><td>Real water samples</td><td>92–99</td><td>MSPE (Elution with ACN)</td><td>HPLC-UV</td><td>[138]</td></tr></tbody></table>

So, to sum up, numerous works have been found in the literature on the PAH adsorption using iron oxide NPs and most of them focus on their functionalisation for this purpose. Nevertheless, it seems that naked bare iron oxide (non-functionalised) has some potential to remove PAHs from water [88,89,92,93,135] and this remains quite unexplored so it is also needed to better understand how iron oxide interacts with PAHs.

On the other hand, contrary to what was the case with graphene and its derivatives, more publications focused on the adsorption of BaP [87, 89,128,132] and other HMW-PAHs [91]. In terms of functionalisation, it is interesting to point out that (poly)dopamine, or even tannic acid,

which is a comparable molecule, have been used multiple times [86,88, 91,129,137]. The use of these molecules to functionalise iron oxide NPs for PAH adsorption is mainly because the aromatic rings present in their structure should promote π-π interactions between adsorbent and adsorbate.

Concerning the adsorption models, different isotherm models fit well to the data depending on the work, mainly Freundlich, Langmuir, linear and Sips isotherms [72,87,89,92,93,128–130,132,133,135] and less frequently, also the Temkin, Toth and DR isotherms [86,92,135]. The kinetics were less explored, with pseudo-second order model generally

<footnote>a Nomenclature: PDA – polydopamine; TCBD – 1,4,7,10-tetrabenzyl-1,4,7,10-tetraazacyclododecane; C₁₀ – decanoic acid; ZIF – zeolitic imidazolate frameworks; NPs – nanoparticles; TCPP – Meso-Tetra (4-carboxyphenyl) porphyrin; CDs – carbon dots; MMTA – 2-mercapto-4-methyl-5-thiazoleacetic acid; 4-MBA – 4-mercapto-benzoic acid; PAH – polycyclic aromatic hydrocarbon; BaP – benzo(a)pyrene; FLU – Fluorene; ANT – Anthracene; FLE – fluoranthene; PYR – Pyrene; BaA – Benz(a)] anthracene; BbF – Benzo(b)fluoranthene; BkF – Benzo(k)fluoranthene; NAP – naphthalene; BbP – benzo(b)pyrene; ACE – acenaphthene; ACY – Acenaphthylene; PHE – Phenanthrene; CHY – Chrysene; IcdP – Indeno(1,2,3-cd)pyrene; DahA – dibenz(ah) anthracene; BghiP – benzo(ghi)perylene; ACN – acetonitrile; THF – tetrahydrofuran; EtOAc – Ethyl Acetate; MSPE – magnetic solid phase extraction; HPLC – high performance liquid chromatography; FLD – fluorescence detector; PDAD – photodiode array detector; GC – gas chromatography; MS – mass spectrometry.</footnote>

<footnote>b in extraction experiments, if the adsorbent concentration was not directly given, it was calculated by dividing the given mass by the solution volume.</footnote>

<footnote>c recovery determined based on the recovered amount of PAH after desorption.</footnote>

<footnote>d elution solvents presented correspond to the optimized conditions.</footnote>
<figure><img src="image_9.png" /><figcaption>**Fig. 7.** Compilation of relevant values of (a) removal efficiency and (b) adsorption capacity as a function of adsorbent concentration for the adsorption of PAHs in water with iron oxide nanoparticles [72,87–89,92,93,128,129,132,133,135]. Non-exhaustive compilation focused on the most studied PAHs (more details in Tables 5 and 6). Color-scheme indicates different PAHs and shape indicates different literature works as reported in the legend. Note: Break in x-axis for the intervals in concentration of 250–700 mg/L in (a), and 250–900 and 1100–1900 mg/L in (b).</figcaption></figure>

fitting the best to the data [89,92,128,129,132]. In terms of adsorption thermodynamics, adsorption with both functionalised and naked iron oxide NPs has been found exothermic [89,129,133].

Finally, the regeneration of these adsorbents was also evaluated in multiple works, up to 12 cycles of reuse [87,89,90,92,93,128,129,131, 132,134,135], with the magnetic separation making them undoubtedly easier to recover. Similarly to graphene adsorbents, solvent desorption was generally used to regenerate the adsorbents, with different organic solvents or mixtures being used, such as acetonitrile [90,134], methanol [89,129], acetonitrile: dichloromethane [128], acetonitrile: tetrahydrofuran [87,132], and methylbenzene: acetonitrile [131]. Only in the works of Wei et al. [92,93,135], thermal desorption (at 250–300°C) was employed to regenerate the adsorbents.

### 6.3. Performance of magnetic iron oxide/graphene nanocomposites for PAH removal

The potential of magnetic graphene-based composites for adsorption of PAHs has been explored in several works in the literature, both in terms of batch experiments or magnetic SPE (MSPE) applications [59, 60, 71–78,138–142]. These works are summarised in Table 7 and Table 8 and will be further discussed. As before, some of the most relevant results on specific PAHs reported are also presented in Fig. 8.

Similarly to what was verified for graphene-based adsorbents and iron oxide nanoparticles, there was a wide range of both adsorbent concentration (5 mg/L [74,77] up to 10 g/L [75] – see x-axis of Fig. 10) and PAHs concentration (10 ng/L [138] up to 30 mg/L [75]).

Yang et al. prepared graphene/iron oxide composites by coprecipitation of ferrous and ferric salts at the surface of GO [72] and used them for the batch adsorption of NAP (see Fig. 8(b) – yellow hexagon) and its derivatives 1-naphthol and 1-naphthylamine [72]. They observed that the adsorption capacity of the three compounds was much higher with the graphene/iron oxide composite than when using iron oxide alone, due to the presence of GO [72]. Furthermore, the adsorption capacity followed the order NAP < 1-NAP-OH < 1-NAP-A (337, 389 and 408 mg/g, respectively), increasing with the increase of polarity of these aromatic compounds, [72], which is not surprising considering also the polar character of GO. Furthermore for NAP, the adsorption was improved with the increase of temperature [72].

Two works studied the adsorption of phenanthrene with different types of magnetic graphene adsorbents. In the work of Huang et al., graphene oxide (in MGO), chemically-reduced graphene (in MCRG) and annealing-reduced graphene (in MARG) were tested as graphene sources [59] and the magnetic composites were prepared by post-synthesis assembly of the graphene and Fe₃O₄ nanoparticles into a composite [59]

(see Fig. 8(b) – green triangle). MCRG had the highest specific surface area (145.31 m²/g) and showed the highest adsorption capacity as it could be expected (30.12 mg/g), even though it was used in much lower concentration (4.2 mg/L) than MGO and MARG [59]. MARG also had a higher adsorption capacity than MGO (26.18 vs 13.65 mg/g), even though it had lower specific surface area (78.72 vs 95.20 m²/g) and was present in lower concentration (12.5 vs 20 mg/L) [59]. Therefore, this confirms previous discussed works that reduced graphene has better adsorption performance than GO, possibly due to less defects on the sp² carbon rings and higher hydrophobicity that lead to more interactions with polyaromatic organic species. In the work of Zhang et al., results on the use of magnetic graphene nanosheets for the adsorption of PHE were also reported [74] (see Fig. 8 – green diamond symbol). The MGNS adsorbents were prepared by the in-situ precipitation of iron oxide in an aqueous medium containing graphene [74] and they were able to increase the PHE removal efficiency from around 25 % to around 96 % by increasing the adsorbent concentration from 5 to 100 mg/L [74]. They obtained a qmax value of 31.53 mg/g, which was similar to that of 30.12 mg/g obtained by Huang et al. for MCRG [59].

Jamshidi et al. reported on a graphene/iron oxide composite functionalised with cyanuric acid and allykamine (GO/Fe₃O₄/CC/AA) for the removal of PYR from water [75] (see Fig. 8 – light blue right pointing triangle). They observed an increase of removal efficiency by increasing the adsorbent concentration from 5 to 10 g/L [75].

Queiroz et al. prepared magnetic chitosan/GO composites (mCS/GO) by both a chemical and a green route and tested them for the adsorption of NAP, ANT and FLE [141] (see Fig. 8 – red and yellow square). They obtained better removal efficiency for NAP (~94 %), and for ANT and FLE it depended on the adsorbent material [141]. The authors explained that the higher stability of NAP possibly led to increased hydrophobic and π-π interactions with the adsorbents [141]. For ANT, the best RE was obtained with the green synthesised composite (~89 % vs ~48 %), while for FLE it was with the composite prepared by the typical chemical route (~76 % vs ~62 %) [141].

Finally, Vaz-Ramos et al. developed composites of few-layers graphene@tannic acid@ iron oxide nanostructures (FLG@TA@RSN) by a modified polyol method where tannic acid was used as a surfactant to ensure the separation of graphene layers and the colloidal stability in water [77]. These composites were used for adsorption of BaP (see Fig. 8 – dark blue star symbol), reaching RE values as high as 99.95 % and an adsorption capacity of 1.27 mg/g in water [77]. In order to increase the concentration of BaP, Vaz-Ramos et al. performed some experiments in 50 % water / 50 % ethanol and reached a qmax of 68.13 mg/g [77]. The high RE and qmax values obtained in this work for BaP with such a low adsorbent concentration (5 mg/L) show the efficiency of this composite
<caption>**Table 7**</caption>

Published works on the adsorption of PAHs from artificial water (under controlled laboratory conditions) by magnetic iron oxide/graphene composite nanomaterials, presented from oldest to newest. The nomenclature is defined at the bottom of the tableª.

<table><thead><tr><th>Adsorbent<sup>a</sup></th><th>Ads. SSA (m²/g)</th><th>Ads. conc. (mg/L)<sup>b</sup></th><th>PAH<sup>a</sup></th><th>PAH conc. (µg/L)</th><th>Adsorption capacity (q<sub>max</sub>, mg/g)/ Removal efficiency (RE, %)/ Recovery (%)<sup>c</sup></th><th>Recyclability</th><th>Process (Ads. Models<sup>b</sup>)</th><th>Analytical method<sup>a</sup></th><th>Ref.</th></tr></thead><tbody><tr><td>Fe<sup>b</sup>/iron oxide-oxyhydroxide/graphene</td><td>138</td><td>300</td><td>16 EPA PAHs<br>Mix PCBs<br>Phthalic acids</td><td>1<br>2<br>4</td><td>Max. Recovery: ~95<br>Max. Recovery: ~93<br>Max. Recovery: ~89</td><td>-</td><td>MSPE (Elution with CH<sub>2</sub>Cl<sub>2</sub>: hexane (9:1))<sup>d</sup></td><td>GC-MS</td><td>[71]</td></tr><tr><td>GO/ FeO•Fe<sub>2</sub>O<sub>3</sub></td><td>272.59</td><td>100</td><td>NAP<br>1NAP-A<br>NAP-OH</td><td>20 × 10<sup>3</sup> (for pH tests); other conc. n. m.</td><td>q<sub>max</sub>: 337 (283 K)-733 (323 K)<br>q<sub>max</sub>: 408 (283 K)<br>q<sub>max</sub>: 389 (283 K)</td><td>20 times</td><td>Batch (F; Endo)</td><td>UV</td><td>[72]</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>@SiO<sub>2</sub>@GO-PEA</td><td></td><td>600</td><td>ACY, FLU, PHE, ANT, PYR, BaA, CHY, BbF, BkF, BaP</td><td>50</td><td>-</td><td></td><td>MSPE (Elution with ACN: C<sub>2</sub>Cl<sub>4</sub> (4:1))</td><td>GC-FID</td><td>[139]</td></tr><tr><td>g-C<sub>3</sub>N<sub>4</sub>/Fe<sub>3</sub>O<sub>4</sub></td><td></td><td>2000</td><td>1-PHE-OH<br>3-PHE-OH<br>1-PYR-OH</td><td>0.5-2.5</td><td>Recovery: 90.1-98.4<br>Recovery: 92.0-97.4<br>Recovery: 99.2-102</td><td>-</td><td>MSPE (Elution with acetone)</td><td>HPLC-FLD</td><td>[140]</td></tr><tr><td>MGO</td><td>95.20</td><td>20</td><td>PHE</td><td>10-1200</td><td>q<sub>max</sub>, exp: 13.65<br>q<sub>max</sub>, exp: 26.18<br>q<sub>max</sub>, exp: 30.12</td><td>-</td><td>Batch (DA)</td><td>HPLC-FLD</td><td>[59]</td></tr><tr><td>MARG</td><td>78.72</td><td>12.5</td><td></td><td></td><td>Recovery: 98.1-100.5</td><td></td><td></td><td></td><td></td></tr><tr><td>MCRG</td><td>145.31</td><td>4.2<br>n.m</td><td></td><td></td><td>RE: ~25<br>RE: ~96</td><td>-</td><td>Batch (F/DA; P2O; Exo)</td><td>Spectrofluorometer</td><td>[74]</td></tr><tr><td>MGNS</td><td>61.98</td><td>5<br>100<br>20<br>n.m.</td><td>PHE</td><td>1000<br>n.m.</td><td>q<sub>max</sub>:31.53<br>Recovery: 95.48 (cycle 1)-81.06 (cycle 6)</td><td>6 times</td><td></td><td></td><td></td></tr><tr><td>GO/Fe<sub>3</sub>O<sub>4</sub>/CC/AA</td><td></td><td>5 × 10<sup>3</sup><br>10 × 10<sup>3</sup><br>10 × 10<sup>3</sup></td><td>PYR</td><td>20 × 10<sup>3</sup><br>(5-30)× 10<sup>3</sup></td><td>RE: 75<br>RE: 99.8<br>q<sub>max</sub>:3.777</td><td>8 times</td><td>Batch (F; P2O)</td><td>UV</td><td>[75]</td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>@GO-PANI</td><td>35.78</td><td>125</td><td>PAHs (FLU, PHE, PYR); N-PAHs (2-NO2-FLU, 9-NO2-ANT, 1-NO2-PYR, 3-NO2-FLE)</td><td>n.m.<br>7.5</td><td>Recovery: 56.2-98.4</td><td>15 times</td><td>MSPE (Elution with ACN)</td><td>GC-MS</td><td>[76]</td></tr><tr><td>G-mCS/GO</td><td></td><td>100</td><td>NAP<br>ANT<br>FLE</td><td>1280<br>1780<br>2020</td><td>q<sub>max</sub>: 11.9 / RE: 93.55<br>q<sub>max</sub>: 15.9 / RE: 89.25<br>q<sub>max</sub>: 12.5 / RE: 62.28</td><td>-</td><td>Batch</td><td>HPLC-UV<br>UV<br>UV</td><td>[141]</td></tr><tr><td>C-mCS/GO</td><td></td><td>100</td><td>NAP<br>ANT<br>FLE</td><td>1280<br>1780<br>2020</td><td>q<sub>max</sub>: 12.0 / RE: 93.89<br>q<sub>max</sub>: 8.6 / RE: 47.80<br>q<sub>max</sub>: 15.4 / RE: 76.07</td><td>10</td><td>Batch</td><td>HPLC-UV<br>UV<br>UV</td><td>[141]</td></tr><tr><td>MNPs@PAMAM-G2.0@GO</td><td></td><td>250</td><td>CBZ, 7-MQL, 9-MCB, DBT, ANT, FLE, PYR, 4,6-DBT</td><td>10</td><td>Recovery: ~75-95</td><td></td><td>MSPE (Elution with ACN)</td><td>HPLC-VWD</td><td>[138]</td></tr><tr><td>FLG@TA@RSN</td><td>34</td><td>5</td><td>BaP</td><td>1-10<br>10-5000</td><td>q: 1.27 / RE: 99.95<sup>d</sup><br>q<sub>max</sub>: 68.13 / RE: 31.70<sup>e</sup></td><td>3 times</td><td>Batch</td><td>UHPLC-FLD</td><td>[77]</td></tr></tbody></table>

<footnote>a Nomenclature: GO – graphene oxide; m-G/CNF – magnetic graphene layers on carbon nanofibers; PEA – 2-phenylethylamine; g-C<sub>3</sub>N<sub>4</sub> – graphitic carbon nitride; MGO – magnetic graphene oxide; MARG – magnetic annealing-reduced graphene; MCRG – magnetic chemically-reduced graphene; MGNS – magnetic graphene nanosheets; CC/AA – cyanuric chloride and allylamine functionalised; PANI – polyaniline; G-mCS – Green-synthesised magnetic chitosan; C-mCS – Chemically-synthesised magnetic chitosan; MNPs – magnetic nanoparticles; PAMAM – polyamidoamine; FLG@TA@RSN – few-layers graphene@tannic acid@ iron oxide raspberry-shaped nanostructures; PAH – polycyclic aromatic hydrocarbon; EPA – Environmental Protection Agency of the USA; PCB – Polychlorinated biphenyl; NAP – naphthalene; 1NAP-A – 1-naphthylamine; NAP-OH – 1-naphthol; FLU – Fluorene; PHE – Phenanthrene; 2MeNAP – 2-methylnaphthalene; 2BrNAP – 2-bromonaphthalene; ACY – Acenaphthylene; ANT – Anthracene; PYR – Pyrene; BaA – Benz(a)anthracene; CHY – Chrysene; BbF – Benzo(b)fluoranthene; BkF – Benzo(k)fluoranthene; BaP – benzo(a)pyrene; 1-PHE-OH – 1-phenanthrol; 3-PHE-OH – 3-phenanthrol; 1-PYR-OH – 1-hydroxypyrene; N-PAH – Nitro polycyclic aromatic hydrocarbon; 2-NO2-FLU – 2-nitro-fluorene; 9-NO2-ANT – 9-nitro-anthracene; 1-NO2-PYR – 1-nitro-pyrene; 3-NO2-FLE – 2-nitro-fluoranthene; FLE – fluoranthene; CBZ – carbazole; 7-MQL – 7-methylquinoline; 9-MCB – 9-methylcarbazole; DBT – dibenzothiophene; 4,6-DBT – 4,6-dimethyldibenzothiophene; ACN – acetonitrile; DMSO – Dimethyl sulfoxide; F – Freundlich isotherm; DA – Dubinin-Astakhov isotherm; P2O – Pseudo-second order model; endo – endothermic; exo – exothermic; MSPE – magnetic solid phase extraction; GC – gas chromatography; MS – mass spectrometry; FID – Flame ionisation detector; HPLC – high performance liquid chromatography; FLD – fluorescence detector; VWD – Variable Wavelength Detectors. bIn extraction experiments, if the adsorbent concentration was not directly given, it was calculated by dividing the given mass by the solution volume. c recovery determined based on the recovered amount of PAH after desorption.</footnote>

<footnote>b experiments performed in water with 1 % of ethanol.</footnote>

<footnote>c experiments performed in water with 50 % of ethanol to stabilise highest concentrations of BaP.</footnote>

<footnote>d elution solvents presented correspond to the optimized conditions.</footnote>
<caption>**Table 8**</caption>

Published works on the extraction of PAHs from PAH-spiked real water samples by magnetic iron oxide/graphene composite nanomaterials, presented from oldest to newest. The nomenclature is defined at the bottom of the table<sup>a</sup>.

<table><thead><tr><th>Adsorbent<sup>a</sup></th><th>Ads. SSA (m²/g)</th><th>Ads. conc. (mg/L)<sup>b</sup></th><th>PAH<sup>a</sup></th><th>PAH conc. (µg/L)</th><th>Relevant experimental conditions</th><th>Recovery (%)<sup>c</sup></th><th>Process<sup>a</sup></th><th>Analytical method<sup>a</sup></th><th>Ref.</th></tr></thead><tbody><tr><td>Fe<sup>0</sup>/iron oxide-oxyhydroxide/graphene</td><td>138</td><td>300</td><td>16 EPA PAHs</td><td>0.5</td><td>Lake water samples</td><td>90–93</td><td>MSPE (Elution with hexane: CH<sub>2</sub>Cl<sub>2</sub> (1:9))</td><td>GC-MS</td><td>[71]</td></tr><tr><td>G/Fe<sub>3</sub>O<sub>4</sub>@PT</td><td>-</td><td>200</td><td>Mix PCBs</td><td>1</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td>Phthalic acids</td><td>2</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td>NAP, ACE, FLU, PHE, ANT</td><td>1</td><td>Seawater samples</td><td>83–107</td><td>MSPE (Elution with toluene)</td><td>GC-FID</td><td>[142]</td></tr><tr><td>m-G/CNF</td><td>-</td><td>1000</td><td>NAP, FLU, PHE, 2MeNAP, 2BrNAP</td><td>0.1</td><td>Tap water samples</td><td>97.4–99.3</td><td>MSPE (Elution with CH<sub>2</sub>Cl<sub>2</sub>)</td><td>GC-FID</td><td>[73]</td></tr><tr><td></td><td></td><td></td><td></td><td>50</td><td></td><td>98.8–99.9</td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td></td><td>0.1</td><td>River water samples</td><td>96.8–98.6</td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td></td><td>50</td><td></td><td>98.5–99.4</td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td></td><td>0.1</td><td>Well water samples</td><td>96.5–98.7</td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td></td><td>50</td><td></td><td>97.2–99.1</td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td></td><td>0.1</td><td>Wastewater samples</td><td>95.5–97.7</td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td></td><td>50</td><td></td><td>96.1–98.7</td><td></td><td></td><td></td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>@SiO<sub>2</sub>@GO-PEA</td><td>-</td><td>600</td><td>ACY, FLU, PHE, ANT, PYR, BaA, CHY, BbF, BkF, BaP</td><td>10</td><td>Hookah water pipe samples</td><td>82.10–96.70</td><td>MSPE (Elution with ACN: C<sub>2</sub>Cl<sub>4</sub> (4:1))</td><td>GC-FID</td><td>[139]</td></tr><tr><td></td><td></td><td></td><td></td><td>50</td><td></td><td>97.06–101.52</td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td></td><td>10</td><td>Wastewater samples</td><td>71.70–101.80</td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td></td><td>50</td><td></td><td>95.46–102.70</td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td></td><td>10</td><td>River water samples</td><td>73.00–97.30</td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td></td><td>50</td><td></td><td>93.88–99.46</td><td></td><td></td><td></td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>/RGO</td><td>100</td><td></td><td>NAP, ACY, ACE, FLU, PHE, ANT, FLE, PYR, BaA, CHY, BbF, BkF, BaP</td><td>0.05–2.5</td><td>Drinking water samples</td><td>82.2–112.4</td><td>MSPE (Elution with hexane: acetone (3:1))</td><td>GC-MS</td><td>[60]</td></tr><tr><td></td><td></td><td></td><td></td><td>0.05–2.5</td><td>River water samples</td><td>75.6–110.9</td><td></td><td></td><td></td></tr><tr><td>GO-Fe<sub>3</sub>O<sub>4</sub>@PS</td><td>750</td><td></td><td>NAP, 2MeNAP, FLU, ANT</td><td>0.05–50</td><td>Tap water samples</td><td>96.5–99.5</td><td>MSPE (Elution with hexane)</td><td>GC-FID</td><td>[78]</td></tr><tr><td></td><td></td><td></td><td></td><td>0.05–50</td><td>River water samples</td><td>95.8–98.1</td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td></td><td>0.05–50</td><td>Well water samples</td><td>96.5–98.7</td><td></td><td></td><td></td></tr><tr><td>GO/Fe<sub>3</sub>O<sub>4</sub>/CC/AA</td><td>-</td><td>n.m.</td><td>PYR</td><td>10–40</td><td>Tap water samples</td><td>99.0–99.92</td><td>Batch</td><td>UV</td><td>[75]</td></tr><tr><td></td><td></td><td></td><td></td><td></td><td>River water samples</td><td>98.0–99.50</td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td></td><td></td><td>Sea water samples</td><td>95.0–99.80</td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td></td><td></td><td>Sewage water samples</td><td>92.0–97.0</td><td></td><td></td><td></td></tr><tr><td>Fe<sub>3</sub>O<sub>4</sub>@GO-PANI</td><td>35.78</td><td>125</td><td>PAHs and N-PAHs</td><td>25</td><td>River water samples</td><td>92.3–112.9</td><td>MSPE (Elution with ACN)</td><td>GC-MS</td><td>[76]</td></tr><tr><td></td><td></td><td></td><td></td><td></td><td>Tap water samples</td><td>92.8–108.0</td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td></td><td></td><td>Mineral water samples</td><td>91.6–105.6</td><td></td><td></td><td></td></tr><tr><td>MNPs@PAMAM-G2.0@GO</td><td>250</td><td></td><td>CBZ, 7-MQL, 9-MCB, DBT, ANT, FLE, PYR, 4,6-DBT</td><td>0.01–1</td><td>Real water samples</td><td>92.5–105.2</td><td>MSPE (Elution with ACN)</td><td>HPLC-VWD</td><td>[138]</td></tr></tbody></table>

and evidence the potential of such materials for PAH adsorption.

On the other hand, a lot of other works tested these types of composites for magnetic SPE of PAH mixtures [71,73,76,134,139,140]. Nian et al. reported the use of graphitic carbon nitride/iron oxide for the extraction of hydroxylated PAHs [140]. They prepared the composite by solvothermal synthesis of Fe<sub>3</sub>O<sub>4</sub> in the presence of graphitic carbon nitride [140]. In terms of extraction performance, they obtained similar recoveries for 1-PHE-OH and 3-PHE-OH (recoveries between 90.1 % and 98.4 %) and a higher affinity for 1-PYR-OH (recovery between 99.2 % and 102 %) [140], indicating that the extraction depends on the hydrophobicity [140]. Karamani et al. also prepared iron/iron

oxide-oxyhydroxide/graphene composites for the extraction of the 16 US EPA PAHs as well as PCBs and phthalic acids and obtained a recovery of around 95 % of PAHs (and 93 and 89 % for PCBs and phthalic acids) [71]. This was higher than the results they obtained both for graphene and iron/iron oxide-oxyhydroxide [71]. Thus, it is possible there is a synergistic effect of the presence of both graphene and iron oxides. In the work of Manousi et al., iron oxide/graphene composite modified with polyaniline was synthesised for the extraction of both PAHs and nitrated PAHs and recoveries between 56.2 % and 98.4 % were obtained depending on the PAH [76]. Likewise, Tong et al. prepared composites of GO and iron oxide modified with polyamidoamine dendrimers for the

<footnote><sup>a</sup> Nomenclature: G – Graphene; PT – polythiophene; m-G/CNF – magnetic graphene layers on carbon nanofibers; GO – graphene oxide; PEA – 2-phenylethylamine; RGO – reduced graphene oxide; PS – polystyrene; CC/AA – cyanuric chloride and allylamine functionalised; PANI – polyaniline; MNPs – magnetic nanoparticles; PAMAM – polyamidoamine; PAH – polycyclic aromatic hydrocarbon; EPA – Environmental Protection Agency of the USA; PCB – Polychlorinated biphenyl; NAP – naphthalene; ACE – acenaphthene; FLU – Fluorene; PHE – Phenanthrene; ANT – Anthracene; 2MeNAP – 2-methylnaphthalene; 2BrNAP – 2-bromonaphthalene; ACY – Acenaphthylene; PYR – Pyrene; BaA – Benz(a)anthracene; CHY – Chrysene; BbF – Benzo(b)fluoranthene; BkF – Benzo(k)fluoranthene; BaP – benzo(a)pyrene; N-PAH – Nitro polycyclic aromatic hydrocarbon; FLE – fluoranthene; CBZ – carbazole; 7-MQL – 7-methylquinoline; 9-MCB – 9-methylcarbazole; DBT – dibenzothiophene; 4,6-DBT – 4,6-dimethyldibenzothiophene; ACN – acetonitrile; MSPE – magnetic solid phase extraction; GC – gas chromatography; MS – mass spectrometry; FID – Flame ionisation detector; HPLC – high performance liquid chromatography; VWD – Variable Wavelength Detectors, n.m. – not mentioned.</footnote>

<footnote><sup>b</sup> in extraction experiments, if the adsorbent concentration was not directly given, it was calculated by dividing the given mass by the solution volume.</footnote>

<footnote><sup>c</sup> recovery determined based on the recovered amount of PAH after desorption.</footnote>
<figure><img src="image_10.png" /><figcaption>**Fig. 8.** Compilation of relevant values of (a) removal efficiency and (b) adsorption capacity as a function of adsorbent concentration for the adsorption of PAHs in water with iron oxide/graphene composites. Non-exhaustive compilation focused on the most studied PAHs (more details in Tables 7 and 8). Color-scheme indicates different PAHs and shape indicates different literature works as reported in the legend. Note: Break in x-axis for the intervals in concentration of 150–4901 and 5099–9901 mg/L in (a), and 150–4901 mg/L in (b).</figcaption></figure>

extraction of PAHs and their derivatives [134] and obtained recoveries between 75 % and 95 %.

As pointed out with graphene and iron oxide, multiple publications explored the use of these composites for the analysis of real water samples (spiked with known concentrations of PAHs) as illustrated in Table 8. In all works, generally good recoveries were obtained, with 71.7 % being the lowest value obtained with Fe₃O₄@SiO₂@GO-PEA with wastewater samples [139]. Rezvani et al. reported the application of magnetic graphene layers synthesised on carbon fibers for the magnetic SPE of spiked tap, river water, well water and wastewater in two different PAH concentrations [143]. They obtained similar high recoveries for all types of water samples and in both concentrations [143]. Similar results were obtained by Amiri et al. [78] and Jamshidi et al. [75].

To briefly sum up, the use of magnetic iron oxide/graphene composites for PAH adsorption in water has been reported. However, if we compare the number of works in Table 7 with Table 3 and Table 5, we see that graphene/iron oxide composite remain less studied for the adsorption of PAHs than their graphene and iron oxide counterparts. The preparation of different types of magnetic graphene materials was studied, as well as their incorporation on supports such as carbon fibres for MSPE applications. In fact, more works were found evaluating the MSPE potential of these materials in comparison to batch adsorption experiments. In a similar way to what was observed for graphene materials, numerous works focused only on PAH mixtures or in single-component experiments with PAHs of lower molecular weight and there is a lack of the study of adsorption of more dangerous PAHs, such as BaP. However, a lot of works were found where the removal of PAH derivatives was studied. Nevertheless, when compared with graphene materials or iron oxide NPs alone, it seems the composites have better performance [71,72], suggesting a synergistic effect of the two nano-materials together.

Concerning the adsorption mechanisms, they were less studied. Generally, Freundlich isotherm was the one that fitted better to the data [72,74,75], possibly because of the heterogeneity of the surface given the presence of both graphene and iron oxide, but Dubinin-Astakhov isotherm also fitted well in certain works [59,74]. Only two works studied the adsorption kinetics, and it was the pseudo-second order model that fit better to the data [74,75]. Similarly, only two works reported on the adsorption thermodynamics and opposite results were obtained: adsorption of NAP on GO/FeO•Fe₂O₃ was endothermic [72], while adsorption of PHE on MGNS was exothermic [74], which could be suggest different types of interactions with PAHs for different graphene materials.

Finally, the regeneration of magnetic graphene/iron oxide-based composites was also evaluated in multiple works [74–77,134,139,143]

and up to 20 cycles of recycling were performed [139]. Here, again, the preferential regeneration method was solvent desorption with acetoni-trile [76,77,138], methanol [75], hexane [74] or a mixture of acetoni-trile: tetrachloroethylene [139], in agreement with the previous sections.

## 6.4. Mechanism of interaction between the adsorbents and PAHs

Multiple works have tried to explain the interactions that occur during the adsorption of PAHs onto magnetic iron oxide/graphene composites nanomaterials [59,72,141], as well as in graphene materials alone [17,79,81,121] and in iron oxide nanoparticles. Additionally, several works used density functional theory (DFT) calculations [144–151] to better understand these interactions and how adsorption works.

### 6.4.1. Interactions between graphene and PAHs

In the case of graphene and graphene-based materials, it is well known that the adsorption of PAHs occurs due to hydrophobic effects and π-π electron-donor-acceptor (EDA) interactions between the graphene sp² carbon network and the aromatic rings of PAHs [17,81]. In addition, in the case of PAHs with functional groups and graphene derivatives, additional n-π* EDA interactions can also be present, which correspond to the delocalisation of lone pair electrons on atoms of oxygen or nitrogen into the π* orbital of an aromatic ring [152]. This was reported as the responsible interaction for the adsorption of 1-naphthol (n-electron donor) on sulfonated graphene (π-electron acceptor) by Zhao et al. [79].

Wang et al. has also reported that the presence of wrinkles on the graphene layers can also have significant influence in the interaction of these materials with PAHs [17].

Theoretical modelling has been employed to understand how PAHs adsorb on the surface of graphene materials and what are the preferred molecular configurations. Zhao et al. reported that NAP adsorbs onto sulfonated graphene in a tilted form [79] with two different states: either with a distance of 3.57 Å or 3.71 Å from the graphene surface [79]. For NAP-OH, they report one state of adsorption, diagonally to the surface (because of the presence of the OH group) at a distance of 3.15 Å [79]. Similarly, Sun et al. described that NAP, ANT and PYR adsorbent into RGO surfaces in a tilted form [81] with one stable state each at distances of 3.389, 3.447 and 3.455 Å, respectively [81].

In the work of Li et al., the adsorption of the 16 US EPA priority PAHs to the surface of graphene was studied by DFT calculations [144]. In their work, they reported that PAHs adsorb parallel to the graphene's surface either in a "top" conformation (i.e., the centre of the aromatic ring is on top of a carbon atom of the graphene) or in a "bridge"
conformation (where the centre of the aromatic ring is in the middle of a C-C bond of graphene) [144] (see Fig. 9), with equilibrium distances between the PAH and graphene varying between 3.38 and 3.51 Å [144]. According to their work and previous works [144–148], the interaction between neutral PAHs and graphene is not only due to π-π interactions, but also a combination of Pauli repulsion, short-range Coulomb interactions and Van der Waals interactions. Furthermore, they reported on an increase of adsorption energies with the increase in total number of atoms (and thus the number of rings) of the PAH [144], indicating that HMW-PAHs are more strongly adsorbed onto graphene than LMW-PAHs. Additionally, they also observed a positive linear relationship between PAH hydrophobicity and the adsorption energy [144]. Interestingly, they also simulated the adsorption behaviour of ACE and IcdP when graphene sheets were under strain [144] and noted that when graphene is under compressive strain (leading to the formation of wrinkles [144]), PAH adsorption is stronger [144], confirming what had been experimentally reported by Wang et al. [17].

However, multiple works have reported contradictory experimental results in whether the interactions between PAHs and graphene materials increase or decrease with the increase on the number of PAH rings, which indicates that the adsorption interaction between PAHs and graphene materials is not completely understood. For example, Sun et al. reported that the adsorption of NAP, PYR and ANT on GO and RGO followed the order ANT < PYR < NAP [81], while Wang et al. reported that adsorption of NAP, PHE and PYR on GNS and GO followed the order NAP < PHE < PYR, depending on PAH concentration [17]. Yi et al. have developed crucial work on trying to understand these opposite results by employing molecular dynamics simulations [149]. In their work, they hypothesise that, given the low solubilities of PAHs in water, PAHs might have self-interactions (also corresponding to hydrophobic and π-π interactions) and form molecular clusters [149]. It is because of the formation of these PAH clusters that they can be found in water in concentrations higher than their solubility, which has been reported for aquatic environments [149]. With this hypothesis, they modelled the formation of PAH molecular clusters in different solvents, using NAP, PHE and PYR as model PAHs, and then studied the adsorption of these PAHs (both individual or in molecular clusters) on graphene by molecular dynamics [149]. They saw that, while in individual state the adsorption of PYR was the lowest, when considering clusters NAP showed the lowest adsorption [149]. Therefore, this suggested that it is important to also consider the self-stacking of PAHs in water when trying to understand the adsorption of these pollutants.

### 6.4.2. Interactions between iron oxide and PAHs

As previously seen, when iron oxide nanoparticles are used to adsorb PAHs, they are usually functionalised. In the case of functionalised iron oxide NPs, the functionalised moieties often contain aromatic rings on their structure, such as dopamine/polydopamine [86,88,129], tannic acid [91], or others, and thus π-π interactions can occur between these functional groups and the aromatic rings of PAHs. On the other hand, the interactions between iron oxide itself and PAHs have not been extensively studied and reported. For example, the works of Wei et al. present DFT results to better understand the adsorption of phenanthrene and

9-phenanthrol [93,135], but only in terms of the interactions between these PAHs and the molecules grafted to the surface of iron oxide NPs, with no results presented for the interactions between iron oxide and PAHs. Likewise, in the work of Deng et al. where Fe₃O₄/Au composites are used for BaP extraction [128], the authors only comment on the interactions between gold and BaP [128].

Nevertheless, it has been reported that PAHs and other non-polar hydrophobic compounds can interact with positively charged mineral surfaces by charge-induced dipole-dipole interactions [150,153] or by ion-π interactions between the cationic sites and the aromatic rings [153–155]. A work performed by Tunega et al. was found where a DFT study was performed to understand the sorption of PAHs onto the surface of goethite (α-FeOOH) [153]. In their work, they reported that PAHs adsorb onto goethite in an almost parallel conformation [153] with distances to the OH surface groups of 2.3–2.7 Å [153]. The adsorption of PAHs on goethite surface was reported to be due to polarisation of the aromatic rings by the polar OH groups [153] and to the formation of weak hydrogen bonds [153], with the formation of weak complexes on the surface of goethite [153]. Out of the studied PAHs (NAP, ANT, PHE, PYR), the highest adsorption was reported for ANT followed by NAP [153] and the authors justified this by the linearity of these two PAHs and the configuration of the OH surface groups of goethite that allows better interaction with linear PAHs [153].

Another work by Dzade et al. was found on the DFT study of the adsorption of, not PAHs, but benzene, on hematite (α-Fe₂O₃) [151]. They compared two benzene conformations (horizontal and vertical) on the hematite surface [151] and observed that the horizontal parallel conformation was the most stable [151]. The authors reported that the interactions correspond to van der Waals interactions at the (0001) surface of hematite and much stronger cation n-π interactions in the (0112) surface of hematite [151].

### 6.4.3. Interactions between magnetic iron oxide/graphene nanocomposites and PAHs

The possible interactions of PAHs with graphene and/or iron oxide materials have been represented in Fig. 10. Nevertheless, in most cases, when describing the adsorption mechanisms between PAHs and magnetic iron oxide/graphene nanocomposites, only the mechanisms between PAHs and graphene are presented, since this nanomaterial is the main responsible for the PAH adsorption. Therefore, the adsorption corresponds primarily to π-π interactions between the aromatic rings of PAHs and sp² carbons in graphene [59,72,141], and to the hydrophobic effect [59,141].

Other interactions that are possible correspond to n-π EDA interactions in the case of PAH derivatives and electron deficient sites in GO composites [72], as well as weaker interactions, such as weak hydrogen bonding and electrostatic interactions [72].

### 6.5. Recommendations for future works

Until here, this review highlighted clearly that both graphene-based and iron oxide nanomaterials and their composites have the potential to adsorb PAHs. Nevertheless, we see that there is a big disparity in the

<figure><img src="image_11.png" /><figcaption>Fig. 9. Example of “top” conformation and “bridge” conformation of PAH adsorbed on graphene (adapted from [144]).</figcaption></figure>
<figure><img src="image_12.png" /><figcaption>**Fig. 10.** Scheme of the possible interactions between PAHs and graphene/iron oxide materials.</figcaption></figure>

experimental conditions of all the works reported, which can make it more difficult to compare the performance of different adsorbents and to understand certain results. This disparity of experimental conditions is clear in Figs. 6, 7 and 8 for the adsorbent concentration, which varied largely between 4.2 mg/L and 20 g/L, but the range of PAH concentrations was also very wide. Future experimental works could benefit from more homogeneous adsorption conditions, but we recognize that completely uniformising the experimental conditions could be a very difficult challenge. Furthermore, the PAH concentration is limited also by the solubility of a specific PAH and the amount of adsorbent can be limited by the synthesis, for example. Therefore, we propose that future works present, at least, both the adsorption capacity (q) and the removal efficiency (RE). This was not the case for multiple works presented in Tables 3, 5 and 7, and it seems crucial in our understanding as these two quantities complement each other. To exemplify, let's hypothesise that one work reports one adsorbent A with a removal efficiency of 99 % and adsorption capacity of 1 mg/g for a certain PAH, while another work reports on adsorbent B with efficiency of 50 % but adsorption capacity of 100 mg/g. If we only consider RE, we might be misled to conclude that adsorbent A is better when in fact adsorbent B can adsorb a higher quantity per g of adsorbent. Similarly, if we only saw the values of q, we might think that adsorbent A is not good, when in certain experimental conditions, it can adsorb almost completely the PAH present in water. If both RE and q are reported, we can have a better understanding of the adsorbent performance and where it is positioned in comparison to others.

Another important matter to consider is which PAHs to be preferentially studied. As seen in Sections 6.1, 6.2 and 6.3, the PAHs in study vary a lot. Ideally, the adsorption of all 16 US EPA Priority PAHs should be performed in experimental works, however this is not always possible. When this is the case, experiments should focus on the PAHs already regulated by the different entities (EU, WHO, US - see Table 2) which include mostly bigger HMW-PAHs like BaP, BbF, BkF, BghiP and IcdP, but also some LMW and lighter HMW-PAHs, like NAP, ANT and FLE, in detriment of others that have been studied a lot such as PYR and PHE.

Finally, it is important to consider the applicability of the reported adsorbents, in particular graphene/ iron oxide composites, for real life applications. At laboratory scale, both batch adsorption experiments and SPE-type experiments have been reported, but the scaling-up of such methodologies for depollution or analytical practices should be considered.

In water depollution applications, processes based on the reported batch adsorption experiments might be preferential as much larger quantities of water will need to be treated. However, high quantities of

adsorbents will also be needed and longer contact times might be necessary for adsorption. Moreover, for such big quantities, very big magnets will be needed to recover the adsorbents and avoid their leaching, and this process might be too long, thus making magnetic separation less advantageous at such scale. So, in fact, batch adsorption might not be feasible for such big volumes and might have to be adapted and optimized into continuous adsorption procedures. These processes could be coupled with analysis of the treated water directly by HPLC methods to ensure the removal of PAHs was successful. In depollution applications, the final disposal of the adsorbed PAHs is also of the utmost importance. In order to promote the reuse of the adsorbents and avoid large quantities of waste, the PAHs will have to be desorbed and possibly degraded. In laboratory scale, this desorption has been tested, mostly with the use of organic solvents but at a much bigger scale, thermal processes might be preferred.

On the other hand, for analytical applications of PAH monitoring, it might be easier to adapt the already-developed composite adsorbents to such applications. For example, Wang et al. already successfully fixed graphene materials on a stainless steel wire for SPME [126]. In monitoring applications, the objective is only to determine the concentration PAHs in an environmental sample. If procedures based on SPME are used, as long as the samples are representative, small volumes of water are needed and thus, so is the amount of adsorbent. Thus, there is here less need to consider large scales and the scale-up process. Such procedures could then be directly integrated with thermodesorption processes and GC-MS techniques for an easy and reliable monitoring of PAHs. Nevertheless, the effect of such desorption processes on the adsorbents should be studied to ensure that they do not lose efficiency significantly.

## 7. Conclusions

PAHs are pollutants of high concern given their health risks and widespread presence in the environment. This review paper highlights that even though multiple common techniques used in wastewater treatment plants can remove PAHs to a certain extent, there is a constant need to develop more efficient tools for their removal.

The possibility to create composites of iron oxide and graphene nanomaterials comes as especially appealing for depollution practices because it couples the easy magnetic separation of iron oxides with the high adsorption performance of graphene materials. In fact, magnetic iron oxide/graphene composites nanomaterials have already caught the attention of the scientific community and appear as promising and efficient materials to adsorb PAHs from aquatic environments that can be applied as tools for pollutant removal, as well as possibly in tools for
their monitoring.

Nevertheless, as previously discussed, there is still more work that should be done before such type of graphene/iron oxide are ready to be used at WWTPs scale. Furthermore, as seen in this review, PAHs in aqueous environments can also be found in particulate phase (i.e. bound to suspended particles), in particular the less soluble HMW-PAHs, which are also generally the more dangerous. Thus, the removal and disposal of this particulate-bound phase should also be considered and studied in the future in order to fully avoid the health risks of this group of organic pollutants.

CRediT authorship contribution statement

**Vaz-Ramos Joana:** Writing – original draft, Methodology, Investi-
gation, Data curation, Conceptualization. **Le Calvé Stéphane:** Writing –
review & editing, Validation, Supervision, Project administration,
Conceptualization. **Bégin-Colin Sylvie Sylvie:** Writing – review &
editing, Validation, Supervision, Project administration, Methodology,
Funding acquisition, Conceptualization.

Declaration of Competing Interest

The authors declare that they have no known competing financial
interests or personal relationships that could have appeared to influence
the work reported in this paper.

Acknowledgments

This work has been conducted in the framework of the FIGHTVIRUS
project financially supported by the region GRAND-EST and ANR (ANR-
20-GES1-0013), of the CAPTALL project funded by the MICA Carnot
Institut and of the ON-LINE PAC ANALYSIS project (ANR-24-CE04-
3932-02). The authors thank the technical department of ICPEES and
IPCMS as well as all the administrative staff of both institutes for their
support.

Data availability

No data was used for the research described in the article.

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