Occurrence and Monitoring of POPs

#### **Chapter 1**

## Endocrine Disruptor Impact on Zebrafish Larvae: Posterior Lateral Line System as a New Target

*Ahmed Nasri, Patricia Aïssa, Ezzeddine Mahmoudi, Hamouda Beyrem and Véronique Perrier*

#### **Abstract**

Endocrine-disrupting chemicals (EDCs), including polychlorinated biphenyls (PCBs), bisphenol A (BPA), pharmaceutical drugs, and pesticides, affect a variety of hormone-regulated physiological pathways in humans and wildlife. The occurrence of these EDCs in the aquatic environment is linked with vertebrates' health alteration. EDCs exhibit lipophilic characteristics and bind to hydrophobic areas of steroid receptors, such as the estrogen receptor, which are involved in vertebrate developmental regulation. Mainly, EDCs modify the transcription of several genes involved in individual homeostasis. Zebrafish conserve many developmental pathways found in humans, which makes it an appreciated model system for EDCs research studies, especially on early organ development. In the current chapter, we emphasize on latest published papers of EDCs effects on lateral line regeneration in zebrafish larvae. Similarly, we describe other special impacts of EDCs exposure. In conclusion, we make the case that the zebrafish lateral line exposed to EDCs can provide important insights into human health.

**Keywords:** zebrafish larvae, posterior lateral line (PLL), endocrine-disrupting chemicals (EDCs), experimental exposure, regeneration

#### **1. Introduction**

During the last two decades, special attention from scientists and the public has been increasingly given to the harmful effects that may result from exposure of humans or wildlife to chemicals, having the property of interfering with the endocrine system. According to McLachlan John [1], one of the very first references to the problematic posed by what are called "endocrine disruptors" is that of Roy Hertz [2] who considered "that the fact should be taken into account that the use of hormones in animal feed risked exposing certain individuals to these substances, when they should never have been in contact with such molecules in their lifetime, that we were in the process of creating a steroid cycle in our environment, and that we had to seriously consider the implications this could have for our development, our growth and perhaps for our reproductive function."

The presence of endocrine disruptors (EDCs) in the environment has raised many questions within the scientific community because of the risks they represent for humans and ecosystems [3]. Indeed, EDCs can cause "adverse effects on the

health of an organism or its progeny in relation to changes in endocrine function" [4]. EDCs can act on all stages of endocrine regulation, from the hormones synthesis to their action in target tissues [3]. They include a large category of chemical substances such as natural and synthetic hormones or synthetic molecules such as pesticides, polychlorinated biphenyls (PCBs), or even alkylphenols from the industrial activities increasing as well as the pharmaceuticals used in everyday life.

Aquatic environments are often the ultimate outlet for many anthropogenic chemicals [5]. Several studies, carried out in the laboratory or in the natural environment, have revealed harmful effects of these substances on fauna and in particular on the physiology of fish reproduction [6]. Numerous studies have established a direct link between the presence of EDCs in the environment and alterations in reproductive function in certain aquatic organisms [6]. In fish, disturbances of hormone-regulated proteins, as well as histological alterations in the male gonads, have been demonstrated and linked to exposure to estrogen mimetics [7]. Indeed, certain EDCs are compounds that act as agonists of estrogen receptors (ERs) such as 17α-ethinylestradiol (EE2). These estrogen-mimetic compounds also include certain pesticides, alkylphenols, or bisphenols. These EDCs can thus mimic the action of endogenous natural estradiol and are able to induce the expression of estrogenregulated genes [8].

While most of the research has focused on the EDCs effects on peripheral organs and more specifically on the gonad [9], this chapter aims to focus on the latest published research on the impact of EDCs on the peripheral nervous system of zebrafish larvae, more precisely, the posterior lateral line PLLs regeneration after exposure to environmental pollutants.

#### **2. Posterior lateral line (PLL)**

#### **2.1 Definition and organization**

The mechanosensory lateral line system is found in more than 25,000 fishes species—the Chondrichthyes (sharks, skates, rays, and chimeras), the "Agnatha" (the jawless fishes, the hagfishes, and lampreys), and the Osteichthyes (the bony fishes). The lateral line systems of the Osteichthyes and Chondrichthyes are more evolved, with superficial neuromasts on the skin and canal neuromasts in a series of lateral line canals on the head and trunk. The lateral line system of the lampreys is composed exclusively of superficial neuromasts located in grooves on the skin of the head and trunk. The lateral line system is a mechanosensory organ that detects the water vibration up to about 300 Hz [10], and relative pressure to the body. The information collected will be used for the detection of prey, schooling behavior, and avoiding predators [11]. Thus, the lateral line is composed of mechanoreceptors organs called neuromasts, which are distributed over the fish body. Neuromasts comprise a core of sensory hair cells surrounded by nonsensory support cells (**Figure 1**). Each neuromast is innervated by a branch of the lateral line nerve (axonal nerve). Generally, the lateral line is subdivided in two—the anterior lateral line (ALL) which comprises the neuromasts present on the head, jaw, and opercle, and its sensory neurons form the ALL ganglion, rostral to the ear, and the posterior lateral line (PLL) which contains the neuromasts on the trunk and tail and its sensory neurons form the PLL ganglion, caudal to the ear. Both the ALL and the PLL comprise several branches.

Each lateral line neuromast contains a central sensory hair cell progenitor embedded within a rosette formed by apical attachment and constriction of surrounding epithelial support cells. Lateral line hair cells are surrounded by a *Endocrine Disruptor Impact on Zebrafish Larvae: Posterior Lateral Line System as a New Target DOI: http://dx.doi.org/10.5772/intechopen.101609*

**Figure 1.** *Zebrafish posterior lateral line organization (PLL).*

group of support cells. Many of these cells are located basally to the hair cells and project interdigitating processes between them, acting to isolate hair cells from one another. In addition to serving as the source for new hair cells in the event of damage, they also provide structural and trophic support. Lateral line hair cells also share structural, functional, and molecular similarities with the hair cells in the vertebrate inner ear [12]. They are innervated by both afferent and efferent nerve fibers on their basal surfaces, which emanate from the lateral line ganglion and branch off at each neuromast [13]. The lateral line system has become increasingly popular as a model for studying hair cell biology relating to human hearing and balance disorders (**Figure 1**).

#### **2.2 PLL development**

In zebrafish, the posterior lateral line system genesis (PLL) is started at the end of 24 hpf by the migration of cells cluster (100 cells) called the primordium PLL (PLLp) under the skin near the ear to the end of the tail [14]. During this step, the PLLp periodically deposits neuromasts (L1, L2 … L7) along the body and will finish its migration, by the establishment of 2–3 terminal neuromasts (TN) at the level of the tail. In addition, each neuromast is formed by the sensory hair cells in the center. The pLLP is prearranged along its migratory axis—the posterior third cells (head zone) are highly proliferative and mesenchymal type, while those of the anterior two-thirds (leakage zone) are placed in epithelial rosettes [15]. These rosette cells called protoneuromasts will give rise to hair cells or support cells after they are deposited (**Figure 2**).

The sensory hair cells are formed from the central progenitor hair cells which after division produce pairs of differentiated hair cells sensitive to water movement. Support cells surround the central sensory hair cells and provide structural support [16]. Also, these internal support cells are themselves surrounded by mantle cells or external support cells. These cells form internal support cells during their proliferation (**Figure 1**) [17].

#### *2.2.1 Genetic control*

The PLLp migration is genetically regulated by several genes including those of the chemokine Cxcl12a and its receptors Cxcr4b and Cxcr7b. Cxcl12a is expressed in cells along the horizontal myoseptum, while PLLp expresses both Cxcr4b and

**Figure 2.** *Zebrafish neuromast anatomy and posterior lateral line development.*

Cxcr7b. The PLLp migration is inhibited following disruption of expression of Cxcl12a and these receptors [18]. Dambly-Chaudière and Ghysen [19] have shown that from an affinity point of view, Cxcr7b has more affinity for Cxcl12a (expressed in the trailing zone) than for Cxcr4b (expressed in the head zone). Cxcr7b to serve as a molecular sink due to this difference in binding affinity, preventing Cxcr4b receptors expressed near the leak area from binding to the chemokine [20]. Thus, through the primordium an expression gradient of Cxcl12a is generated, the binding of which can involve the polymerization of actin in the direction of migration [21].

The PLLp migration is dependent on the canonical signaling interaction between Wnt and FGF. In the head zone, the Wnt signaling is important while in the tail zone it is the FGF signaling which is essential. Wnt signaling results in Cxcr4b expression and proliferation mediation in the head area, allowing the primordium to maintain its size throughout its migration [22]. These proliferating cells migrate from the head area and grow throughout the pPLL, thus depositing several protoneuromasts. The number of neuromasts generated is reduced and the speed of neuromast placement decreases following disruption of proliferation [23, 24]. Wnt signaling also controls the expression of Fgf3 and Fgf10a ligands [25]. FGF signaling drives the morphogenesis of epithelial rosettes which will give rise to neuromasts at the level of the leakage zone [26]. Studies have shown that inhibition of FGF signaling inhibits the formation of rosettes, and consequently the formation of neuromasts [27]. Recently, Yanicostas et al. [28] reported that primordial migration is inhibited following the inactivation of kal1a, which is a homologous zebrafish gene encoding the extracellular matrix protein Anosmin-1a and known to be an activator of FGF signaling [29]. Likewise, the expression of kal1a is similar to that of cxcr4b, important in the head region and less essential in the tail region, but independent of CXCR4b/SDF1a signaling.

#### *2.2.2 Estrogen receptors implication*

In zebrafish, the most studied nuclear receptors are estrogen receptors (ERs). Three genes encoding these receptors have been identified—one encoding an ERα ortholog in mammals and two orthologs encoding ERβ (called ERβ1 and ERβ2) [30]. These receptors are found scattered throughout several regions of the body such as the gonads, liver, and nervous system [30]. Several studies have shown that all estrogen receptor isoforms exhibit high expression levels specifically in lateral line neuromasts [31]. Research work on the importance of ERs in the establishment of PLL has shown that a disturbance in the development of neuromasts by the absence of hair cells, occurs following the temporary suppression of the

*Endocrine Disruptor Impact on Zebrafish Larvae: Posterior Lateral Line System as a New Target DOI: http://dx.doi.org/10.5772/intechopen.101609*

expression of ERβ2 by a morpholino, which could be linked to aberrant activation of the Notch signaling pathway in embryos [32]. The temporary suppression by a morpholino of the expression of ERβ2 led to the disturbing development of these neuromasts (absence of hair cells), which could be related to aberrant activation of the Notch signaling pathway in embryos treated with morpholino [32]. Likewise, developmental defects and early embryo mortality occur following the suppression of ERα expression by application of morpholino from the translation of maternal transcripts [33]. Recently, a mutation in the gene encoding ERβ2 made it possible to identify a new mutant zebrafish line [34]. Deformed sexual intercourse (dominance of the adult male population), testes of altered morphologies, an imbalance in hormone levels, and an altered immune system, are the results of this mutation [34].

#### **3. Axonal nerve PLL**

#### **3.1 Axonal nerve regeneration**

Several tissues in fish exhibit a remarkable capacity for regeneration after injury, including the retina, cardiac tissue, and neurons. The lateral line which is a sensory system located on the surface and used to detect the movement of water shows a robust regenerative capacity. In addition, O'Brien et al. [35] approved that all cell types in this system can be genetically, physically, or chemically modified. Neuromasts deposited on the body surface of zebrafish larvae are innervated by sensory axons (PLL nerve) [11]. The superficial development of the nerve allows localized lesion, thus, the dynamics of complete regeneration of axons has been studied in zebrafish larvae 24 hours after axotomy [36]. Several works have studied the involvement of different cell types in the dynamics of degeneration and regeneration of the lateral line nerve (PLL). The inhibition of Schwann cell expression after chemical depletion inhibits the binding of neuregulin to the Erb receptor, which causes the exhaustion of nerves in peripheral Schwann cells [37]. In addition, the use of a transgenic mutant fish "named leo1" (no development of neural crest derivatives according to Nguyen et al. [38] leads to incoherent axonal regeneration. Also, inhibition of macrophages (by morpholino: molecule used to modify gene expression) causes very slow regeneration of the PLL nerve (more than 6 hours). In doubly mutant individuals characterized by the absence of Schwann cells and macrophages, axotomy of the nerve is followed by the death of these individuals [36].

The PLL nerve undergoes Wallerian degeneration (WD) after axotomy (**Figure 3**) with a succession of three phases; a delay phase (phase 1), a fragmentation phase (phase 2) followed by a final clearance phase (phase 3). The two phases of fragmentation and clearance begin approximately 3–5 hours post-axotomy respectively. In zebrafish, Wallerian degeneration (WD) of posterior lateral line axons occurs much faster than that studied in mammals and Drosophila [36]. Wallerian degeneration (WD) occurs in an orderly and stereotypical fashion involving described genetic control in the central nervous system (CNS) and peripheral (PNS) after trauma, stroke, or infection [39]. After axon cleavage, acute axonal degeneration (AAD) can occur at both adjacent ends of the lesion [40]. After ADA, it has been reported that the detached fragment remains intact during the delay phase or "lag phase." Following this latency phase, the axons will rapidly degenerate followed by a cut in the endoplasmic reticulum, degradation of the neurofilament, and swelling of the mitochondria of the axonal fragments. During the final phase, these fragments are removed using phagocytic cells. At the level of the PLL system, Schwann cells and macrophages play an important role in the process of Wallerian degeneration (WD). Thus, Schwann cells decrease

**Figure 3.** *Wallerian degeneration (WD) steps.*

myelin lipid synthesis in the first 12 hours after axotomy [41] and inhibit myelin protein production for 48 hours [42].

#### **4. Sensory hair cells**

The hair cells of neuromasts are mechanosensory cells that are able to detect the water movements and transform the energy generated into electrical signals transmitted automatically to the brain. Usually, the hair cell is highly polarized, both apicobasal and in the plane of the epithelium. In its structure, it is characterized by a crescent-shaped stereociliary bundle and a large single kinocilium, on the apical side of the cell facing the otic lumen (**Figure 4**). Due to the morphology and function of hair cells in the lateral line system, these cells are very similar to those in the inner ear of mammals [12]. Numerous researchers have demonstrated that lateral line hair cells are sensitive to exposure to aminoglycosides [43, 44], *in vivo* imaging of fish lateral line hair cells zebra was first used by [45] who observed the death and regeneration of hair cells induced by neomycin. Harris et al. [17] then developed additional assays to quantify the death and regeneration of hair cells in the lateral line, establishing a basis for genetic and chemical studies aimed at detecting modulators of hair cell sensitivity to ototoxic exposures and to perform further testing. Other research works [46] have been studied the lateral line response of zebrafish following exposure to ototoxic compounds, such as aminoglycosides and cisplatin. The zebrafish lateral line system is, therefore, a rapid and efficient model for evaluating the effects of a large number of pharmaceuticals on mechanicalsensory hair cells [32].

#### **4.1 Sensory hair cell regeneration**

#### *4.1.1 Cell differentiation and proliferation*

Neuromast hair cells are functional in zebrafish 3 days after fertilization [19] and contain 8–20 hair cells 5 days after fertilization [17]. They are surrounded by *Endocrine Disruptor Impact on Zebrafish Larvae: Posterior Lateral Line System as a New Target DOI: http://dx.doi.org/10.5772/intechopen.101609*

#### **Figure 4.**

*Hair cell of neuromast showing its functional asymmetry and its afferent/efferent innervation.*

nonsensory support cells, with basal nuclei and apical projections that intersect between them [47]. Rubel et al. [48] report that hair cells regenerate after damage via trans-differentiation or proliferation of carrier cells. Studies performed using tritium and bromodeoxyuridine (BrdU) labeling techniques have shown that lateral line hair cells can undergo continuous proliferation [49]. After acoustic trauma, fish can regenerate hair cells within 1–2 weeks [50]. However, the regenerative potential of neuromast hair cells can be considered a dose-dependent response depending on the level of damage [51]. Several studies have shown that hair cells can regenerate from mitotic divisions and the proliferation of supporting cells. At the zebrafish lateral line, hair cells normally undergo programmed cell death during development but are restored from support cells to the periphery after the S phase is produced [45]. The proliferating supporting cells can either remain on the periphery or migrate inward and their number increases after druginduced hair cell death. Further studies on the zebrafish lateral line have shown that the newly formed hair cells are the result of the proliferation of supporting (carrier) cells and that there are two sets of these cells within the neuromasts [52]; one group of cells is centrally located and considered the progenitor of hair cells, the other is peripheral whose function is unknown. This suggests that there may be functional specializations between populations of neuromast support cells. Hair cells can also regenerate from the trans-differentiation of carrier cells. By applying high levels of damage to neuromasts, hair cell replacement allows surrounding support cells to divide [53].

#### *4.1.2 Regeneration mechanisms*

The determination of genes and molecular mechanisms controlling the regeneration of hair cells via differentiation and proliferation of support cells has received great interest. Recently, DNA microarrays and next-generation sequencing (high throughput sequencing) have been used to identify which genes are activated after the destruction of hair cells. Thus, following the exposure of zebrafish to noise, a DNA chip was produced in order to follow the change in gene transcripts [54]. A modification of the genes encoding growth hormone and genes for myosin (light and heavy chains) and the major histocompatibility complex have been observed. Liang et al. [55] have shown that the "stat3/socs3" pathway can modulate the production of lateral line hair cells during development and of the adult inner ear during regeneration. In mouse models, Stat3 effectors may be involved in hair cell survival [56], but no role of Stat3 in hair cell regeneration has yet been reported in mammals.

Chemical screening techniques (chemical screening) have also been used to identify compounds that increase or inhibit hair cell regeneration in the lateral line of zebrafish. Chemical screening has been used to identify synthetic glucocorticoid activators that promote hair cell regeneration by increasing mitotic activity [57]. This study also identified inhibitors that reduced hair cell regeneration or prevented cell proliferation. Research on zebrafish has shown that the "Wnt" signaling pathway is involved in the regeneration of hair cells. At the neuromast level, inhibition of "Wnt/β-catenin" signaling reduces proliferation and differentiation of hair cells while activation of "Wnt" increases the number of hair cells and promotes reintegration of support cells in the cycle of hair cells and their proliferation [58]. In addition, activation of Wnt/β-catenin causes increased regeneration of hair cells [59]. The size of the neuromast is also regulated by a negative feedback loop that integrates the "Wnt" signaling activity [60] and promotes the proliferation of surrounding cells.

Jiang et al. [61] showed that analysis of RNA transcripts expressed in the zebrafish lateral line following neomycin-induced damage showed that Wnt/β-catenin signaling is weakly regulated at onset, but becomes highly regulated later, suggesting that "Wnt" is necessary for hair cell proliferation, but not immediately after hair cell damage. The research focused on "Wnt" signaling in zebrafish indicates that the "Sox2" transcription factor is involved in the proliferation and trans-differentiation of hair cells [59]. Thus, it has been shown that the newly formed hair cells originate from the proliferation of "Sox2 positive" cells. The "Sox2" factor is highly expressed in most progenitor cells of proliferating neuromasts [53] and is also required for trans-differentiation of carrier cells [62]. Another regulator of "Wnt/β-catenin" signaling is "ErbB/Neuregulin," which may act to regulate zebrafish lateral line interneuromast cells proliferation and neuromast development [63]. Other work suggests that the "Notch" pathway modulates regeneration because inhibition of "Notch" signaling can cause regeneration of hair cells [64]. Jiang et al. [61] studying lateral line hair cell regeneration in zebrafish found that "Notch" signaling is inhibited immediately after hair cell damage. In addition, constitutive expression of "Notch" may prevent the proliferation of hair cells from carrier progenitor cells, while the elimination of "Notch" activity produces an increased number of cellular progenitors and hair cells [16]. Wada et al. [60] reported that the proliferation pathways of hair cells, Wnt, Notch, and Erb are key components common in zebrafish and mouse models involved in the regulation of hair cell regeneration. In addition, the cell cycle inhibitor p27kip is known to regulate the regeneration of mammalian hair cells [65].

#### **4.2 Regeneration after contaminant exposure**

The lateral line system has been used to study the various compounds' toxicity, including aminoglycosides [44], trace metals [51], cisplatin [66], and endocrine disruptors [67–69] during the zebrafish early life stages. The hair cells death by necrosis and apoptosis, the new hair cells differentiation and proliferation following exposure to toxic substances have been described by several studies [44, 51, 66–69].

#### *4.2.1 Metal*

Hernandez et al. [51] have shown that copper is toxic to the zebrafish lateral line following exposure to concentrations varying between 1 and 50 M resulting in the death of hair cells of neuromasts. During the first 5 minutes of copper sulfate "CuSO4" exposure (5 M), they appear signs of damage [70]. Thus, morphological changes resulting from the onset of apoptosis and necrosis have been recorded [51]. Exposure of zebrafish larvae to "CuSO4" for 2 hours, followed by assessment of hair cell regeneration over the next 5 days, resulted in robust regeneration in neuromasts of the anterior lateral line, whereas posterior lateral line neuromasts showed little regeneration, suggesting a differential regeneration in the lateral line within the same animal [51]. In other work, they have shown that copper attenuates hair cell regeneration in part by reducing cell proliferation [71]. Also, neuromasts did not regenerate upon continuous exposure to copper [71]. These results suggest that copper is toxic to both hair cells and support cells, and its presence in waterways can negatively affect fitness. Other work has shown that the exposure of zebrafish larvae to cadmium causes an alteration of the lateral line, including a regeneration deficit associated with a change in behavior such as rheotaxis or escape reactions [72]. Exposure to 5 mg/L cadmium for 2 days severely damaged the nervous system of sea bass (*Dicentrarchus labrax*) [73].

#### *4.2.2 Aminoglycosides*

It was during the twenty-first century that the zebrafish lateral line was considered a model system for determining the toxicity of therapeutic drugs via hair cells. Williams and Holder [45] agreed that neomycin is toxic to lateral line hair cells. In addition, Harris et al. [17] showed that the response to neomycin is dose-dependent, identical in each neuromast, and that the sensitivity of hair cells depends on the age of the fish. Thus, in fish aged 4 days post fertilization (dpf), hair cells are less sensitive to neomycin than those of 5 dpf or more [74].

Van Trump et al. [75] also showed that hair cells in the lateral line are sensitive to damage caused by aminoglycosides. Research on the toxicity mechanisms of its molecules has shown that the swelling of mitochondria, loss of mitochondrial membrane potential, and the need for Bcl2 proteins associated with mitochondria suggest that aminoglycosides activate mitochondrial-dependent cell death pathways [76]. Also, the duration of hair cell death differs considerably for different aminoglycoside molecules, showing that toxicity pathways are initiated via the activation of different intracellular signaling cascades [46]. Actually, treatment with aminoglycosides lateral line of fish species has become a standard tool in behavioral studies designed to study lateral line function [77].

#### *4.2.3 Cisplatin*

The cisplatin toxicity study, which is a platinum-based chemotherapeutic agent used in the treatment of various tumors, has shown that there is a proportional

relationship between the dose and the time of hair cell loss [66]. Thus, hair cells damage is related to intracellular accumulation of cisplatin as well as little work that reports cascading activation of cell death in lateral line hair cells treated with cisplatin. Molecules with antioxidant properties, such as N-acetyl L-cysteine and D-methionine, exhibit protective capacity for hair cells in zebrafish exposed to cisplatin, suggesting that cisplatin can induce oxidative stress pathways. These oxidative stress pathways have been implicated in the ototoxicity of cisplatin and aminoglycosides, suggesting some conservation of cell death mechanisms between different classes of ototoxic drugs [78]. Nuclear condensation and mitochondrial swelling are the consequences of apoptotic cell death [66].

#### *4.2.4 Endocrine disruptors*

Endocrine disruptors chemicals (EDCs) are natural or synthetic compounds found in the environment that disrupt the levels and distribution of endogenous hormones in exposed organisms [79], which can alter development and/ or reproduction in humans and wildlife. Among the endocrine-disrupting compounds, we can cite natural hormones (17 beta-estradiol, E2) and synthetic [17α-ethinylestradiol (EE2)], pesticides [for example, dichlorodiphenyltrichloroethane (DDT)], polychlorinated biphenyls (PCB), bisphenol A (BPA), phthalates, flavonoids and polycyclic musks [80]. Zebrafish are used as a model organism to study endocrine disruptors and assess environmental risks [81]. The main mechanism of action triggered by these molecules is an agonist or antagonist interaction with ER. The concentration of vitellogenin (VTG), an egg yolk precursor protein, which is produced and secreted by the liver, absorbed by the ovary, and changed by developing eggs to form egg yolk, is the most common biomarker most widely used for EDC activity [81].

Bisphenol-A (BPA), polychlorinated biphenyls (PCBs), 17α-ethinylestradiol (EE2), and pesticides are widespread aquatic pollutants and can deeply affect the lateral line of zebrafish via disruption of endocrine system signaling. Bisphenol-A (BPA) is found in polycarbonate plastics and epoxy resins, as well as in the coatings of some cans. Researchers examined its impact on lateral line regeneration and found that it is toxic to hair cells in zebrafish larvae and that exposure delays regeneration [67]. BPA also attenuates hair cell regeneration after aminoglycoside damage, suggesting that BPA is toxic to supporting cells [71]. Hayashi et al. [67] found in another study that PCB-95 had no effect on lateral line development or hair cell survival. Rather, BPA had a significant effect on the survival and regeneration of hair cells. BPA-induced hair cell loss is both dose-dependent and temporal. Experimental laboratory studies suggest that BPA kills hair cells via activation of oxidative stress pathways, similar to previous reports of BPA toxicity in other tissues. In addition, Hayashi et al. [67] observed that hair cells killed with neomycin did not regenerate normally when BPA was present, suggesting that BPA in aquatic environments could interfere with innate regenerative responses in fish.

In other studies, Nasri et al. [68] examined the effect of the pesticide A6 derived from naturally-occurring α-terthienyl, and structurally related to the endocrinedisrupting pesticides anilinopyrimidines, on living zebrafish larvae. Results show that A6 decreases larval survival and affects central neurons at micromolar concentrations. In the lateral line system, researchers found that A6 alters the axonal and sensory cell regeneration at nanomolar concentrations. In addition, A6 has accumulated in lateral line neurons and hair cells. In addition, the examination of 17α-ethinylestradiol (EE2) effects at pico- to nano-molar concentrations on early nervous system development of zebrafish larvae showed that EE2 disrupted axonal nerve regeneration and hair cell regeneration. Upregulation of gene expression of

*Endocrine Disruptor Impact on Zebrafish Larvae: Posterior Lateral Line System as a New Target DOI: http://dx.doi.org/10.5772/intechopen.101609*

estrogen and progesterone receptors has been recorded. In contrast, downregulation of the tyrosine hydroxylase, involved in the synthesis of neurotransmitters, occurred concomitant with diminution of proliferating cells. Collectively, EE2 modifies nervous system development, both centrally and peripherally, with negative effects on regeneration and swimming behavior [69].

#### **5. Conclusions**

The lateral line is a sensory system utilized at a variety of aquatic vertebrates especially in zebrafish, to detect changes in surrounding water flow. This sense, which utilizes mechanotransduction, mediates a wide variety of behaviors from predator detection to schooling. Its position on the body surface allows experimental laboratory studies. The regeneration of functional mechanosensory cells after damage or following pollutants exposure offers the potential to uncover processes involved in the maintenance, proliferation, and differentiation of sensory precursors. The latest research has approved that estrogen receptors are involved in the control of lateral line development and regeneration. These results support the use of this sensory system as a target for research on environmental estrogenic endocrine disruptors.

#### **Conflict of interest**

The authors declare no conflict of interest.

#### **Author details**

Ahmed Nasri1,2\*, Patricia Aïssa1 , Ezzeddine Mahmoudi1 , Hamouda Beyrem1 and Véronique Perrier2

1 Faculty of Sciences of Bizerta (FSB), Laboratory of Environment Biomonitoring, University of Carthage, Bizerta, Tunisia

2 U1198 MMDN (Molecular Mechanisms of Neurodegenerative Diseases), Inserm (National Institute for Health and Medical Research), MGX (Montpellier GenomiX), BioCampus, University of Montpellier, Montpellier, France

\*Address all correspondence to: a7mednas@gmail.com; ahmed.nasri@fsb.u-carthage.tn

© 2021 The Author(s). Licensee IntechOpen. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/ by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

### **References**

[1] McLachlan JA. Environmental signaling: What embryos and evolution teach us about endocrine disrupting chemicals? Endocrine Reviews. 2001;**22**(3):319-341

[2] Hertz R. Accidental ingestion of estrogens by children. Pediatrics. 1958;**21**(2):203-206

[3] Kavlock RJ et al. Research needs for the risk assessment of health and environmental effects of endocrine disruptors: A report of the US EPAsponsored workshop. Environmental Health Perspectives. 1996;**104**: 715-740

[4] Birnbaum LS. State of the science of endocrine disruptors. 2013:a107-a107. DOI: 10.1289/ehp.1306695

[5] Sumpter JP. Xenoendocrine disrupters - Environmental impacts. Toxicology Letters. 1998;**103**:337-342

[6] Vos JG et al. Health effects of endocrine-disrupting chemicals on wildlife, with special reference to the European situation. Critical Reviews in Toxicology. 2000;**30**(1):71-133

[7] Jobling S et al. Widespread sexual disruption in wildfish. Environmental Science and Technology. 1998;**32**(17): 2498-2506

[8] Nash JP et al. Long-term exposure to environmental concentrations of the pharmaceutical ethinylestradiol causes reproductive failure in fish. Environmental Health Perspectives. 2004;**112**(17):1725-1733

[9] Mills LJ, Chichester C. Review of evidence: are endocrine-disrupting chemicals in the aquatic environment impacting fish populations ? Science of the Total Environment. 2005;**343**(1-3): 1-34

[10] Van Trump WJ, McHenry MJ. The morphology and mechanical sensitivity of lateral line receptors in zebrafish larvae (*Danio rerio*). Journal of Experimental Biology. 2008;**211**(13): 2105-2115

[11] Ghysen A, Dambly-Chaudiere C. Development of the zebrafish lateral line. Current Opinion in Neurobiology. 2004;**14**(1):67-73

[12] Nicolson T. The genetics of hearing and balance in zebrafish. Annual Review of Genetics. 2005;**39**:9-22

[13] Raible DW, Kruse GJ. Organization of the lateral line system in embryonic zebrafish. Journal of Comparative Neurology. 2000;**421**(2):189-198

[14] Gompel N et al. Pattern formation in the lateral line of zebrafish. Mechanisms of Development. 2001;**105**:69-77

[15] Chitnis AB, Nogare DD, Matsuda M. Building the posterior lateral line system in zebrafish. Developmental Neurobiology. 2012;**72**:234-255

[16] Wibowo I et al. Compartmentalized Notch signaling sustains epithelial mirror symmetry. Development. 2011;**138**:1143-1152

[17] Harris JA et al. Neomycin-induced hair cell death and rapid regeneration in the lateral line of zebrafish (*Danio rerio*). Journal of the Association for Research in Otolaryngology. 2003;**4**(2): 219-234

[18] Haas P, Gilmour D. Chemokine signaling mediates selforganizing tissue migration in the zebrafish lateral line. Developmental Cell. 2006;**10**(5):673-680

[19] Dambly-Chaudière C, Cubedo N, Ghysen A. Control of cell migration in the development of the posterior lateral *Endocrine Disruptor Impact on Zebrafish Larvae: Posterior Lateral Line System as a New Target DOI: http://dx.doi.org/10.5772/intechopen.101609*

line: Antagonistic interactions between the chemokine receptors CXCR4 and CXCR7/RDC1. BMC Developmental Biology. 2007;**7**:23

[20] Venkiteswaran G et al. Generation and dynamics of an endogenous, self-generated signaling gradient across a migrating tissue. Cell. 2013;**155**(3): 674-687

[21] Xu H et al. Gb1 controls collective cell migration by regulating the protrusive activity of leader cells in the posterior lateral line primordium. Developmental Biology. 2014;**385**(2): 316-327

[22] Laguerre L, Ghysen A, Dambly-Chaudière C. Mitotic patterns in the migrating lateral line cells of zebrafish embryos. Developmental Dynamics. 2009;**238**(5):1042-1051

[23] Gamba L et al. Lef1 controls patterning and proliferation in the posterior lateral line system of zebrafish. Developmental Dynamics. 2010;**239**(12):3163-3171

[24] Matsuda M et al. Lef1 regulates Dusp6 to influence neuromast formation and spacing in the zebrafish posterior lateral line primordium. Development (Cambridge, England). 2013;**140**(11):2387-2397

[25] Aman A, Piotrowski T. Supplemental data Wnt/b-catenin and fgf signaling control collective cell migration by restricting chemokine receptor expression. Gene Expression. 2008;**15**(5):749-761

[26] Lecaudey V et al. Dynamic Fgf signaling couples morphogenesis and migration in the zebrafish lateral line primordium. Development (Cambridge, England). 2008;**135**:2695-2705

[27] Galanternik MV, Kramer KL, Piotrowski T. Heparan sulfate proteoglycans regulate Fgf signaling and cell polarity during collective cell migration. Cell Reports. 2015;**10**(3): 414-428

[28] Yanicostas C et al. Essential requirement for zebrafish anosmin-1a in the migration of the posterior lateral line primordium. Developmental Biology. 2008;**320**(2):469-479

[29] Hardelin JP, Dodé C, et al. The complex genetics of Kallmann syndrome: KAL1, FGFR1, FGF8, PROKR2, PROK2. Sexual Development. 2008;**2**(4-5):181-193

[30] Menuet A et al. Molecular characterization of three estrogen receptor forms in zebrafish: Binding characteristics, transactivation properties, and tissue distributions. Biology of Reproduction. 2002;**66**(6): 1881-1892

[31] Tingaud-Sequeira A et al. Expression patterns of three estrogen receptor genes during zebrafish (Danio rerio) development: Evidence for high expression in neuromasts. Gene Expression Patterns. 2004;**4**(5):561-568

[32] Froehlicher M et al. Estrogen receptor subtype β2 is involved in neuromast development in zebrafish (Danio rerio) larvae. Developmental Biology. 2009;**330**(1):32-43

[33] Celeghin A et al. The knockdown of the maternal estrogen receptor 2a (esr2a) mRNA affects embryo transcript contents and larval development in zebrafish. General and Comparative Endocrinology. 2011;**172**(1):120-129

[34] López-Muñoz A et al. Estrogen receptor 2b deficiency impairs the antiviral response of zebrafish. Developmental & Comparative Immunology. 2015;**53**(1):55-62

[35] O'Brien GS et al. Developmentally regulated impediments to skin reinnervation by injured peripheral

sensory axon terminals. Current Biology. 2009;**19**(24):2086-2090

[36] Villegas R et al. Dynamics of degeneration and regeneration in developing zebrafish peripheral axons reveals a requirement for extrinsic cell types. Neural Development. 2012;**7**(1): 1-14

[37] Gilley J, Coleman MP. Endogenous Nmnat2 is an essential survival factor for maintenance of healthy axons. PLOS Biology. 2010;**8**(1):e1000300

[38] Nguyen CT et al. The PAF1 complex component Leo1 is essential for cardiac and neural crest development in zebrafish. Developmental Biology. 2010;**341**(1):167-175

[39] Luo L, O'Leary DD. Axon retraction and degeneration in development and disease. Annual Review of Neuroscience. 2005;**28**:127-156

[40] Coleman MP, Freeman MR. Wallerian degeneration, wlds, and nmnat. Annual Review of Neuroscience. 2010;**33**:245-267

[41] White FV et al. Lipid metabolism during early stages of Wallerian degeneration in the rat sciatic nerve. Journal of Neurochemistry. 1989;**52**(4): 1085-1092

[42] Trapp BD, Hauer P, Lemke G. Axonal regulation of myelin protein mRNA levels in actively myelinating Schwann cells. Journal of Neuroscience. 1988;**8**(9):3515-3521

[43] Kaus S. The effect of aminoglycoside antibiotics on the lateral line organ of *Aplocheilus lineatus* (Cyprinodontidae). Acta Oto-Laryngologica. 1987;**103**(3-4): 291-298

[44] Song J, Yan HY, Popper AN. Damage and recovery of hair cells in fish canal (but not superficial) neuromasts after gentamicin exposure. Hearing Research. 1995;**91**(1):63-71

[45] Williams JA, Holder N. Cell turnover in neuromasts of zebrafish larvae. Hearing Research. 2000;**143**(1):171-181

[46] Coffin AB et al. Extracellular divalent cations modulate aminoglycoside-induced hair cell death in the zebrafish lateral line. Hearing Research. 2009;**253**(1):42-51

[47] Metcalfe WK, Kimmel CB, Schabtach E. Anatomy of the posterior lateral line system in young larvae of the zebrafish. Journal of Comparative Neurology. 1985;**233**(3):377-389

[48] Rubel EW, Furrer SA, Stone JS. A brief history of hair cell regeneration research and speculations on the future. Hearing Research. 2013;**297**:42-51

[49] Sun H, Lin CH, Smith ME. Growth hormone promotes hair cell regeneration in the zebrafish (Danio rerio) inner ear following acoustic trauma. PLoS One. 2011;**6**(11):e28372

[50] Faucher K et al. Damage and functional recovery of the Atlantic cod (*Gadus morhua*) inner ear hair cells following local injection of gentamicin. International Journal of Audiology. 2009;**48**(7):456-464

[51] Hernandez PP et al. Sub-lethal concentrations of waterborne copper are toxic to lateral line neuromasts in zebrafish (Danio rerio). Hearing Research. 2006;**213**(1):1-10

[52] Ma EY, Rubel EW, Raible DW. Notch signaling regulates the extent of hair cell regeneration in the zebrafish lateral line. Journal of Neuroscience. 2008;**28**(9):2261-2273

[53] Hernández PP et al. Regeneration in zebrafish lateral line neuromasts: Expression of the neural progenitor cell marker sox2 and proliferationdependent and-independent mechanisms of hair cell renewal. Developmental Neurobiology. 2007;**67**(5):637-654

*Endocrine Disruptor Impact on Zebrafish Larvae: Posterior Lateral Line System as a New Target DOI: http://dx.doi.org/10.5772/intechopen.101609*

[54] Schuck JB et al. Transcriptomic analysis of the zebrafish inner ear points to growth hormone mediated regeneration following acoustic trauma. BMC Neuroscience. 2011;**12**(1):88

[55] Liang J et al. The stat3/socs3a pathway is a key regulator of hair cell regeneration in zebrafish stat3/socs3a pathway: Regulator of hair cell regeneration. Journal of Neuroscience. 2012;**32**(31):10662-10673

[56] Hertzano R et al. Transcription profiling of inner ears from Pou4f3 ddl/ ddl identifies Gfi1 as a target of the Pou4f3 deafness gene. Human Molecular Genetics. 2004;**13**(18):2143-2153

[57] Namdaran P et al. Identification of modulators of hair cell regeneration in the zebrafish lateral line. Journal of Neuroscience. 2012;**32**(10):3516-3528

[58] Head JR et al. Activation of canonical Wnt/β-catenin signaling stimulates proliferation in neuromasts in the zebrafish posterior lateral line. Developmental Dynamics. 2013;**242**(7): 832-846

[59] Jacques BE et al. The role of Wnt/βcatenin signaling in proliferation and regeneration of the developing basilar papilla and lateral line. Developmental Neurobiology. 2014;**74**(4):438-456

[60] Wada H et al. Wnt/Dkk negative feedback regulates sensory organ size in zebrafish. Current Biology. 2013;**23**(16): 1559-1565

[61] Jiang L et al. Gene-expression analysis of hair cell regeneration in the zebrafish lateral line. Proceedings of the National Academy of Sciences. 2014;**111**(14):E1383-E1392

[62] Millimaki BB, Sweet EM, Riley BB. Sox2 is required for maintenance and regeneration, but not initial development, of hair cells in the zebrafish inner ear. Developmental Biology. 2010;**338**(2):262-269

[63] Lush ME, Piotrowski T. Sensory hair cell regeneration in the zebrafish lateral line. Developmental Dynamics. 2014;**243**(10):1187-1202

[64] Moon IS et al. Fucoidan promotes mechanosensory hair cell regeneration following amino glycoside-induced cell death. Hearing Research. 2011;**282**(1): 236-242

[65] Walters BJ et al. Auditory hair cell-specific deletion of p27Kip1 in postnatal mice promotes cellautonomous generation of new hair cells and normal hearing. Journal of Neuroscience. 2014;**34**(47):15751-15763

[66] Ou HC, Raible DW, Rubel EW. Cisplatin-induced hair cell loss in zebrafish (Danio rerio) lateral line. Hearing Research. 2007;**233**(1):46-53

[67] Hayashi L et al. The effect of the aquatic contaminants bisphenol-A and PCB-95 on the zebrafish lateral line. Neurotoxicology. 2015;**46**:125-136

[68] Nasri A et al. Neurotoxicity of a biopesticide analog on zebrafish larvae at nanomolar concentrations. International Journal of Molecular Sciences. 2016;**17**(12):2137

[69] Nasri A et al. Ethinylestradiol (EE2) residues from birth control pills impair nervous system development and swimming behavior of zebrafish larvae. Science of The Total Environment. 2021;**770**:145272

[70] Olivari FA, Hernández PP, Allende ML. Acute copper exposure induces oxidative stress and cell death in lateral line hair cells of zebrafish larvae. Brain Research. 2008;**1244**:1-12

[71] Mackenzie SM, Raible DW. Proliferative regeneration of zebrafish lateral line hair cells after different ototoxic insults. 2012;**7**(10):e47257

[72] Montgomery JC, Coombs S, Baker CF. The mechanosensory lateral line system of the hypogean form of Astyanax fasciatus. The biology of hypogean fishes. Dordrecht: Springer; 2001. pp. 87-96

[73] Faucher K et al. Impact of acute cadmium exposure on the trunk lateral line neuromasts and consequences on the "C-start" response behaviour of the sea bass (Dicentrarchus labrax L.; Teleostei, Moronidae). Aquatic Toxicology. 2006;**76**(3-4):278-294

[74] Santos F et al. Lateral line hair cell maturation is a determinant of aminoglycoside susceptibility in zebrafish (Danio rerio). Hearing Research. 2006;**213**(1):25-33

[75] Van Trump WJ et al. Gentamicin is ototoxic to all hair cells in the fish lateral line system. Hearing Research. 2010;**261**(1):42-50

[76] Owens KN et al. Ultrastructural analysis of aminoglycoside-induced hair cell death in the zebrafish lateral line reveals an early mitochondrial response. Journal of Comparative Neurology. 2007;**502**(4):522-543

[77] Coombs S. Smart skins: Information processing by lateral line flow sensors. Autonomous Robots. 2001;**11**(3):255-261

[78] Rybak LP, Ramkumar V. Ototoxicity. Kidney International. 2007;**72**(8): 931-935

[79] Colborn T, Vom Saal FS, Soto AM. Developmental effects of endocrinedisrupting chemicals in wildlife and humans. Environmental Health Perspectives. 1993;**101**(5):378-384

[80] Gore AC et al. EDC-2: The Endocrine Society's second scientific statement on endocrine-disrupting chemicals. Endocrine Reviews. 2015;**36**(6):E1-E150

[81] Dang M et al. Long-term drug administration in the adult zebrafish using oral gavage for cancer preclinical studies. Disease Models & Mechanisms. 2016;**9**(7):811-820

#### **Chapter 2**

## Role and Effect of Persistent Organic Pollutants to Our Environment and Wildlife

*Nisha Gaur, Dhiraj Dutta, Aman Jaiswal, Rama Dubey and Dev Vrat Kamboj*

#### **Abstract**

Persistent organic pollutants (POPs) are toxic substances composed of carbon-based chemical compounds and mixtures. In the recent times, persistent organic pollutants (POPs) came as a threat for the wildlife and environmental world. POPs are chemically stable, remain intact in the environment for long periods, are recalcitrant in nature, and are lipophilic in nature. Therefore, they accumulate in fatty tissue of living organisms and reside longer period of time finally affecting the human and wildlife. It is believed that approximately 90% of human beings are exposed to POPs from their diets that contain animal products. These foods lead to bioaccumulation in fat tissues that then cause health hazard. There are many studies such that its behavior in photocatalytic oxidation reactions are available; also many research studies are going to combat from these toxic substances. In this chapter, we will take you through how persistent organic pollutants are affecting our environment and wildlife and what are its roles.

**Keywords:** persistent organic pollutants, food chain, grasshopper, health, environment, Stockholm convention, global warming, marine

#### **1. Introduction**

As the name implies, persistent organic pollutants (POPs) are extremely persistent in the atmosphere, with a half-life of over a decade in the soil, sediments, air, and biota [1]. The importance of research into persistent organic pollutants is illustrated by the Stockholm Convention, adopted in 2001 by conference of Plenipotentiaries and came into force on May 17, 2004. POPs have now become the focus of different growing national and international concern as they show toxic effects on animal reproduction, development, and immunological function. Some national agencies are still not taking it seriously and call it differently as "Bio accumulative chemicals of concern" (BCCs) [2]. Only those compounds that get the extreme end of the distribution in degree of persistence, mobility, and toxicity will be ranked as POPs.

Chlorinated substances stay in the environment for a long time. With the introduction of electron capture detectors, several organochlorine pesticides such as dichlorodiphenyltrichloroethane (DDT), dichlorodiphenyldichloroethane (DDE), dieldrin, and toxaphene began to be detected [3]. These have been reported at such places where never been used before such as the earth's pole.


#### **Table 1.**

*POPs recognized in Stockholm convention [4].*

These bio-accumulate in the food chain (animals and humans), causing a slew of well-known health and environmental problems. These pollutants are causing great concern among scientists, governments, and nongovernmental organizations (NGOs). There are different types of POPs such as heptachlor, chlordane, aldrin, dieldrin, hexachlorobenzene, endrin, mirex, chlordecone, toxaphene, lindane, hexa- and penta-bromodiphenyl ethers (commercial octabromodiphenyl ether), tetra- and penta-bromodiphenyl ethers (commercial pentabromodiphenyl ether), hexabromobiphenyl, pentachlorobenzene, polychlorinated biphenyls (PCBs), α- and β-hexachlorocyclohexane, α- and β-endosulfans (technical endosulfan and its isomers), perfluorooctane sulfonic acid and its salts (PFOS), perfluorooctane sulfonyl fluoride (PFOSF), DDT, pentachlorobenzene, hexachlorobenzene, polychlorinated dibenzofurans (PCDFs), polychlorinated dibenzo-p-dioxins (PCDDs), polychlorinated biphenyls (PCBs), hexachlorobutadiene, chlorinated naphthalenes, pentachlorophenol hexabromocyclododecane (HBCD), short-chained chlorinated paraffins were recognized by United Nations Environment Program (UNEP), Montreal, International POPs Elimination Network (IPEN), and Stockholm Convention. These pollutants are classified into four categories based on their toxicity and are shown in **Table 1** [4].

#### **2. POPs: a general aspect**

It is important to discuss the aspects of POPs before coming to the effects of these to understand the notorious effects of POPs.

Persistent chemicals generally have higher concentrations because they usually present in the environment for longer time and get their steady state. If emissions decline to near-zero after a particular duration of use, the amount of chemical in the environment will decrease exponentially. The overall transformation rate constant in the environment determines the rate of elimination, and persistent substances are removed more slowly [1].

The impact of POPs can be well understood by its deposition process. This innovative method combines aerosol remote sensing data with POP aerosol phase

#### *Role and Effect of Persistent Organic Pollutants to Our Environment and Wildlife DOI: http://dx.doi.org/10.5772/intechopen.101617*

observations. By this process, POPs impact remote oceanic areas, raising environmental concerns because of their toxicity and accumulation in aquatic food webs. Because POPs are semi-volatile chemicals that can exist in both the gas and aerosol phases, precipitation scavenging will deposit POPs in both dissolved and particulate forms [5]. Persistent organic pollutants separate into gas and aerosol phases once they are released and are subject to long-range atmospheric transport (LRAT). In the transport and fate of POPS at the regional and global scale, atmospheric depositional processes play a key role. Then, by adopting any of the three primary processes such as (1) dry deposition of particulate-bound pollutants, (2) diffusive gas exchange between the atmospheric boundary layer (ABL), and (3) the surface ocean and rain scavenging (either from gas or particulate phases), transport of semi-volatile organic compounds from the atmosphere to the ocean takes place [6]. Additionally, this can be concluded that the dry aerosol and gaseous deposition contribute to aquatic ecosystems pollutant burden and finally support POP accumulation in marine food web.

Climate change has already sparked a slew of environmental issues, and there is a direct link between pollutant emissions, dispersion, and toxicity and climate change. Both the IPCC report and the annual report of the United Nations Environment Program (UNEP) stressed the importance of paying attention to the problem of environmental pollution, particularly in light of global warming [7]. The series of POPs has already been discussed above, and the transport processes for persistent organic pollutants can be seen in **Figure 1**. POPs are transported globally in two ways: atmospheric circulation and ocean currents. POPs can exist in both gaseous and particle forms in the atmosphere. As a result, POPs in the atmosphere can be distributed globally via gaseous and particulate thanks to atmospheric circulation. POPs are more likely to be deposited on the land surface when the temperature drops while they will evaporate back into the atmosphere and migrate again, when the temperature rises.

POPs are transported globally in two ways: atmospheric circulation and ocean currents. POPs can exist in both gaseous and particle forms in the atmosphere. Hence, gaseous and particulate POPs in atmosphere can be globally spread with the help of atmospheric circulation.

**Figure 1.** *Fate of persistent organic pollutants in the environment.*

This cycle repeats itself, allowing POPs to be transported and deposited in far-flung locations. The grasshopper effect is what it is called. Furthermore, some POPs with a significantly higher solubility, such as hexachlorocyclohexanes (HCH) and perfluoro octane sulfonates (PFOS), can penetrate surface waters, feed into ocean currents, and travel around the world. Changes in climatic parameters such as temperature, wind speed, wind direction, and precipitation occur as a result of climate change. The intensity and pathways of POPs transported by air and ocean will undoubtedly change as these conditions change [8].

#### **3. United Nations Stockholm convention**

The text of the Stockholm Convention was adopted by the Conference of the Plenipotentiaries on May 22, 2001 and came into effect on May 17, 2004. In May 1995, the UNEP Governing Council recognized that persistent organic pollutants (POPs) pose serious and escalating dangers to human health and the environment and recommended that an international assessment process of an initial list of 12 POPs be performed in its resolution 18/32. The Intergovernmental Forum on Chemical Safety (IFCS) develops international action recommendations that must be considered by the UNEP Governing Council and the World Health Assembly by 1997. The Stockholm Convention's 12 initial POPs were divided into three categories.


Another mechanism known as bioaccumulation concentrates POPs in living creatures [9]. The term "bioaccumulation" has two meanings: a) to define a dynamic process in which an element or compound is passively or actively taken up and concentrated within an organism; b) to indicate a currently high concentration as a result of prior accumulation activities [10].

### **4. Role and effect of persistent organic pollutants**

Many developed countries have taken initiatives in recent years to restrict and limit the flow of PCBs into the atmosphere. The most important element in establishing these limits was a 1973 suggestion from the Organization for Economic Co-operation and Development (OECD) (WHO, 1976; IARC, 1978; OECD, 1982). Since then, the 24 OECD member countries have set manufacture, sales, importation, exportation, and use limitations, as well as a labeling system. POPs have severe effect on wildlife as well as on environment and hence to humans also. Reproductive impairment and malformations, increased risk of tumors, altered liver enzyme function are some of the many bad effects of POPs.

In **Figure 2**, it can be seen that POPs have an anthropogenic origin and are discharged into the air, water, and land, where they deposit in water and eventually enter the food chain through sediment. These are disseminated over the world by air and ocean currents and hence, travel a long distance. They subsequently penetrate atmospheric processes, air-water exchange, and cycles such as rain, snow, and dry particles, exposing even the most remote groups of humans and animals who rely on aquatic food.

#### **4.1 Effect of POPs on environment**

As of now, we know that POPs are a global hazard since POPs discharged into the environment can travel a considerable distance through the air and water via evaporation and redeposition from their initial source. The most important factor in the transportation of global POPs is the atmosphere. Because of the semi-volatile nature of the atmosphere, these chemicals are found in atmospheric gases. Once these POPs get encountered into the gases, they go under some other processes such that degradation, soil deposition, vegetation, bioaccumulation, sedimentation, and many more (**Figure 3**). Because these POPs are temperature-dependent, the Global Distillation Effect theory predicts that gas-phase contaminants will be transported from warm regions of the planet, such as tropical or temperate source areas, to colder, higher-latitude regions, affecting vapor pressure and Henry's constant,

#### **Figure 3.**

*Environmental progressions during long-range atmospheric transport of POPs.*

resulting in condensation and accumulation of POPs in soil, vegetation, and other places, from where they can enter into the food chains [11].

Because reactivity, adsorption, and accumulation are all temperature-dependent activities, they can be influenced by climate change at any point along the transport and redistribution paths. Climate-change-related activities are predicted to modify POP exposure patterns for native and resident human populations in the long run. The majority of POPs have been produced and released in enormous quantities in the Northern Hemisphere, primarily by agriculture and industry. The emission pattern for most legacy POPs began around 1940–1950, according to published data on POPs. Following a significant increase in emissions, some countries banned or restricted the use of these compounds or found techniques to eliminate them as byproducts, resulting in a period of reduction (about 1970). Variations in climate and ambient temperature have a direct impact on a number of deciding environmental elements as summarized in **Table 2** [12].

Several persistent organic pollutants (POPs), such as polychlorinated biphenyls (PCBs) and organochlorine pesticides (OCPs), have been widely utilized in Asian developing countries for industry, agriculture, and vector control throughout the last few decades. Pesticide usage in India is at 85,000 metric tons per year, with DDTs, HCHs, and malathion accounting for 70% of that [13]. Due to a lack of efficient facilities in Asian developing countries, substantial amounts of municipal trash from populous regions are directly thrown into open dumping sites with poor management. The public is concerned about the potential negative consequences for local communities and the environment as a result of this behavior. When recent rigorous research revealed increased human health risk from exposure to harmful substances such as dioxins and similar compounds, as well as heavy metals in these dumping sites, these worries became more serious [14].

Since these POPs are highly volatile and resistant to photolytic, biological, and chemical degradation, they were found at high concentrations around the globe, including open oceans, deserts, the Arctic, and the Antarctic. Many studies have looked at the spread of PCBs around the world. POP levels were found to be extremely high. Many studies have looked at the spread of PCBs around the world. Few studies have discovered high levels of some organochlorines in ocean, rainfall, and wild animals. The concentration of HCB in Antarctic fish was found to be comparable to that of North Sea fish in a study [15].


*Role and Effect of Persistent Organic Pollutants to Our Environment and Wildlife DOI: http://dx.doi.org/10.5772/intechopen.101617*

**Table 2.**

*Influences of climate change on environmental processes related to transport and distribution of POPs [12].*

#### **4.2 Effect of POPs on wildlife**

As we previously discussed, POPs are organic molecules with high lipid solubility, allowing them to stay in the environment for long periods of time, be transported substantial distances from their source, and bio-accumulate in food chains. They are deemed to have a danger of generating negative effects on human and wildlife health because of these traits.

Multiple exposure routes can expose wildlife species living in contaminated areas to complex combinations of pollutants. Wild species can be utilized as biomonitors of environmental pollution in a place since they have varied ecological, etiological, and physiological properties. Hazardous material exposure at various organizational levels also threatens the long-term viability of wildlife populations. As a result, pollution is currently considered one of the most serious threats to biological diversity. Exposure and effect biomarkers can be examined and integrated simultaneously to provide more information about probable toxicity pathways and ecosystem component health [16].

Marine wildlife numbers, particularly megafauna species, have dropped dramatically in recent decades, according to several research studies. For example, census studies in Eastern Australian seas show that loggerhead turtle numbers have plummeted by up to 86% during the 1970s. Similarly, in the last 90 years, the global population of dugongs appears to have fallen by at least 20%. (approximately three dugong generations). These and many more marine wildlife species have been added to the Red List of Threatened Species as a result of their critical population status, which spans from endangered to vulnerable to extinct, and are high conservation priorities [17].

Disease has been highlighted as a major contributor to the loss of marine wildlife populations, among the many concerns. Chronic sickness, which can lead to mortality, has been found to be on the rise in marine wildlife populations, reaching panzootic levels in some cases [18]. POPs are distributed in the marine environment as a result of subsequent transport mechanisms and source discharges. An increasing corpus of research examines the relationship between tissue loads of dominating POP groups and functional outcomes that have been degraded (**Table 3**).


*AhR: aryl hydrocarbon receptor; p,p′-DDE: 4,4′-dichlorodiphenyldichloroethylene; DDTs: dichlorodiphenyltrichloroethane; HCHs: hexachlorocyclohexanes- hexanes; OCPs: organochlorine pesticides; PCBs: polychlorinated biphenyls; PBDEs: polybrominated diphenyl ethers. TT3: total triiodothyrine; FT3: free triiodothyrine; TT4: total thyroxine; FT4: free thyroxine; T3: triiodothyrine; cyp1A1: cytochrome P450A1.*

#### **Table 3.**

*Summary of correlative studies associating functional health parameters with body burdens of persistent organic pollutants in marine wildlife.*

Hydrophobic (water-hating) and lipophilic (fat-loving) compounds are the most common types of POPs. They bond firmly to solids, particularly organic matter, in both marine and terrestrial environments, avoiding the aqueous section. They also enter lipids more easily than the aqueous system inside cells and accumulate in fatty tissue. Chemicals are stored in fatty tissue, allowing them to remain in biota where metabolism is slow. As a result, POPs may move up the food chain. Under ambient temperatures, POPs tend to shift into the gas phase. As a result, they may volatilize from soils, plants, and aquatic systems into the air and migrate vast distances before being re-deposited due to their resistance to breakdown reactions in air [28].

POP loads can be passed from the mother to the child not only during placental development, but also through breast feeding in placental viviparous species, which are the only placental mammals. The importance of maternal transfer in terms of early exposure should not be ignored; various studies have shown that juvenile placental mammals acquire higher amounts of PCDD/Fs or PCBs via milk than they do from prenatal exposure during placental development. According to the studies, the offspring of pregnant rats administered six PCDD/Fs, including 2,3,7,8,-TCDD/F and the non-ortho PCBs 77, 126, and 169, got 7–28% of their doses lactationally and just 0.5–3% through the placenta. Fasting mothers can increase their children's dietary exposure to POPs from milk. Although the trend in female polar bear body burdens was not consistent—DDT and HCH decreased during fasting, while

#### *Role and Effect of Persistent Organic Pollutants to Our Environment and Wildlife DOI: http://dx.doi.org/10.5772/intechopen.101617*

chlordane and PCBs increased—the ratio of plasma/adipose tissue and milk/adipose tissue OC concentrations did not change during the fast, indicating that POPs were probably at steady state among the various tagging systems [29].

All birds, most reptiles, most amphibians, and the rare monotreme mammals, for example, will deliver POPs to their eggs via maternal transfer. The direct transfer of the contaminant burden from the female to the eggs via the reallocation of the female's lipid storage is the principal source of POP exposure to the growing embryo. Both biological processes and chemical features of the pollutants induce the deposition of lipids and proteins (together with the POPs associated with the lipids and proteins) in the developing egg. The energy required for egg production can come from the female's older body reserves, her energy consumption during the egg formation period, or a combination of both [29].

#### **4.3 Effect of POPs on human health**

Among the many POPs that are abundant in our surroundings, a "black list" of POPs has been recognized under the diplomatic signature of the Stockholm Convention in 2001. Pesticides, such as aldrin, DDT, dieldrin, endrin, heptachlor, chlordane, mirex, and toxaphene; industrial compounds, such as hexachlorobenzene and polychlorinated biphenyls (PCBs); and other chemical by-products, such as polychlorinated dibenzodioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs)—the general name "dioxins" is used for PCDDs and PCDFs. These POPs are known to be especially hazardous, having a high proclivity for biomagnification in the food chain, and have been linked to carcinogenic and endocrine disrupting effects in a variety of biota [9].

POP residues have been found in human adipose tissue from people living in a variety of countries, including Europe, Asia, Africa, and North America, for many years. Pesticide toxicity and persistence are beneficial for killing their target organisms, but they can cause difficulties for humans and the environment. Organochlorine pesticides (OCPs) have been found in the environment and, as a result, in the food chains of humans and wildlife since the early 1960s. Dichlorodiphenyltrichloroethane (DDT) and its derivatives have already been discovered in almost all environmental media and are the most common OCPs detected in human tissues, particularly adipose tissue. DDT is an organochlorine chemical initially synthesized in 1874 in Germany. The insecticidal effects of DDT were discovered in 1939, and commercial use began in 1945. DDT is dechlorinated in the human body to tetrachlorodiphenylethane (DDD), which is water-soluble and less hazardous to humans. Dichlorodiphenyldichloroethanes (DDEs) are another class of DDT derivatives that quickly accumulate in human adipose tissue and constitute a considerable health risk due to their long half-life. DDE can be acquired through DDT metabolism in the body or from intake of DDE-tainted foodstuffs [30].

While human data analysis has raised concerns, it has yet to produce conclusive evidence of causal relationships between low-level exposure to environmental chemicals, endocrine disrupting activities, and harmful human health impacts. All relevant data, including experimental animal data and wildlife observations, must be examined. The difficulties of correlating prenatal, postnatal, and childhood exposure to adult functioning are particularly concerning [31].

Reduced semen quality (i.e., reduced numbers, motility, and altered morphology of sperm), male reproductive tract abnormalities (e.g., hypospadias and cryptorchidism), altered sex ratio, endometriosis, precocious puberty, and early menarche) have been the focus of much of the human health concerns resulting from EDC exposure. A reduction in sperm counts has been recorded in a number of research studies in a lot of countries [32].

Increased rates of some hormone-related malignancies in many regions of the world are frequently cited as proof that EDC exposure has had negative health consequences. Increases in breast cancer and testicular cancer are particularly concerning. Several human epidemiological studies and experimental laboratory investigations have been undertaken to see if organochlorine pesticides are linked to an increased risk of breast cancer [31].

Organometals that bind to protein, particularly organomercurials; lipophilic contaminants such as dioxins, PCBs, polybrominated diphenyl ethers (PBDEs), and chlorinated pesticides; and persistent non-lipophilic compounds such as per fluorinated compounds used as repellents are all sources of concern for human health. Toxicity does not necessitate the persistence of a chemical. Many volatile organic chemicals, phthalates, and bisphenol A, which are present and leach from typical industrial products, are also found throughout the food supply and in the bodies of many of the world's populations. It has recently been discovered that exposure to certain chlorinated POPs increases the likelihood of developing type 2 diabetes, an insulin-related disease [33]. Surprisingly, this increased risk appears to occur at extremely low concentrations and does not appear to follow a linear dose–response curve [34]. DDE and hexachlorobenzene appear to have the strongest links, although PCBs also greatly enhance the risk [35]. Because type 2 diabetes is an insulin receptor disease, the processes underlying this link are unknown; however, it is most likely the result of gene induction. Obesity is frequently cited as a major risk factor for diabetes. Obese people who do not have high levels of POPs, on the other hand, do not have an increased risk of diabetes, according to certain research [33].

#### **5. Conclusion**

By now it is very much clear that how severe the POPs can be in future for the upcoming generations. It is not affecting only environment but also through environment to wildlife animals and directly or indirectly to human. Because of rising industrial use, persistent organic pollutants are becoming a major concern, causing their accumulation and persistence in living beings and the environment. Its exposure has numerous deadly implications for organs and tissues, including oxidative stress and cell death, due to several mechanisms. Several international plans have been developed to reduce the use of POPs and eliminate them completely; however, many developing countries around the world continue to ignore the Aarhus and Stockholm conventions. All of the class 1 and 2 POPs listed by these conventions must be banned. There are many agencies and NGOs working to accomplish their goals, but as humans, we must also recognize how damaging a single error can be. When scientific information is unknown, decision-making on the possible detrimental effects of chemical exposure is increasingly regulated by the precautionary principle, both internationally and nationally. The impact of endocrine disruptors on negative health consequences has yet to be proven conclusively. However, it is evident that the risk of endocrine disruptors is significant at specific times (preconception, pregnancy, and postpartum), and further research and development are needed to determine the health problems that should always be prioritized.

#### **Conflict of interest**

The authors declare no conflict of interest.

*Role and Effect of Persistent Organic Pollutants to Our Environment and Wildlife DOI: http://dx.doi.org/10.5772/intechopen.101617*

### **Author details**

Nisha Gaur1 , Dhiraj Dutta1 , Aman Jaiswal<sup>2</sup> , Rama Dubey1 \* and Dev Vrat Kamboj1

1 Defence Research Laboratory, DRDO, Tezpur, Assam, India

2 Indian Institute of Science Education and Research Mohali, Punjab, India

\*Address all correspondence to: r\_dubey172@rediffmail.com

© 2022 The Author(s). Licensee IntechOpen. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/ by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

### **References**

[1] Jones KC, De Voogt P. Persistent organic pollutants (POPs): State of the science. Environmental Pollution. 1999;**100**(1):209-221

[2] Wania F, Mackay D. Tracking the Distribution of persistent organic pollutants. Environmental Science & Technology. 1996;**30**(9):390-396. DOI: 10.1016/s0926-3373(97)80026-4

[3] Muir DCG, Howard PH. Are there other persistent organic pollutants? A challenge for environmental chemists. Environmental Science & Technology. 2006;**40**(23):7157-7166. DOI: 10.1021/ es061677a

[4] Alharbi OML, Basheer AA, Khattab RA, Ali I. Health and environmental effects of persistent organic pollutants. Journal of Molecular Liquids. 2018;**263**:442-453. DOI: 10.1016/j.molliq.2018.05.029

[5] Jurado E, Jaward F, Lohmann R, Jones KC, Simó R, Dachs J. Wet deposition of persistent organic pollutants to the global oceans. Environmental Science & Technology. 2005;**39**(8):2426-2435. DOI: 10.1021/ es048599g

[6] Jurado E, Jaward FM, Lohmann R, Jones KC, Simó R, Dachs J. Atmospheric dry deposition of persistent organic pollutants to the Atlantic and inferences for the global oceans. Environmental Science & Technology. 2004;**38**(21): 5505-5513. DOI: 10.1021/es049240v

[7] Climate Change 2013: The Physical Science Basis. Available from: https:// www.ipcc.ch/report/ar5/wg1/ [Accessed July 22, 2021]

[8] Wang XP, Sun DC, Yao TD. Climate change and global cycling of persistent organic pollutants: A critical review. Science China Earth Sciences. 2016;**59**(10):1899-1911. DOI: 10.1007/ s11430-016-5073-0

[9] Stockholm Convention. Available from: http://chm.pops.int/ Implementation/ProgrammeofWork/ tabid/6247/Default.aspx [Accessed July 23, 2021]

[10] Streit B. Fachbereich Biologie, University of Frankfurt, Siesmayerstrasse 70, D-6000 Frankfurt Am Main (Federal Republic of Germany). Experientia. 1992;**48**:955-970

[11] Fernández P, Grimalt JO. On the global distribution of persistent organic pollutants. Chimia (Aarau). 2003; **57**(9):514-521. DOI: 10.2533/ 000942903777679000

[12] Kallenborn R, Hung H, Brorström-Lundén E. Atmospheric Long-Range Transport of Persistent Organic Pollutants (POPs) into Polar Regions. Comprehensive Analytical Chemistry. Vol. 67. Elsevier; 2015. DOI: 10.1016/ B978-0-444-63299-9.00013-2

[13] Minh NH, Minh TB, Kajiwara N, Kunisue T, Subramanian A, Iwata H, et al. Contamination by persistent organic pollutants in dumping sites of Asian developing countries: Implication of emerging pollution sources. Archives of Environmental Contamination and Toxicology. 2006;**50**(4):474-481. DOI: 10.1007/s00244-005-1087-3

[14] Agusa T, Kuntio T, Nakashima E, Minh TB, Tanabe S, Subramanian A, et al. Preliminary studies on trace element contamination in dumping sites of municipal wastes in India and Vietnam. Journal de Physique IV. 2003;**107**(1):21-24

[15] Jacob J. A review of the accumulation and distribution of persistent organic pollutants in the environment. International Journal of Bioscience, Biochemistry and Bioinformatics. 2013;**3**(6):657-661. DOI: 10.7763/ijbbb.2013.v3.297

*Role and Effect of Persistent Organic Pollutants to Our Environment and Wildlife DOI: http://dx.doi.org/10.5772/intechopen.101617*

[16] González-Mille DJ, Ilizaliturri-Hernández CA, Espinosa-Reyes G, Cruz-Santiago O, Cuevas-Díaz MDC, Martín Del Campo CC, et al. DNA damage in different wildlife species exposed to persistent organic pollutants (POPs) from the delta of the Coatzacoalcos River, Mexico. Ecotoxicology and Environmental Safety. 2018;**2019**(180):403-411. DOI: 10.1016/j.ecoenv.2019.05.030

[17] Jin L, Gaus C, Escher BI. Bioanalytical Approaches to Understanding Toxicological Implications of Mixtures of Persistent Organic Pollutants in Marine Wildlife. Comprehensive Analytical Chemistry. Vol. 67. Elsevier; 2015. DOI: 10.1016/ B978-0-444-63299-9.00002-8

[18] Cunningham, andrew A. A walk on the wild side—Emerging wildlife diseases. BMJ. 2005;**331**(7527):1214- 1215. DOI: 10.1136/bmj.331.7527.1214

[19] Bayen S, Gong Y, Chin HS, Lee HK, Leong YE, Obbard JP. Androgenic and Estrogenic response of green mussel extracts from Singapore's coastal environment using a human cell-based bioassay. Environmental Health Perspectives. 2004;**112**(15):1467-1471. DOI: 10.1289/ehp.6990

[20] Asawasinsopon R, Prapamontol T, Prakobvitayakit O, Vaneesorn Y, Mangklabruks A, Hock B. The association between organochlorine and thyroid hormone levels in cord serum: A study from Northern Thailand. Environment International. 2006;**32**(4): 554-559. DOI: 10.1016/j.envint.2006. 01.001

[21] Vanden Berghe M, Weijs L, Habran S, Das K, Bugli C, Pillet S, et al. Effects of polychlorobiphenyls, polybromodiphenylethers, organochlorine pesticides and their metabolites on Vitamin A status in lactating grey seals. Environmental

Research. 2013;**120**:18-26. DOI: 10.1016/ j.envres.2012.09.004

[22] Ostertag S. Bridging Knowledge Systems to Monitor Beluga Whale Health & Habitat Use in the Beaufort Sea. In: North. Contam. Progr. 21st Annu. Results Work. Canadian Science Publishing, Arctic Science; 2015. p. 2015

[23] Van Dolah FM, Neely MG, McGeorge LE, Balmer BC, Ylitalo GM, Zolman ES, et al. Seasonal variation in the skin transcriptome of common Bottlenose Dolphins (Tursiops Truncatus) from the Northern Gulf of Mexico. PLoS One. 2015;**10**(6):1-21. DOI: 10.1371/journal.pone.0130934

[24] Montie EW, Fair PA, Bossart GD, Mitchum GB, Houde M, Muir DCG, et al. Cytochrome P4501A1 expression, polychlorinated biphenyls and hydroxylated metabolites, and adipocyte size of Bottlenose Dolphins from the Southeast United States. Aquatic Toxicology. 2008;**86**(3):397- 412. DOI: 10.1016/j.aquatox. 2007.12.004

[25] Tornero V, Borrell A, Aguilar A, Forcada J, Lockyer C. Organochlorine contaminant and retinoid levels in Blubber of common Dolphins (Delphinus Delphis) off Northwestern Spain. Environmental Pollution. 2006;**140**(2):312-321. DOI: 10.1016/j. envpol.2005.07.006

[26] Sonne C, Dietz R, Born EW, Riget FF, Kirkegaard M, Hyldstrup L, et al. Is bone mineral composition disrupted by organochlorines in East Greenland Polar Bears (Ursus Maritimus)? Environmental Health Perspectives. 2004;**112**(17):1711-1716. DOI: 10.1289/ehp.7293

[27] Komoroske LM, Lewison RL, Seminoff JA, Deheyn DD, Dutton PH. Pollutants and the Health of Green Sea Turtles resident to an urbanized Estuary in San Diego, CA. Chemosphere. 2011;**84**(5):544-552. DOI: 10.1016/j. chemosphere.2011.04.023

[28] Ashraf MA. Persistent organic pollutants (POPs): A global issue, a global challenge. Environmental Science and Pollution Research. 2017;**24**(5):4223-4227. DOI: 10.1007/ s11356-015-5225-9

[29] de Solla SR. Exposure, bioaccumulation, metabolism and monitoring of persistent organic pollutants in terrestrial wildlife. The Handbook of Environmental Chemistry. 2016;**49**:203-252. DOI: 10.1007/698\_ 2015\_450

[30] Li QQ, Loganath A, Chong YS, Tan J, Obbard JP. Persistent organic pollutants and adverse health effects in humans. The Journal of Toxicology and Environmental Health, Part A, Current Issues. 2006;**69**(21):1987-2005. DOI: 10.1080/15287390600751447

[31] Damstra T. Potential effects of certain persistent organic pollutants and endocrine disrupting chemicals on the health of children. Journal of Toxicology: Clinical Toxicology. 2002;**40**(4):457-465. DOI: 10.1081/ CLT-120006748

[32] Jørgensen N, Andersen AG, Eustache F, Irvine DS, Suominen J, Petersen JH, et al. Regional Differences in Semen Quality in Europe. Human Reproduction. 2001;**16**(5):1012-1019. DOI: 10.1093/humrep/16.5.1012

[33] Lee DH, Lee IK, Porta M, Steffes M, Jacobs DR. Relationship between serum concentrations of persistent organic pollutants and the prevalence of metabolic syndrome among nondiabetic adults: Results from the National Health and Nutrition Examination Survey 1999-2002. Diabetologia. 2007;**50**(9):1841-1851. DOI: 10.1007/s00125-007-0755-4

[34] Lee DH, Steffes MW, Sjödin A, Jones RS, Needham LL, Jacobs DR. Low dose of some persistent organic pollutants predicts type 2 diabetes: A nested case-control study. Environmental Health Perspectives. 2010;**118**(9):1235-1242. DOI: 10.1289/ ehp.0901480

[35] Wang SL, Tsai PC, Yang CY, Guo YL. Increased risk of diabetes and polychlorinated biphenyls and dioxins A 24-year follow-up study of the Yucheng Cohort. Diabetes Care. 2008;**31**(8):1574- 1579. DOI: 10.2337/dc07-2449

#### **Chapter 3**

## Impact of PCBs, Furan and Dioxin on Hepatocarcinogenesis

*Mohamed Helal, Sara Ghanem and Amany El-Sikaily*

#### **Abstract**

Stockholm Convention defined polychlorinated biphenyls (PCBs) as a group of persistent organic pollutants (POPs) such as dioxin/furan, dichlorodiphenyltrichloroethane, polychlorinated biphenyls, polychlorinated dibenzo-p-dioxins, aldrin, polychlorinated dibenzofurans and organometallic compounds (such as organotin and organomercury) which share the same characteristics of being persistent, bioaccumulative and toxic and can travel long distance through various media. They have diverse health impacts with different underlined molecular mechanisms. Recently, PCBs were referred as potent carcinogens with persistent existence in the environment. As the liver is the organ of detoxification, it is the major target organ for toxic effects induced by environmental contaminants, including PCBs. PCBs, furan and dioxin exert their hepatocarcinogenic effect through different mechanisms such as induction of oxidative stress, an increase of reactive oxygen species (ROS), mutagenic induction to oncogenes and epigenetic alteration to hepatic cells. In this chapter, we will provide an updated overview about PCBs, furan and dioxins, their impact on liver cancer initiation and progression on various *in vivo* and *in vitro* systems and its underlined molecular mechanisms. Also, a special emphasis will be directed to highlight zebrafish as *in vivo* model system to analyse the hepatocarcinogenic effect of these pollutants.

**Keywords:** polychlorinated biphenyls, Dioxin, Furan, hepatocarcinogenesis, zebrafish

#### **1. Introduction**

Polychlorinated biphenyls (PCBs) are related to a larger group of environmental pollutants named persistent organic pollutants (POPs) with intense ecological burden and toxicological problems [1, 2]. PCBs are halogenated aromatic hydrocarbons with special chemical formulas encompassing 209 congeners [3]. Based on the number of chlorine atoms and their location at the biphenyl, PCBs can be divided into lower and higher chlorinated congeners. Lower and higher chlorinated PCBs have different bioaccumulation rates [4]. PCBs are made of two cis-carbon rings linked by a single carbon bond and biphenyl molecule. Each PCB molecule consists of 12 carbon atoms alongside chlorine atoms substituted for hydrogen ones at any of 10 possible positions. Hence, theoretically, 209 possible PCB congeners can be found [5]. Due to their chemical characteristics, they have been used extensively in the industry as electric insulators, plasticizers, heat exchange and hydraulic fluids. The main problem of PCBs lies in their persistent nature, high degree of lipophilicity, slow transformation rate and low environmental degradation, which made them associated with a broad spectrum of human diseases such as reproductive,

immunological and carcinogenic [3, 6–8]. In 1987, International Agency for Research on Cancer and in 1996, the US environmental Protection Agency reported that PCBs are carcinogen in laboratory and wild animals and a possible carcinogen in humans [9–15].

Absorption of PCBs is feasible through ingestion, inhalation and dermal routes. The highest exposure pattern occurs through inhalation and skin absorption [16]. Once inside the body PCBs are transported to the liver. The main target of PCBs metabolism is the liver by the action of hepatic cytochrome p450 (CYP450). According to their chemical structure, PCBs can bind to different receptors [17] such as aryl hydrocarbon receptor (AhR), constitutive androstane receptor (CAR) and pregnane X receptor (PXR) [3]. Therefore, research was focused to study the impact of PCBs on liver function alteration and carcinogenicity [18]. Based on PCBs' chemical structure and receptor affinity, they can be categorized into dioxinlike and non-dioxin-like PCBs. Among different PCBs, Aroclors 1016, 1242, 1254 and 1260 are the most produced and used PCBs in the US during the 1958–1977 period [19].

One of the earliest incidences of human direct PCB ingestion was reported in Japan in 1968 and known as Yusho accident (oil syndrome in Japanese). This incident affected more than 1800 people who ingested rice oil contaminated with kanechlor-400 [20]. The average intake of PCBs in Yusho patients was 633 mg. The PCB concentration in adipose tissue was 46–76 ppm. After 34 years, the titre of non-ortho and mono-ortho PCBs in the blood of those patients reached 320 and 76 pg./g lipid, respectively [21]. Another famous PCB toxicity outbreak was reported a few years later after the Yusho incident in 1979 in Taiwan, known as the Yucheng incident (oil syndrome in Chinese). In this incident, around 2000 individuals consumed the same rice oil contaminated with Kanechlor-500 from one store [22]. Later, they developed skin problems and various health diseases. The concentration of PCBs in blood ranged from 3 to 1156 ppb [23]. A meta-analysis study of Yusho and Yucheng incidents showed that most patients exhibit a high degree of mortality due to lung, liver and skin cancers in both men and women [24, 25].

Before 1996, different assumptions have been made regarding PCBs' carcinogenicity. Sometimes all PCBs were considered carcinogens. Other assumptions indicated that mixtures with high chlorine content are only carcinogenic [26]. The carcinogenic potential of different PCBs is attributed to PCBs potency, which is affected by environmental processes (partitioning, chemical reaction, transformation and preferential bioaccumulation). Partitioning includes the fractionation of PCB mixture into different environmental compartments (air, water, sediment and soil). PCBs adsorption rate increases proportionally with their chlorine and organic content where low-chlorine content PCBs tend to be more volatile and hydrophilic and high-chlorine content is more persistent and lipophilic. Chemical transformation of PCBs in the environment occurs through the biodegradation process by the action of anaerobic bacteria in the sediments. These bacteria remove chlorines from *para* and *ortho* positions leading to a reduction of PCB toxicity [27]. PCBs bioaccumulation tends to concentrate in the adipose tissue with long-term stability and toxicity [28]. Due to the above-mentioned characteristics, EPA developed a new test approach to assess the cancer risk of different PCBs considering both environmental processes and PCBs toxicity [29]. Furthermore, Stecca *et al*. used different human cell lines (HuH6, HepG2 and DLD-1) to develop a suitable *in vitro* test battery to evaluate the cumulative effect of different PCBs mixtures. Their results showed that the best *in vitro* model to study toxicity is HepG2 and DLD-1 cell lines in terms of expression panel of several genes such as *AhR*, *AR*, *PXR*, *PPARγ*, *ERβ* and *THRα* (which showed the best representative expression upon PCBs mixture treatment) [30]. Most importantly, many of the genes identified by the Stecca group are also

implicated in lipogenesis and lipid homeostasis in the liver indicating a significant role of PCBs on lipid metabolism in the liver [31].

Based on several reports from different agencies, PCBs were referred as possible human carcinogens [32]. This assumption is based on experimental data from rodents where PCBs treatment increased neoplasm formation in different rodent tissues.

Commercial PCB mixtures such as Aroclors 1016, 1242, 1254 and 1260 (dietary PCBs concentration ranged from 25 to 200 ppm over a period of 24-month treatment) were found to induce not only alteration in liver function tests (AST, ALT and GGT) but also liver tumours with bile duct carcinoma (cholangiocarcinoma) in rats (female tumour incidence higher than male rats) after long-term feeding regime [33]. In addition, rats exposed to commercial mixtures with 60% chlorine through a dietary lifetime regime developed benign liver tumours that eventually progressed to malignant ones [34]. Furthermore, a mixture of PCB126 and PCB153 caused a mild increase in neoplastic liver lesions in mice. This was accompanied by an up-regulation of Cyp1a1 and Cyp2b10 (RNA and protein level) [35]. The same promotion of liver carcinogenesis was observed in partially hepatectomized rats challenged with a single DEN dose and subsequent intraperitoneal injection of PCB105, 126 or 153 [36]. Moreover, rats receiving a single dose of DEN followed by intraperitoneal injection of PCB77 or PCB153 (150 μmol/kg) alongside seleniumenriched diets feeding developed hepatic neoplasm. Selenium administration enhances the carcinogenic induction of PCB77 more than that of PCB153 as the number of positive placental glutathione s0transferase (PGST+) hepatic regions was higher in the former than the latter, respectively [37].

Single PCB compounds have a preferential binding affinity to different receptors. For example, PCB126 binds to AhR while PCB153 binds to CAR. Rats treated with a single dose of DEN followed by PCB126, PCB153 or in combination developed hepatic neoplasms most profoundly in single PCBs treatment only (PCBs combination treatments showed antagonistic results on liver neoplasm formation) as indicated with positive GST-P liver areas [38]. Cultured mouse hepatocytes treated with PCB126 exhibited reduced hepatocyte glycogen content in a dosedependent manner and suppressed forskolin-stimulated gluconeogenesis from lactate. Interestingly, glycogen treatment of cells restored PCB127 effects, indicating that PCB127 could affect the terminal players in the gluconeogenesis cycle. Finally, PCB126 could activate AhR and its downstream effector phosphoenolpyruvate carboxylase. This suggests a possible role of PCBs as energy metabolism disruptor agents [39]. Other studies showed that PCB153 could induce hepatocarcinoma through induction of NF-kB in mice (this was inhibited by deleting the p50 subunit of NF-kB) [40] or induce mutation in β-catenin (*Catnb*) [41] or *ras* [42] oncogenes as a promoter of tumorigenesis.

PCBs administration could interfere with metals accumulation in the liver and affect their transport and excretion through kidneys. Mice fed different concentrations of PCBs with Cadmium (Cd)-enriched diets showed a reduction of Cd concentration in the liver. Also, liver histology of those mice revealed a characterized centrilobular enlargement of hepatocytes, hepatic focal necrosis and clear cytological signs of malignancy than the control group [43]. On the other hand, female rats fed a diet enriched with high-dose Aroclor-1254 and Aroclor-1260 for 78 weeks developed initial iron accumulation in the liver by week 52, induced hepatocyte proliferation and eventually liver carcinoma by the 78th week, indicating that iron accumulation in the liver is an early sign of hepatic neoplasm transformation induced by PCBs [44].

Human exposure cohort studies were also conducted to monitor the pathological aspects of PCBs. Workers in capacitor factories exposed to Aroclors mixtures

with 41–54% chlorine content had increased mortality rates from liver tumours (gall bladder and biliary tract) [45]. The same finding was reported in HCC Italian patients settled in areas highly polluted with PCBs [46, 47]. The burden of PCBs concentration in liver, lung and kidney tissues of Chinese cancer patients residing near e-waste disassembly sites was very high (257.9 to 455.1 ng g−1), indicating a possible correlation between PCBs exposure and cancer incidence in those patients [48]. Another long-term cohort study was conducted in Germany in former PCBexposed workers. The study linked the change in liver enzymes and morphology with PCB exposure level. There was a significant inverse connection between PCB concentration and ɤGT and a significant association between liver enlargement and PCBs level [49]. Another cohort study in the USA linked elevated levels of orthosubstituted PCBs and liver toxicant-associated steatohepatitis (TASH) in the former worker of PCB manufacturing complex. The authors reported that the increase in PCBs exposure was connected with an increase in liver disease burden, inflammation, steatohepatitis induction and hepatocyte apoptosis and fibrosis [50]. A large and extensive cohort study was conducted on 138,905 electricity workers exposed to insulating liquids of PCBs at five different electricity companies between the period of 1950 and 1986. Poisson regression was utilized to examine mortality of skin cancer (melanoma) and liver cancer in relation to PCBs exposure. Results showed that PCBs exposure was linked to melanoma development and in some workers hepatic cancers [51]. A controlled study was nested with two large prospective cohorts (one from Northern California Multiphasic Health Check-up (MHC) comprising 408 HCC cases and Norwegian Janus group comprising 84 HCC cases) from 1960 to 1980. Measuring 37 different congeners with GC-MS, the authors found that among measured congeners, PCBs (151, 170, 172,180,177 and 195) congeners were the highest with a concentration in HCC patients 1.3 to 1.4 ng/g lipid for the first group and 1.9 ng/g lipid for the second group, confirming a significant link between PCBs levels and HCC development [52].

PCBs can be indirectly accumulated in the human body through food chain by ingestion of aquatic animals contaminated with PCBs. For example, Delistraty study showed that PCBs titre in different aquatic animals in Columbia River, USA was significantly high. Sturgeon liver, whitefish fillet, carbs and smallmouth bass all showed significant high level of dioxin-like PCBs, non-dioxin like PCBs and total PCBs [53]. Another study showed that Bottlenose dolphin was stranded alive with high levels of different PCBs such as PCB 153,180, 187 and 138. Finally, large cell immunoblastic lymphoma was observed in the hepatic sinuses of these dolphins accompanied with liver enlargement. All previous studies indicate a direct correlation between carcinogenesis induction and levels of PCBs in those dolphins [9].

#### **1.1 Possible hepatocarcinogenic mechanisms of PCBs**

PCBs mode of action and the underlined molecular mechanisms of toxicity and carcinogenicity have not been deciphered so far [7]. Yet, studies on different animal models, *in vitro* cell lines and human cohort studies could give us a glimpse of the key molecular players responsible for different pathogenic outcomes.

2,3′4,4′,5-Pentachlorobiphenyl known as PCB118, one of the most persistent congener members, was found to promote hepatocellular carcinoma SMMC-7721 cell proliferation and glycolysis through AhR, which subsequently elevates the expression of pyruvate kinase M2 (PKM2) and stimulation of reactive oxygen species (ROS) production through nicotinamide adenine dinucleotide phosphate (NADPH). These effects were inhibited by treating cells with PKM2 shRNA and superoxide dismutase, respectively [6].

PCB126 (3,3′,4,4′,5-Pentachlorobiphenyl), a non-ortho-chlorinated congener, was found to increase the synthesis of ROS specifically. Treatment of HepG2 cells with this congener enhance their carcinogenicity by inducing an oxidative stress response that was underlined by activation of mitogen-activated protein kinases (extracellular signal-regulated kinase 1/2), p38, c-Jun phosphorylation, activating protein-1 (AP-1) and finally an expression of antioxidant-responsive element (ARE)-dependent genes [7]. In addition, Faust *et al*. demonstrated that rat progenitor liver cells (WB-F344) treated with PCB126 exhibit a differential transcriptional response over the treatment period. At 6-hour post-treatment (hpt), about 146 significant deregulated genes were identified under AhR direct targets. The number of deregulated genes was 658 and 968 after 24 and 72 hpt, respectively. The most identified genes through gene ontology analysis were affiliated to developmental, cell cycle, growth control and drug metabolism. The main targeted pathways were Wnt and TGF-β. Finally, they have also identified a novel target gene under the AhR signalling pathway such as *Fst*, *Btg2*, *Ctgf* and *Hbegf* [54]. AhR is an essential receptor controlling liver response to environmental toxicants. By using rat and human hepatocytes as *in vitro* cellular models to study PCBs toxicity, researchers found that rats fed Arochlor 1254 exhibit liver carcinoma through activation of AhR and downstream induction of *raf* effector in a MAPK-dependent pathway [55].

PCB47, 49, 52, 77 and 153 have a tumour promoting activity [32]. Also, some PCBs induce liver toxicity through induction of mixed function oxidases (phenobarbital, 3-methylcholanthrene) [56] and inhibition of anti-oxidant production such as PCB154, 155, 184 and 153 inhibit paraoxonase 1 (PON1) in treated rats [57].

Most PCBs mixtures with high chlorine content and their derived metabolites showed superior tumour-promoting characteristics. Yet, concern over low-chlorine content PCBs was raised after experiments showing that dihydroxy metabolites induce breast cancer by inducing oxidative DNA damage in breast cancer cells [58]. Another possible mechanism of PCBs carcinogenicity is their ability to suppress the immune system and cause endocrine disruption [59]. PCB104, 188 and their hydroxylated forms 4′-50, 4′-30, 4′-72, 4′-112 and 4′-121 disturb endocrine pathways in rainbow trout cultured hepatocytes and induced vitellogenin synthesis indicating altered liver physiology [60]. Human MCF-7 cells exposed to PCBs analogues showed a reduction in catechol-O-methyltransferase (COMT) activity on the transcriptional and translational level *via* the oestrogenic receptor. This could explain the PCB mode of liver tumour induction *via* modulation of oestrogen receptor response [61]. A comparative metabolomic study was conducted on rats fed a control diet and choline-deficient diet (as an inducer of liver non-alcoholic steatohepatitis) and subsequently exposed to PCB126. The addition of PCB127 promoted fatty liver development through dysregulation of glycerophospholipid metabolism, CoA biosynthesis pathway and glutathione metabolism. In addition to lipid metabolism disturbance, PCB127 down-regulated redox genes, and induced oxidative stress and mitochondrial dysfunction [62].

HepG2 cells co-treated with benzo-a-pyrene and different doses of Aroclor 1254 had a high degree of DNA damage (as indicated by DNA migration assay and formation of 8-hydroxy-2′-deoxyguanosine (8-OHdG)), oxidative stress and elevated CYP1A activity [63]. In another experiment, HepG2 cells exposed to various PCBs concentration (0.01-10µM) exhibit aggressive carcinogenic behavious underlined by pERK Tyr204 and pMdm2 Ser166 which attenuated P53 activity in those cells [64].

Dioxin-like PCBs such as PCB 77 and 81 were shown to have direct genotoxic effects on Chinese hamster V79-derived cell line by inducing micronuclei formation, and induced expression of CYP1A1, CYP2E1 and γ-H2AX protein (a marker of DNA double-strand breaks) [65, 66].

Another surprising finding of PCBs-induced hepatic carcinogenicity is their ability to inhibit intercellular community between liver cells. Mouse hepatoma cell line (Hepa1c1c7) treated with TCDD and different PCBs showed a rapid intercellular inhibition after 2 hrs. of treatment accompanied with AhR activation and induction of ethoxyresorufin O-deethylase (EROD) activity (an early marker of PCBs induced oxidative stress) [67]. Moreover, by using a quantitative polymerase chain reaction (qPCR) to quantify relative telomere length in lung and liver samples collected from rats treated with different PCBs (126, 153 and a mixture of them) showed larger relative telomere length, which is an early indication of euplastic or non-neoplastic pathogenic disease development [68].

#### **2. Furan**

Furan, a heterocyclic organic chemical, is considered as a human carcinogen and a liver toxicant in rodents [69]. It is found in a broad spectrum of common heat-treated and jarred foods in addition to tobacco smoke. It is also generated from numerous precursors such as amino acids, ascorbic acid and carbohydrates [70]. Infants received the highest furan exposure from ready-to-eat meals, while adults are exposed to furan by the dietary intake of coffee [69]. Furan is found mainly in the liver and is metabolized to the reactive metabolite, cis-but-2-ene-1,4-dialdehyde (BDA) through cytochrome P450 2E1 (CYP2E1). Reported studies have indicated that humans can convert furan to its reactive metabolite and cis-2-butene-1,4-dial (BDA), and consequently may be subjected to furan toxicity [71].

Being hepatotoxic, researchers [72] stated that furan is associated with cholangiofibrosis in rats and HCC & adenomas in mice. They also indicated that oxidative stress, alterations in gene expression, epigenetic modifications, inflammation and increased cell proliferation represent indirect mechanisms that are included in carcinogenesis. The carcinogenic effects of furan have been referred to as genotoxic and non-genotoxic modes of action. Epigenetic alterations are among the most important non-genotoxic alterations induced by furan since they are related to all other non-genotoxic events [69]. As a genotoxic furan could be linked to furan-into carcinogenicity, current human exposure levels to this hepatotoxicant may represent a risk to human health and required the necessity for its mitigation [73].

#### **2.1 Possible hepatocarcinogenic mechanisms of furan**

Metabolism of furan leads to the formation of protein adducts in the target organ. The first bioactivation step comprises the oxidation of cytochrome P450 catalysed of furan, which generates cis-2-butene-1,4-dial (BDA). BDA can react with lysine to form pyrrolin-2-one adducts [70]. This metabolite directly reacts with DNA nucleophiles and proteins [74]. It is also known as a bacterial mutagen in Ames assay strain TA104. According to metabolic studies, this reactive metabolite is formed *in vivo* [74]. BDA was found to react with glutathione (GSH) generating 2-(S-glutathionyl) butanedial (GSH-BDA), which reacts in turn with lysine forming GSH-BDA-lysine cross-links. Relative reactivity of these two intermediates was explored by the reaction of cytochrome *c* with BDA in the existence and absence of GSH [75]. Using MALDI-TOF mass spectrometry, BDA was found to react widely with cytochrome *c* forming adducts (which add 66 Da to the protein) according to pyrrolinone adducts formation. On the other hand, when GSH was added to the reaction, the overall extent of adduct formation was reduced. Briefly, the majority of adducts arose on lysine residues contributing to the carcinogenic hazard of furan [71]. By using liquid chromatography, tryptic peptides analysis clarified a cross-link between GSH-BDA and lysine 107 of histone H2B isolated from male F344 rats' liver exposed to carcinogenic doses of furan. This cross-link was detected before the identification of epigenetic changes and occurred at a lysine residue that is known as a target for epigenetic modifications and crucial for nucleosome stability [76].

Being a hepatocarcinogen in mice and rats, furan induced an enhancement for cytotoxic pathways represented by signalling of stress-activated protein kinase (SAPK) and death receptor (DR5 and TNF-alpha), and proliferation through extracellular signal-regulated kinases (ERKs) and TNF-α. In addition, NF-kappa B and c-Jun (genes essential for liver regeneration) were involved in response to furan [76]. Previous studies applying furan high doses revealed that it induced tumours at nearly 100% incidence at all doses [77]. Fraction of H-ras codon 61 CAA to AAA mutation was increased in liver tumours of furan-treated mice [78]. Besides, furan has a deleterious impact on the activity of crucial target enzymes included in ATP synthesis, glycolysis, redox regulation as well as b-oxidation in rat liver. After treatment with a high dose of furan, it was found that glyceraldehyde-3-phosphate dehydrogenase was significantly inhibited and observed some metabolic changes reliable to blockage of the glycolytic breakdown of glucose in the liver of the rat. Despite an increase in enoyl-CoA hydratase activity, an enhancement of ketone bodies production and a reduction in the activity of succinate dehydrogenase were recorded as a result of furan treatment. These enzymatic changes were linked to impairments occurring in cellular processes affecting the metabolic pathways and antioxidant defence and indicate mitochondrial dysfunction as a serious incident in furan toxicity [79]. Moreover, targets of putative protein of furan reactive metabolites induced functional damage of numerous individual proteins and interference with pathways, especially that of mitochondrial energy production, redox regulation, and protein folding. This damage represented critical targets of furan toxicity and can combine together to disturb cell homeostasis and cause the cell death of hepatocytes [80].

The liver is the main target organ affected by furan as indicated by serum biomarkers changes, change in liver weights and histological lesions after exposure to furan. Accordingly, a dose of 0.03 mg/kg bw of furan was proposed to be the non-detectable serious effect for hepatic toxicity [80]. In addition, Selmanoğlu *et al.* [81] revealed a significant increase of LDL levels, a significant decrease in ALT and ALP levels and insignificant changes in liver MDA levels, catalase activities and superoxide dismutase in the liver of rat groups treated with furan comparing with control groups. They also indicated a significant change in liver weights of furan-treated groups and observed hyperaemic blood vessels in their hepatic tissue under the light microscope. Histopathologically, multifocal hepatocellular necrosis intermingled with pigment-laden Kupffer cells and reactive leukocytes, oval-cell hyperplasia enhancement, hepatocyte mitoses increase and hepatocyte injury were also observed in livers from furan-treated mice as a result of furan induction [82]. Furthermore, furan-enhanced Ki-67 and PCNA expression in hepatic tissues increased the content of ROS in addition to indices of oxidative damage and decreased the TAC in the serum level of exposed rats. Finally, exposure to furan was found to be linked to changes in the mRNA expression pattern of intermediate filament proteins in hepatic tissues and promoted fibrosis and proliferation of hepatocytes in the liver [82].

In addition, analysis of liver rats treated with furan by Comet assays showed breaks in both strands of DNA, an increase in oxidized purines and pyrimidines at cancer bioassay dosage represented by a near-linear dose-responsive manner [83]. Consequently, these findings postulated that furan induces cancer mainly in rats' liver through a secondary genotoxic mechanism including oxidative stress, a down-regulation in the expression of apoptotic, cell-cycle checkpoints as well as DNA-repair genes accompanied by inflammation and cell proliferation dosage [83].

Furthermore, glutathione S-transferase placental form-positive (GST-P) foci are considered as preneoplastic lesions markers in the hepatocarcinogenic rats. Using reporter gene transgenic rats, it was found that furan rapidly induces GST-Pþ foci formation without reporter gene mutation after short exposure [84]. On the other hand, GST-P foci development is probably due to cell proliferation other than the genotoxic mode of action in furan-treated rats. Based on the close association between neoplastic hepatocytes and GST-P, Hibi *et al*. [85] postulated that cell proliferation following signal transduction other than the pathway of mitogenactivated protein kinase (MAPK)/ERK may contribute in the early stage of furaninduced hepatocarcinogenecity.

Cholangiofibrosis is defined as a physical anomaly that occurs before cholangiocarcinoma development in some rodents. Some reports explained that severely affected areas of the liver representing injury due to furan administration were extended into the portal and capsular parts, resulting in a rapid ductular cells proliferation that extended into the parenchyma accompanied by a subtype of liver fibroblasts. These ductules were differentiated into hepatocytes lacking fibroblasts or developed to form tortuous ductular structures replacing much of the parenchyma, leading to cholangiofibrosis [86]. Moreover, furan-induced cholangiocarcinomas were proposed to develop from cholangiofibrosis areas as a consequence of indirect and chronic damage to DNA through oxygen radicals joined with persistent proliferative signals, including loss of connexin 32, which acts to translate this DNA damage to fixed mutations [87].

#### *2.1.1 Epigenetic alterations and the non-genotoxic mechanism of furan in liver*

The carcinogenic effect of furan has been referred to as a genotoxic and nongenotoxic mechanism comprising epigenetic alterations in liver tissue [88]. Some reports postulated that furan carcinogenicity is caused by a non-genotoxic mechanism since it was not genotoxic in *in vivo* or *in vitro* micronucleus assay [89, 90]. Other studies indicated that BDA is not directly responsible for the effects of furan on mutational spectra *in vivo.* Therefore, an indirect mechanism of genotoxicity was hypothesized in which chronic toxicity was followed by inflammation and secondary cell proliferation that triggers the development of cancer in furan-exposed models [82].

In addition, epigenetic alterations involving DNA and microRNA (miRNA) methylation play a fundamental role in inducing furan carcinogenicity. It was indicated that DNA methylation changes and miRNA modulation followed by a DNA-damage response are the most pronounced alterations resulted from the use of 3-month furan treatment at a carcinogenic dose suggesting that non-genotoxic mechanisms are crucial for furan carcinogenicity [91]. It was found that gene-specific DNA methylation alterations have an essential role in the contribution of furan hepatotoxicity and hepatocarcinogenicity [88]. Other studies indicated that aberrations in microRNAs (miRNAs) expression are one of the non-genotoxic alterations induced by furan exposure, which highlighted the role of epigenetic impairments in the furan hepatotoxicity mechanism [69].

Moreover, Conti *et al*. [92] mentioned that epigenetic modifications which occurred in hepatotoxicity and carcinogenicity of furan are dose and timedependent. They noted some epigenetic aberrations represented by DNA methylation, histone lysine acetylation and methylation, gene-specific methylation and alteration of chromatin-modifying genes expression in male Fisher rats treated with furan. Their findings indicated that sustained alterations in histone lysine acetylation (which is responsible for the ability of cells to maintain and control correctly the expression of genetic information) represent the adverse effects of furan induction. Some reports indicated that gene expression alterations resulted from furan exposure were irreversible [91]. Using whole-genome transcriptomic analysis, Tryndyak *et al*. [88] demonstrated differential gene expression alterations in liver lesions induced in male rats treated with furan. These alterations are essential in key pathways linked with the diverse aspects of liver pathology. Furthermore, it was noted that the continuous exposure to furan induced noticeable changes in the expression of miRNA represented by over-expression of hepatic miRNAs (miR-34a, miR-93, miR-200a, miR-200b and miR-224), and down-regulation of miR-375. In addition, hypermethylation of cytosine DNA and the lysine methylation of histone H3K9 and H3K27 at the MiR-375 genes were increased due to the reduction in miR-375 expression. It was revealed that the significant miR-375 inhibition was accompanied by toxicity and carcinogenicity of furan-induced liver leading to an up-regulation in Yes-associated protein 1 (YAP1), which is one of the principal events in liver carcinogenesis [69].

Since the mammalian genome is transcribed into mRNAs that code for protein and other non-coding RNA products [93]. Long non-coding-RNAs (lncRNA) are known as ncRNA species >200 nucleotides long, which represent significant epigenetic regulators of gene expression and are included in a wide spectrum of biological processes related to toxicology. Recio *et al*. [93] indicated that lncRNAs are transcriptional targets in the cytotoxic levels of furan exposure inducing cell proliferation. They also hypothesized that lncRNAs are considered as epigenetic biomarkers of carcinogenic exposure.

#### **3. Dioxins**

Dioxins are considered as representative toxic agents among persistent organic pollutants and a large family of halogenated aromatic hydrocarbons, which composed of tricyclic aromatic compounds [94]. These compounds are produced by industrial wastes and can accumulate in soil, sediments as well as food chains with long half-life of numerous years, affecting human health [95]. 2,3,7,8-Tetrachlorodibenzo -p-dioxin (TCDD) is a typical representative and the most toxic substance of dioxins, which exhibits systemic hepatotoxicity, carcinogenicity, immunotoxicity, teratogenicity, endocrine disruption and also affects pathology and physiology of human skin [96]. Being with four chlorine atoms in lateral positions, 2,3,7,8 Tetrachlorodibenzop-dioxin (TCDD) is the most biologically active isomer of dioxins [97]. It is a widespread and persistent pollutant in the environment originated from waste incineration or metal industries, plastics manufacturing and paper processing [98]. Moreover, it plays a significant role by binding to AhR for endocrine changes in experimental animals [99]. Besides, Türkez, Türkez *et al.* [100] suggested that oxidative stress has a crucial role for toxic effects of TCDD with AhR.

#### **3.1 TCDD exposure and dioxin receptor**

The aryl hydrocarbon receptor (AhR) is considered as a ligand-activated receptor which enables environmental pollutant toxicity like 2,3,7,8-tetrachlorodibenzop-dioxin (TCDD) [101]. It is also known as xenobiotic receptor or dioxin receptor and is a member of the basic helix-loop-helix/period AhR nuclear translocator single-minded family [102]. AhR translocates to the nucleus after binding to TCDD, dimerizes with AhR nuclear-translocator protein, binds to dioxin-responsive elements and up-regulates a series of genes expression that encode xenobiotic-metabolizing enzymes, such as cytochrome P450s (e.g. CYP1A1, CYP1A2), NAD(P)H

quinone oxidoreductase as well as a form of UDPglucoronosyl-transferase-6 [103]. Even though AhR may serve as part of an adaptive chemical response, numerous studies reported that this dioxin receptor has important functions in liver, cell proliferation, cardiac development [104, 105] and the ubiquitin-proteasome system [106]. AhR plays a fundamental role in three biological aspects including xenobiotics metabolism, the toxic responses related to TCDD (dioxin) exposure and the vascular remodelling of the developing embryo. Using *Cre*-lox technology, Walisser *et al*. [107] examined the role of AhR signalling in hepatocytes and endothelial cells. They revealed that AhR signalling in hepatocytes is crucial to produce adaptive and toxic hepatic responses due to TCDD exposure.

Being a generally expressed ligand-dependent transcription factor, AhR mediates cellular responses to dioxins. Boutros *et al*. [108] demonstrated that AhR mediated all effects of dioxin on hepatic mRNA levels and revealed the alteration of 297 genes including many well-established AhR target genes due to dioxin exposure in mice liver. They also indicated that AhR genotype remodelled hepatic transcriptomes suggesting the existence of a basal AhR gene battery. Results of Boutros *et al*. [108] highlighted the fundamental role of this dioxin receptor in the liver tissue.

In response to dioxin*,* Kennedy *et al*. [109] also explained that the signal transduction pathways that mediate tumour promotion of liver by 2,3,7,8-tetrachlorodibenzo-*p*-dioxin are accomplished by the linked action of two receptor systems, the AhR and the receptors for the "IL-1-like" cytokines. However, Yamaguchi and Hankinson [110] indicated that TCDD might suppress the cell growth of liver cancer through numerous signalling pathways, mediated by AhR and its related co-factors. In addition, they found that the impact of TCDD was accomplished by gemcitabine (responsible for nuclear DNA damage in cancer cells), suggesting that their use as a combination may be considered as a suppressor of tumour cell growth *in vitro*.

TCDD can induce hepatic fibrosis through a sequential events of steatosis followed by steatohepatitis. Lee, Wada [102] investigated the role of AhR in liver steatosis in wild type and transgenic mice. They concluded that AhR activated in liver cells induced CD36 expression, enhanced the uptake of fatty acids and steatosis induciton [102].

Cytochrome P4502E1 (CYP2E1) mainly expressed in liver, is involved in the metabolic activation of carcinogens and hepatotoxins such as TCCD and CCl4. At post-transcriptional levels, CYP2E1 is induced and exerted mostly through mRNA and protein stabilization, while xenobiotic induction is found to be very limited at the transcriptional level [101]. Since the effect of xenobiotics on CYP2E1 liver, expression is of significant attention. Therefore, Mejia-Garcia *et al*. [101] studied the effect of TCDD on CYP2E1 liver of mouse and the impact on CCl4 that induced hepatotoxicity. They found that TCDD augmented levels of mRNA and protein in hepatic tissue of mouse CYP2E1 in an AhR-dependent manner and CYP2E1 was induced causing CCl4-induced hepatotoxicity.

Recent studies revealed that TCDD exposure had caused increased productions of lipid peroxidation, reactive oxygen species and histopathological injury in the liver of both rats and mice [111]. This exposure also enhances oxidative stress and diminishes the fluidity of hepatic membrane and glutathione (GSH) content, as well as imbalances the antioxidant enzymes in the liver [112, 113]. Moreover, an increase in the relative weight of the liver, a significant increase in all of the hepatic biomarker levels (glucose, cholesterol, triglycerides, AST, ALT and LDH) in the serum and a decrease of the antioxidant enzyme activities (catalase, glutathione peroxidase and superoxide dismutase) were observed under dioxin effect in hepatic tissue of rat [114]. Additionally, Bentli *et al.* [99] revealed that immunotoxicity associated with altered cytokine levels is among the other TCDD-induced toxicity prominent symptoms. Finally, Ciftci and Ozdemir [115] indicated that one of the

main regulated pathways of TCDD toxicity is the elevated levels of the inflammatory cytokine. Using the real-time polymerase chain reaction (PCR), several studies indicated that heat shock proteins (mortalin, α-B-crystallin, Hsp105, Hsp27 and Hsp90s) and antioxidant enzymes (GST, SOD-3 and catalase) in livers of rats were induced suggesting protective mechanisms against 2,3,7,8-TCDD which induced hepatotoxicity [116]. Moreover, Czepiel *et al*. [117] mentioned that TCDD impaired the liver of rats and the activity of CYP1A1 in a dose-dependent manner. Parenchymal degeneration of hepatic lobules, hepatocytes vacuolation in prominent and peripheralized nuclei, hepatocellular hypertrophy and turgor of the vein in the centriacinar regions were also observed in rats' liver that received a high dose of dioxin [114]. TCDD also induces CYP1A1 activity by elevating the immunohistochemical reactivity of central areas of hepatic lobules located around the central vein in the rat liver [117].

#### **4. Zebrafish models of PCBs-, furan- and dioxin-induced hepatocarcinogenesis**

Viluksela and Pohjanvirta [118] reported that paternal exposure to TCDD was considered as the most effective congenator of dioxins in laboratory rodents and zebrafish as it can lower the reproductive performance and reduce the male/female ratio of offspring. Therefore, it will affect subsequent generations *via* both paternal and maternal germlines. These adverse effects have been accompanied by epigenetic alterations in sperm cells and/or placenta, including variations in methylation patterns of imprinted genes.

Besides, previous studies have demonstrated that dioxins broadly alter hepatic mRNA levels [119]. Unexpectedly, Boverhof, Burgoon [120] found that responses of mouse and rats to TCDD exposure revealed that rat and mouse responses diverge significantly through analysis of a limited portion of their transcriptome. Accordingly, it was suggested that both mice and rat models should be applied to detect the acute hepatotoxicity of xenobiotics [120].

Under fabp10a promoter, Zhang [121] also established a line of transgenic zebrafish line (LiPan) characterized with the expression of liver-specific red fluorescent protein (DsRed), which enables the observation of liver in live LiPan fry. They revealed that TCDD could significantly increase both liver red fluorescence and size in LiPan fry. Thus, LiPan transgenic fry offers a suitable and rapid hepatotoxicity assay *in vivo* that should be used to monitor the effect of environmental contaminants from the chemical mixture [121].

Furthermore, by using the inducible *kras* transgenic zebrafish model of hepatocarcinogenesis, Qiqi *et al.* have shown that PCB12 and TCDD alongside other environmental pollutants could accelerate HCC induction and inflammation of the liver [122]. Zebrafish is a very promising model of environmental and molecular toxicology. Further studies and more deep studies should be carried out using this model to provide more insightful information about carcinogenic potential and mechanisms of PCBs, furan and dioxins.

#### **5. Conclusion**

Environmental pollutants are a severe persistent burden, which cause a broad spectrum of health problems not only to aquatic animals but also to humans. Among them, PCBs, furan and dioxin are organic pollutants that were widely used in different applications before they were banned due to their carcinogenic

potential. Different studies using different model animals and screening systems (*in vivo* and *in vitro*) indicated their correlation with liver tumour induction and promotion. In this chapter, we highlighted the collective and recent updates linking these groups of pollutants with the pathology of hepatocellular carcinoma. Although extensive research has been done, yet the exact potential and molecular mechanisms of these pollutants are to be discovered and deciphered.

### **Author details**

Mohamed Helal\*, Sara Ghanem and Amany El-Sikaily National Institute of Oceanography and Fisheries (NIOF), Cairo, Egypt

\*Address all correspondence to: m.helalf@gmail.com

© 2022 The Author(s). Licensee IntechOpen. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/ by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

*Impact of PCBs, Furan and Dioxin on Hepatocarcinogenesis DOI: http://dx.doi.org/10.5772/intechopen.101526*

#### **References**

[1] Dirinck E, Jorens PG, Covaci A, Geens T, Roosens L, Neels H, et al. Obesity and persistent organic pollutants: Possible obesogenic effect of organochlorine pesticides and polychlorinated biphenyls. Obesity (Silver Spring, Md). 2011;**19**(4):709-714

[2] Markowitz G, Rosner D. Monsanto, PCBs, and the creation of a "world-wide ecological problem". Journal of Public Health Policy. 2018;**39**(4):463-540

[3] Faroon O, Ruiz P. Polychlorinated biphenyls: New evidence from the last decade. Toxicology and Industrial Health. 2016;**32**(11):1825-1847

[4] Safe S. Toxicology, structurefunction relationship, and human and environmental health impacts of polychlorinated biphenyls: Progress and problems. Environmental health perspectives. 1993;**100**:259-268

[5] Carpenter DO. Polychlorinated biphenyls (PCBs): Routes of exposure and effects on human health. Reviews on Environmental Health. 2006;**21**(1):1-23

[6] Liang W, Zhang Y, Song L, Li Z. 2,3′4,4′,5-Pentachlorobiphenyl induces hepatocellular carcinoma cell proliferation through pyruvate kinase M2-dependent glycolysis. Toxicology Letters. 2019;**313**:108-119

[7] Song MO, Freedman JH. Activation of mitogen activated protein kinases by PCB126 (3,3′,4,4′,5-pentachlorobiphenyl) in HepG2 cells. Toxicological Sciences. 2005;**84**(2):308-318

[8] Knerr S, Schrenk D. Carcinogenicity of "non-dioxinlike" polychlorinated biphenyls. Critical Reviews in Toxicology. 2006;**36**(9):663-694

[9] Jaber JR, Pérez J, Carballo M, Arbelo M, Espinosa de los Monteros A, Herráez P, et al. Hepatosplenic large cell immunoblastic lymphoma in a bottlenose dolphin (Tursiops truncatus) with high levels of polychlorinated biphenyl congeners. Journal of Comparative Pathology. 2005;**132**(2-3):242-247

[10] Program NT. Toxicology and carcinogenesis studies of 2,3,4,7,8 pentachlorodibenzofuran (PeCDF) (Cas No. 57117-31-4) in female Harlan Sprague-Dawley rats (gavage studies). National Toxicology Program Technical Report Series. 2006;**525**:1-198

[11] Program NT. Toxicology and carcinogenesis studies of a binary mixture of 3,3′,4,4′,5 pentachlorobiphenyl (PCB 126) (Cas No. 57465-28-8) and 2,2′,4,4′,5,5′ hexachlorobiphenyl (PCB 153) (CAS No. 35065-27-1) in female Harlan Sprague-Dawley rats (gavage studies). National Toxicology Program Technical Report Series. 2006;(530):1-258

[12] Program NT. Toxicology and carcinogenesis studies of a mixture of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) (Cas No. 1746-01-6), 2,3,4,7,8-pentachlorodibenzofuran (PeCDF) (Cas No. 57117-31-4), and 3,3′,4,4′,5-pentachlorobiphenyl (PCB 126) (Cas No. 57465-28-8) in female Harlan Sprague-Dawley rats (gavage studies). National Toxicology Program Technical Report Series. 2006;(526):1-180

[13] Program NT. Toxicology and carcinogenesis studies of a binary mixture of 3,3′,4,4′,5 pentachlorobiphenyl (PCB 126) (Cas No. 57465-28-8) and 2,3′,4,4′,5 pentachlorobiphenyl (PCB 118) (Cas No. 31508-00-6) in female Harlan Sprague-Dawley rats (gavage studies). National Toxicology Program Technical Report Series. 2006;**531**:1-218

[14] Program NT. Toxicology and carcinogenesis studies of

2,3′,4,4′,5-pentachlorobiphenyl (PCB 118) (CAS No. 31508-00-6) in female harlan Sprague-Dawley rats (gavage studies). National Toxicology Program Technical Report Series. 2010;**559**:1-174

[15] N Program NT. TP toxicology and carcinogenesis studies of 3,3′,4,4′,5-pentachlorobiphenyl (PCB 126) (CAS No. 57465-28-8) in female Harlan Sprague-Dawley rats (Gavage Studies). National Toxicology Program Technical Report Series. 2006;**520**:4-246

[16] Bosetti C, Negri E, Fattore E, La Vecchia C. Occupational exposure to polychlorinated biphenyls and cancer risk. European Journal of Cancer Prevention: The Official Journal of the European Cancer Prevention Organisation (ECP). 2003;**12**(4):251-255

[17] Oliver JD, Roberts RA. Receptormediated hepatocarcinogenesis: Role of hepatocyte proliferation and apoptosis. Pharmacology & Toxicology. 2002;**91**(1):1-7

[18] Ludewig G, Robertson LW. Polychlorinated biphenyls (PCBs) as initiating agents in hepatocellular carcinoma. Cancer Letters. 2013;**334**(1):46-55

[19] Mayes BA, McConnell EE, Neal BH, Brunner MJ, Hamilton SB, Sullivan TM, et al. Comparative carcinogenicity in Sprague-Dawley rats of the polychlorinated biphenyl mixtures Aroclors 1016, 1242, 1254, and 1260. Toxicological Sciences. 1998;**41**(1):62-76

[20] Kashimoto T, Miyata H, Fukushima S, Kunita N, Ohi G, Tung TC. PCBs, PCQs and PCDFs in blood of yusho and yu-cheng patients. Environmental Health Perspectives. 1985;**59**:73-78

[21] Todaka T, Hirakawa H, Hori T, Tobiishi K, Iida T, Furue M. Concentrations of polychlorinated dibenzo-p-dioxins, polychlorinated dibenzofurans, and non-ortho and

mono-ortho polychlorinated biphenyls in blood of Yusho patients. Chemosphere. 2007;**66**(10):1983-1989

[22] Hsu ST, Ma CI, Hsu SK, Wu SS, Hsu NH, Yeh CC, et al. Discovery and epidemiology of PCB poisoning in Taiwan: A four-year followup. Environmental Health Perspectives. 1985;**59**:5-10

[23] Guo YL, Ryan JJ, Lau BP, Yu ML, Hsu CC. Blood serum levels of PCBs and PCDFs in Yucheng women 14 years after exposure to a toxic rice oil. Archives of Environmental Contamination and Toxicology. 1997;**33**(1):104-108

[24] Onozuka D, Nakamura Y, Tsuji G, Furue M. Mortality in Yusho patients exposed to polychlorinated biphenyls and polychlorinated dibenzofurans: a 50-year retrospective cohort study. Environmental Health. 2020;**19**(1):119

[25] Li MC, Chen PC, Tsai PC, Furue M, Onozuka D, Hagihara A, et al. Mortality after exposure to polychlorinated biphenyls and polychlorinated dibenzofurans: A meta-analysis of two highly exposed cohorts. International Journal of Cancer. 2015;**137**(6):1427-1432

[26] Cogliano VJ. Assessing the cancer risk from environmental PCBs. Environmental Health Perspectives. 1998;**106**(6):317-323

[27] Abramowicz DA. Aerobic and anaerobic PCB biodegradation in the environment. Environmental health perspectives. 1995;**103**(Suppl 5):97-99

[28] Gilbert JM, Baduel C, Li Y, Reichelt-Brushett AJ, Butcher PA, McGrath SP, et al. Bioaccumulation of PCBs in liver tissue of dusky Carcharhinus obscurus, sandbar C. plumbeus and white Carcharodon carcharias sharks from south-eastern Australian waters. Marine Pollution Bulletin. 2015;**101**(2):908-913

*Impact of PCBs, Furan and Dioxin on Hepatocarcinogenesis DOI: http://dx.doi.org/10.5772/intechopen.101526*

[29] Cogliano VJ. PCBs: Cancer Dose-Response Assessment and Application to Environmental Mixtures. Washington, DC, EPA/600/P-96/001F., editor: U.S. Environmental Protection Agency OoRaD, National Center for Environmental Assessment, Washington Office; 1996

[30] Stecca L, Tait S, Corrado F, Esposito M, Mantovani A, La Rocca C. Development of an in vitro test battery model based on liver and colon cancer cell lines to discriminate PCB mixtures by transcription factors gene expression analysis. Toxicology In Vitro. 2016;**34**:204-211

[31] Shan Q, Li H, Chen N, Qu F, Guo J. Understanding the Multiple Effects of PCBs on Lipid Metabolism. Diabetes, Metabolic Syndrome and Obesity: Targets and Therapy. 2020;**13**:3691-3702

[32] Silberhorn EM, Glauert HP, Robertson LW. Carcinogenicity of polyhalogenated biphenyls: PCBs and PBBs. Critical Reviews in Toxicology. 1990;**20**(6):440-496

[33] Brunner MJ, Sullivan TM, Singer AW, Ryan MJ, Toft Ii JD, Menton RS, et al. An Assessment of the Chronic Toxicity and Oncogenicity of Aroclor-1016, Aroclor-1242, Aroclor-1254, and Aroclor-1260 Administered in Diet to Rats: Chronic Toxicity and Oncogenicity Report. Vol. 1. Columbus, OH: Battelle; 1997 Report No.: Battelle Study No.SC920192

[34] Norback DH, Weltman RH. Polychlorinated biphenyl induction of hepatocellular carcinoma in the Sprague-Dawley rat. Environmental Health Perspectives. 1985;**60**:97-105

[35] Rignall B, Grote K, Gavrilov A, Weimer M, Kopp-Schneider A, Krause E, et al. Biological and tumorpromoting effects of dioxin-like and non-dioxin-like polychlorinated biphenyls in mouse liver after single or combined treatment. Toxicological Sciences. 2013;**133**(1):29-41

[36] Haag-Grönlund M, Johansson N, Fransson-Steen R, Hâkansson H, Scheu G, Wärngård L. Interactive effects of three structurally different polychlorinated biphenyls in a rat liver tumor promotion bioassay. Toxicology and Applied Pharmacology. 1998;**152**(1):153-165

[37] Stemm DN, Tharappel JC, Lehmler HJ, Srinivasan C, Morris JS, Spate VL, et al. Effect of dietary selenium on the promotion of hepatocarcinogenesis by 3,3′, 4,4′-tetrachlorobiphenyl and 2,2′, 4,4′, 5,5′-hexachlorobiphenyl. Experimental Biology and Medicine (Maywood, NJ). 2008;**233**(3):366-376

[38] Dean CE Jr, Benjamin SA, Chubb LS, Tessari JD, Keefe TJ. Nonadditive hepatic tumor promoting effects by a mixture of two structurally different polychlorinated biphenyls in female rat livers. Toxicological Sciences. 2002;**66**(1):54-61

[39] Zhang W, Sargis RM, Volden PA, Carmean CM, Sun XJ, Brady MJ. PCB 126 and other Dioxin-like PCBs specifically suppress hepatic PEPCK expression via the Aryl hydrocarbon receptor. PLoS One. 2012;**7**(5):e37103

[40] Glauert HP, Tharappel JC, Banerjee S, Chan NL, Kania-Korwel I, Lehmler HJ, et al. Inhibition of the promotion of hepatocarcinogenesis by 2,2′,4,4′,5,5′-hexachlorobiphenyl (PCB-153) by the deletion of the p50 subunit of NF-kappa B in mice. Toxicology and Applied Pharmacology. 2008;**232**(2):302-308

[41] Strathmann J, Schwarz M, Tharappel JC, Glauert HP, Spear BT, Robertson LW, et al. PCB 153, a nondioxin-like tumor promoter, selects for beta-catenin (Catnb)-mutated mouse liver tumors. Toxicological Sciences. 2006;**93**(1):34-40

[42] Vincent F, de Boer J, Pfohl-Leszkowicz A, Cherrel Y, Galgani F. Two cases of ras mutation associated with liver hyperplasia in dragonets (Callionymus lyra) exposed to polychlorinated biphenyls and polycyclic aromatic hydrocarbons. Molecular Carcinogenesis. 1998;**21**(2):121-127

[43] Andersen O, Lindegaard P, Unger M, Nordberg GF. Effects of liver damage induced by polychlorinated biphenyls (PCB) on cadmium metabolism in mice. Environmental Research. 1985;**38**(2):213-224

[44] Whysner J, Wang CX. Hepatocellular iron accumulation and increased cell proliferation in polychlorinated biphenyl-exposed Sprague-Dawley rats and the development of hepatocarcinogenesis. Toxicological Sciences. 2001;**62**(1): 36-45

[45] Brown DP. Mortality of workers exposed to polychlorinated biphenyls- -An update. Archives of Environmental Health. 1987;**42**(6):333-339

[46] Donato F, Moneda M, Portolani N, Rossini A, Molfino S, Ministrini S, et al. Polychlorinated biphenyls and risk of hepatocellular carcinoma in the population living in a highly polluted area in Italy. Scientific Reports. 2021;**11**(1):3064

[47] Zani C, Gelatti U, Donato F, Capelli M, Portolani N, Bergonzi R, et al. Polychlorinated biphenyls in serum, liver and adipose tissue of subjects with hepatocellular carcinoma living in a highly polluted area. Chemosphere. 2013;**91**(2):194-199

[48] Zhao G, Wang Z, Zhou H, Zhao Q. Burdens of PBBs, PBDEs, and PCBs in tissues of the cancer patients in the e-waste disassembly sites in Zhejiang, China. Science of the Total Environment. 2009;**407**(17):4831-4837 [49] Kaifie A, Schettgen T, Gube M, Ziegler P, Kraus T, Esser A. Functional and structural liver abnormalities in former PCB exposed workers – analyses from the HELPcB cohort. Journal of Toxicology and Environmental Health, Part A. 2019;**82**(1):52-61

[50] Clair HB, Pinkston CM, Rai SN, Pavuk M, Dutton ND, Brock GN, et al. Liver disease in a residential Cohort with elevated polychlorinated biphenyl exposures. Toxicological Sciences. 2018;**164**(1):39-49

[51] Loomis D, Browning SR, Schenck AP, Gregory E, Savitz DA. Cancer mortality among electric utility workers exposed to polychlorinated biphenyls. Occupational and Environmental Medicine. 1997;**54**(10):720-728

[52] Niehoff NM, Zabor EC, Satagopan J, Widell A, O'Brien TR, Zhang M, et al. Prediagnostic serum polychlorinated biphenyl concentrations and primary liver cancer: A case-control study nested within two prospective cohorts. Environmental Research. 2020;**187**:109690

[53] Delistraty D. Ecotoxicity and risk to human fish consumers of polychlorinated biphenyls in fish near the Hanford Site (USA). Science of the Total Environment. 2013;**445-446**:14-21

[54] Faust D, Vondráček J, Krčmář P, Smerdová L, Procházková J, Hrubá E, et al. AhR-mediated changes in global gene expression in rat liver progenitor cells. Archives of Toxicology. 2013;**87**(4): 681-698

[55] Borlak J, Jenke HS. Cross-talk between aryl hydrocarbon receptor and mitogenactivated protein kinase signaling pathway in liver cancer through c-raf transcriptional regulation. Molecular Cancer Research. 2008;**6**(8):1326-1336

[56] Buchmann A, Kunz W, Wolf CR, Oesch F, Robertson LW. Polychlorinated *Impact of PCBs, Furan and Dioxin on Hepatocarcinogenesis DOI: http://dx.doi.org/10.5772/intechopen.101526*

biphenyls, classified as either phenobarbital- or 3-methylcholanthrene-type inducers of cytochrome P-450, are both hepatic tumor promoters in diethylnitrosamineinitiated rats. Cancer Letters. 1986;**32**(3):243-253

[57] Shen H, Robertson LW, Ludewig G. Regulatory effects of dioxin-like and non-dioxin-like PCBs and other AhR ligands on the antioxidant enzymes paraoxonase 1/2/3. Environmental Science and Pollution Research International. 2016;**23**(3):2108-2118

[58] Oakley GG, Devanaboyina U, Robertson LW, Gupta RC. Oxidative DNA damage induced by activation of polychlorinated biphenyls (PCBs): Implications for PCB-induced oxidative stress in breast cancer. Chemical Research in Toxicology. 1996;**9**(8):1285-1292

[59] Birnbaum LS. Endocrine effects of prenatal exposure to PCBs, dioxins, and other xenobiotics: Implications for policy and future research. Environmental Health Perspectives. 1994;**102**(8):676-679

[60] Andersson PL, Blom A, Johannisson A, Pesonen M, Tysklind M, Berg AH, et al. Assessment of PCBs and hydroxylated PCBs as potential xenoestrogens: In vitro studies based on MCF-7 cell proliferation and induction of vitellogenin in primary culture of rainbow trout hepatocytes. Archives of Environmental Contamination and Toxicology. 1999;**37**(2):145-150

[61] Ho PW, Garner CE, Ho JW, Leung KC, Chu AC, Kwok KH, et al. Estrogenic phenol and catechol metabolites of PCBs modulate catechol-O-methyltransferase expression via the estrogen receptor: Potential contribution to cancer risk. Current Drug Metabolism. 2008;**9**(4):304-309

[62] Deng P, Barney J, Petriello MC, Morris AJ, Wahlang B, Hennig B.

Hepatic metabolomics reveals that liver injury increases PCB 126-induced oxidative stress and metabolic dysfunction. Chemosphere. 2019;**217**:140-149

[63] Yuan J, Lu WQ, Zou YL, Wei W, Zhang C, Xie H, et al. Influence of aroclor 1254 on benzo(a)pyreneinduced DNA breakage, oxidative DNA damage, and cytochrome P4501A activity in human hepatoma cell line. Environmental Toxicology. 2009;**24**(4): 327-333

[64] Al-Anati L, Högberg J, Stenius U. Non-dioxin-like-PCBs phosphorylate Mdm2 at Ser166 and attenuate the p53 response in HepG2 cells. Chemico-Biological Interactions. 2009;**182**(2-3): 191-198

[65] Chen Y, Wu Y, Xiao W, Jia H, Glatt H, Shi M, et al. Human CYP1B1 dependent genotoxicity of dioxin-like polychlorinated biphenyls in mammalian cells. Toxicology. 2020;**429**:152329

[66] Hu K, Yu H, Li Z, Jin G, Jia H, Song M, et al. Human CYP2E1-activated mutagenicity of dioxin-like PCBs 105 and 118-Experimental data consistent with molecular docking results. Toxicology. 2020;**437**:152438

[67] De Haan LH, Simons JW, Bos AT, Aarts JM, Denison MS, Brouwer A. Inhibition of intercellular communication by 2,3,7,8-tetrachlorodibenzo-p-dioxin and dioxin-like PCBs in mouse hepatoma cells (Hepa1c1c7): Involvement of the Ah receptor. Toxicology and Applied Pharmacology. 1994;**129**(2):283-293

[68] VanEtten SL, Bonner MR, Ren X, Birnbaum LS, Kostyniak PJ, Wang J, et al. Telomeres as targets for the toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and polychlorinated biphenyls (PCBs) in rats. Toxicology and Applied Pharmacology. 2020;**408**:115264

[69] de Conti A, Beland FA, Pogribny IP. The role of epigenomic alterations in furan-induced hepatobiliary pathologies. Food and Chemical Toxicology. 2017;**109**:677-682

[70] Lu D, Sullivan MM, Phillips MB, Peterson LA. Degraded protein adducts of cis-2-butene-1, 4-dial are urinary and hepatocyte metabolites of furan. Chemical Research in Toxicology. 2009;**22**(6):997-1007

[71] Grill AE, Schmitt T, Gates LA, Lu D, Bandyopadhyay D, Yuan J-M, et al. Abundant rodent furan-derived urinary metabolites are associated with tobacco smoke exposure in humans. 2015;**28**(7):1508-1516

[72] Knutsen HK, Alexander J, Barregård L, Bignami M, Brüschweiler B, Ceccatelli S, et al. Risks for public health related to the presence of furan and methylfurans in food. EFSA journal European Food Safety Authority 2017;**15**(10):e05005

[73] Mariotti MS, Granby K, Rozowski J, Pedreschi FJF. Furan: a critical heat induced dietary contaminant. Food & Function. 2013;**4**(7):1001-1015

[74] LAJ P. Electrophilic intermediates produced by bioactivation of furan. Drug Metabolism Reviews. 2006;**38**(4):615-626

[75] Phillips MB, Sullivan MM, Villalta PW, Peterson LA. Covalent modification of cytochrome c by reactive metabolites of furan. Chemical Research in Toxicology. 2014;**27**(1):129-135

[76] Nunes J, Martins IL, Charneira C, Pogribny IP, de Conti A, Beland FA, et al. New insights into the molecular mechanisms of chemical carcinogenesis: In vivo adduction of histone H2B by a reactive metabolite of the chemical carcinogen furan. 2016;**264**:106-113

[77] Jackson AF, Williams A, Recio L, Waters MD, Lambert IB, Yauk CLJT, et al. Case study on the utility of hepatic global gene expression profiling in the risk assessment of the carcinogen furan. 2014;**274**(1):63-77

[78] Gill S, Kavanagh M, Barker M, Weld M, Vavasour E, Hou Y, et al. Subchronic oral toxicity study of furan in B6C3F1 Mice. Toxicol Pathol. 2011;**39**(5):787-794

[79] Ramm S, Limbeck E, Mally A. Functional and cellular consequences of covalent target protein modification by furan in rat liver. Toxicology. 2016;**361**:49-61

[80] Moro S, Chipman JK, Antczak P, Turan N, Dekant W, Falciani F, et al. Identification and pathway mapping of furan target proteins reveal mitochondrial energy production and redox regulation as critical targets of furan toxicity. 2012;**126**(2):336-352

[81] Selmanoğlu G, Karacaoğlu E, Kılıç A, Koçkaya EA, Akay MT. Toxicity of food contaminant furan on liver and kidney of growing male rats. Environmental Toxicology. 2012;**27**(10):613-622

[82] Terrell AN, Huynh M, Grill AE, Kovi RC, O'Sullivan MG, Guttenplan JB, et al. Mutagenicity of furan in female Big Blue B6C3F1 mice. 2014;**770**:46-54

[83] Ding W, Petibone DM, Latendresse JR, Pearce MG, Muskhelishvili L, White GA, et al. In vivo genotoxicity of furan in F344 rats at cancer bioassay doses. 2012;**261**(2):164-171

[84] Takasu S, Ishii Y, Kijima A, Ogawa K, Nakane S, Umemura T. Furan induced characteristic glutathione S-transferase placental form-positive Foci in terms of cell kinetics and gene expression. Toxicologic Pathology. 2020;**48**(6):756-765

[85] Hibi D, Yokoo Y, Suzuki Y, Ishii Y, Jin M, Kijima A, et al. Lack of genotoxic *Impact of PCBs, Furan and Dioxin on Hepatocarcinogenesis DOI: http://dx.doi.org/10.5772/intechopen.101526*

mechanisms in early-stage furaninduced hepatocellular tumorigenesis in gpt delta rats. 2017;**37**(2):142-149

[86] Hickling K, Hitchcock J, Chipman J, Hammond T, Evans JJ. Induction and progression of cholangiofibrosis in rat liver injured by oral administration of furan. Toxicologic Pathology. 2010;**38**(2):213-229

[87] Hickling KC, Hitchcock JM, Oreffo V, Mally A, Hammond TG, Evans JG, et al. Evidence of oxidative stress and associated DNA damage, increased proliferative drive, and altered gene expression in rat liver produced by the cholangiocarcinogenic agent furan. Toxicol Pathol. 2010;**38**(2):230-243

[88] Tryndyak V, de Conti A, Doerge DR, Olson GR, Beland FA, Pogribny IP. Furan-induced transcriptomic and gene-specific DNA methylation changes in the livers of Fischer 344 rats in a 2-year carcinogenicity study. Archives of Toxicology. 2017;**91**(3):1233-1243

[89] Durling JKL, Svensson K, Abramsson-Zetterberg L. Furan is not genotoxic in the micronucleus assay in vivo or in vitro. Toxicology Letters. 2007;**169**(1):43-50

[90] McDaniel LP, Ding W, Dobrovolsky VN, Shaddock JG Jr, Mittelstaedt RA, Doerge DR, et al. Genotoxicity of furan in Big Blue rats. 2012;**742**(1-2):72-78

[91] Chen T, Williams TD, Mally A, Hamberger C, Mirbahai L, Hickling K, et al. Gene expression and epigenetic changes by furan in rat liver. 2012;**292**(2-3):63-70

[92] Conti A, Kobets T, Escudero-Lourdes C, Montgomery B, Tryndyak V, Beland FA, et al. Dose-and time-dependent epigenetic changes in the livers of Fisher 344 rats exposed to furan. 2014;**139**(2):371-380

[93] Recio L, Phillips SL, Maynor T, Waters M, Jackson AF, Yauk CL. Differential expression of long noncoding RNAs in the livers of female B6C3F1 mice exposed to the carcinogen furan. Toxicological Sciences. 2013;**135**(2):369-379

[94] Wong MH, Armour M-A, Naidu R, Man M. Persistent toxic substances: Sources, fates and effects. Reviews on Environmental Health. 2012;**27**(4):207-213

[95] English JSC, Dawe RS, Ferguson JJ. Environmental effects and skin disease. British Medical Bulletin. 2003;**68**(1): 129-142

[96] Ju Q, Zouboulis CJB. Effects of 2, 3, 7, 8-tetrachlorodibenzo-p-dioxin on the skin. 2013;**2, 113** 2167-0331.1000113

[97] Pelclová D, Urban P, Preiss J, Lukáš E, Fenclová Z, Navrátil T, et al. Adverse health effects in humans exposed to 2, 3, 7, 8-tetrachlorodibenzop-dioxin (TCDD). 2006;**21**(2):119-138

[98] Çiftçi O. The Investigation of effect mechanism, chemical structure and toxicokinetics properties of dioxins compounds. Annals of Medical reaserch. 2010;**17**(4):413-422

[99] Bentli R, Ciftci O, Cetin A, Otlu AJT. Montelukast, an antiinflamattory agent, can prevent toxic effects of 2, 3, 7, 8-tetrachlorodibenzo-p-dioxin in terms of oxidative stress and histological alterations in liver and serum cytokine levels. Toxicology and Industrial Health. 2013

[100] Türkez H, Geyikoğlu F, Yousef MI, Toğar B, Vançelik SJT. Propolis alleviates 2, 3, 7, 8-Tetrachlorodibenzo-p-dioxininduced histological changes, oxidative stress and DNA damage in rat liver. Toxicology and Industrial Health. 2013;**29**(8):677-685

[101] Mejia-Garcia A, Sanchez-Ocampo EM, Galindo-Gomez S, Shibayama M, Reyes-Hernandez O, Guzman-Leon S, et al. 2, 3, 7, 8-Tetrachlorodibenzo-p-dioxin enhances CCl4-induced hepatotoxicity in an aryl hydrocarbon receptor-dependent manner. 2013;**43**(2):161-168

[102] Lee JH, Wada T, Febbraio M, He J, Matsubara T, Lee MJ, et al. A novel role for the dioxin receptor in fatty acid metabolism and hepatic steatosis. 2010;**139**(2):653-663

[103] Gonzalez JF, Fernandez-Salguero P. The aryl hydrocarbon receptor: Studies using the AHR-null mice. Drug Metabolism and Disposition. 1998;**26**(12):1194-1198

[104] Elizondo G, Fernandez-Salguero P, Sheikh MS, Kim G-Y, Fornace AJ, Lee KS, et al. Altered cell cycle control at the G2/M phases in aryl hydrocarbon receptor-null embryo fibroblast. 2000;**57**(5):1056-1063

[105] Sato S, Shirakawa H, Tomita S, Ohsaki Y, Haketa K, Tooi O, et al. Low-dose dioxins alter gene expression related to cholesterol biosynthesis, lipogenesis, and glucose metabolism through the aryl hydrocarbon receptormediated pathway in mouse liver. 2008;**229**(1):10-19

[106] Reyes-Hernandez O, Mejia-Garcia A, Sánchez-Ocampo E, Cabanas-Cortes M, Ramirez P, Chávez-González L, et al. Ube2l3 gene expression is modulated by activation of the aryl hydrocarbon receptor: Implications for p53 ubiquitination. 2010;**80**(6):932-940

[107] Walisser JA, Glover E, Pande K, Liss AL, Bradfield CA. Aryl hydrocarbon receptor-dependent liver development and hepatotoxicity are mediated by different cell types. Proceedings of the National Academy of Sciences of the United States of America. 2005;**102**(49):17858-17863

[108] Boutros PC, Bielefeld KA, Pohjanvirta R, Harper PA.

Dioxin-dependent and dioxinindependent gene batteries: Comparison of liver and kidney in AHR-null mice. Toxicological Sciences. 2009;**112**(1):245-256

[109] Kennedy GD, Nukaya M, Moran SM, Glover E, Weinberg S, Balbo S, et al. Liver tumor promotion by 2, 3, 7, 8-tetrachlorodibenzo-p-dioxin is dependent on the aryl hydrocarbon receptor and TNF/IL-1 receptors. 2014;**140**(1):135-143

[110] Yamaguchi M, Hankinson O. 2, 3, 7, 8-Tetrachlorodibenzo-p-dioxin suppresses the growth of human liver cancer HepG2 cells in vitro: Involvement of cell signaling factors. International Journal of Oncology. 2018;**53**(4): 1657-1666

[111] Türkez H, Geyikoglu F, Yousef MI. Ameliorative effect of docosahexaenoic acid on 2, 3, 7, 8-tetrachlorodibenzo-pdioxin-induced histological changes, oxidative stress, and DNA damage in rat liver. Toxicology and Industrial Health. 2012;**28**(8):687-696

[112] Alsharif NZ, Hassoun EA. Protective effects of vitamin A and vitamin E succinate against 2, 3, 7, 8-tetrachlorodibenzo-p-dioxin (TCDD)-induced body wasting, hepatomegaly, thymic atrophy, production of reactive oxygen species and DNA damage in C57BL/6J mice. Basic & Clinical Pharmacology & Toxicology. 2004;**95**(3):131-138

[113] Hassoun EA, Vodhanel J, Abushaban A. The modulatory effects of ellagic acid and vitamin E succinate on TCDD-induced oxidative stress in different brain regions of rats after subchronic exposure. Journal of Biochemical and Molecular Toxicology. 2004;**18**(4):196-203

[114] Ahd K, Dhibi S, Akermi S, Bouzenna H, Samout N, Elfeki A, et al. Protective effect of ginger (Zingiber

*Impact of PCBs, Furan and Dioxin on Hepatocarcinogenesis DOI: http://dx.doi.org/10.5772/intechopen.101526*

officinale) against PCB-induced acute hepatotoxicity in male rats. 2019;**9**(50):29120-29130

[115] Ciftci O, Ozdemir I. Protective effects of quercetin and chrysin against 2, 3, 7, 8-tetrachlorodibenzo-p-dioxin (TCDD) induced oxidative stress, body wasting and altered cytokine productions in rats. 2011;**33**(3):504-508

[116] Kim H-S, Park S-Y, Yoo K-Y, Lee SK, Jung W-W. Induction of heat shock proteins and antioxidant enzymes in 2, 3, 7, 8-TCDD-induced hepatotoxicity in rats. The Korean Journal of Physiology & Pharmacology. 2012;**16**(6):469-476

[117] Czepiel J, Biesiada G, Gajda M, Szczepański W, Szypuła K, Dąbrowski Z, et al. The effect of TCDD dioxin on the rat liver in biochemical and histological assessment. 2009;**58**(1-2):85-90

[118] Viluksela M, Pohjanvirta R. Multigenerational and transgenerational effects of dioxins. International Journal of Molecular Sciences. 2019;**20**(12):2947

[119] Boutros PC, Yan R, Moffat ID, Pohjanvirta R, Okey AB. Transcriptomic responses to 2, 3, 7, 8-tetrachlorodibenzo-p-dioxin (TCDD) in liver: comparison of rat and mouse. BMC Genomics. 2008;**9**(1):1-17

[120] Boverhof DR, Burgoon LD, Tashiro C, Sharratt B, Chittim B, Harkema JR, et al. Comparative toxicogenomic analysis of the hepatotoxic effects of TCDD in Sprague Dawley rats and C57BL/6 mice. 2006;**94**(2):398-416

[121] Zhang X, Li C, Gong Z. Development of a convenient in vivo hepatotoxin assay using a transgenic zebrafish line with liver-specific DsRed expression. 2014;**9**(3):e91874

[122] Yang Q, Salim L, Yan C, Gong Z. Rapid Analysis of Effects of

Environmental Toxicants on Tumorigenesis and Inflammation Using a Transgenic Zebrafish Model for Liver Cancer. Marine Biotechnology (New York, NY). 2019;**21**(3):396-405

#### **Chapter 4**

## Persistent Organic Pollutants in the Bizerte Lagoon Ecosystem: Occurrence, Distribution, and Ecotoxicological Assessment Using Marine Organism

*Ahmed Nasri, Takoua Mhadhbi, Mohamed Allouche, Amel Hannachi, Hary Demey, Patricia Aïssa, Hamouda Beyrem and Ezzeddine Mahmoudi*

#### **Abstract**

Marine ecosystem represents an ecologically and economically important water bodies for human and animal living. Their increasing pollution by persistent organic pollutants has represented a major environmental alarm during the last years. In the current study, we examined the occurrence, local distribution and ecotoxicological menace of organic pollutants, comprising brominated flame retardants (BFR), polychlorinated biphenyls (PCBs), polycyclic aromatic hydrocarbons (PAHs), and organochlorine pesticides (OCPs) in different matrices from the Northern Tunisian Coastal Ecosystem (Bizerte lagoon). The pollutant existence in this biome is related with a negative impact on the biocenosis health. Many approach including (i) chemical analyses; (ii) taxonomic structure and ecological indices analyses; (iii) and biochemical experimental studies, were investigated to determine the ecosystem quality and the contaminant effects. Our chapter introduces the baseline information on the organic contaminations extent and toxicological impact, as well as, it contribute to evaluate the ecological quality of this marine coastal ecosystem.

**Keywords:** Persistent Organic Pollutants (POPs), Mediterranean coastal lagoon, Occurrence, Distribution, Ecotoxicological evaluation

#### **1. Introduction**

The expansion of the anthropoid industrial activities has led to the pollution of many ecosystems. The daily release of agricultural, industrial and commercial chemicals into the aquatic ecosystem has induced various toxic effects on marine organisms [1]. Among Mediterranean coastal environments, many lagoons are exposed to human-induced pressures. The Bizerte lagoon is one of them, located near an industrial and agricultural area of Northern Tunisia, known as a receptor of numerous industrial wastes, chemical fertilizers and pesticides via runoff and soil erosion, conducting to a lessening in fish and bivalves' production [2].

These organisms accumulate chemicals directly from contaminated water and indirectly via the food chain. In addition, these are used as sentinels' species because they have the capacity to bio-concentrate toxic compounds. Their main advantage is the rapid response to low concentrations of aquatic contaminant [3].

Biomonitoring of the aquatic environment has based both on the quantification of pollutant concentrations in bioindicator organisms and on the biological analyses (ecological indices and biomarkers). The chemical analyses were performed after pollutants extraction from different matrices (water, sediment, and marine organism), using instruments by Gas chromatography separation connected or not to a mass spectrometer (GC–MS) [4–15]. Nevertheless, the biological analyses were performed both firstly; by nematodes taxonomic structure and ecological indices determination to assess the chemicals toxicity, secondly, by biomarkers assessment who are known as tools for contamination impacts evaluation in the marine environment. Some of these have been incorporated in ecological monitoring programs [16].

The biomarkers have been classified in different types, such as biomarkers exposure and response, or as general and specific biomarkers [17]. Antioxidant enzymes such as catalase (CAT), reduced glutathione (GSH) and Glutathione-S-transferase (GST) have been used as biomarkers reflecting the exposure and toxicity of contaminants [18]. The Acetylcholinesterase (AChE) activity has been used as a biomarker of exposure to several pesticides (organophosphate and carbamate insecticides) in aquatic environments [19]. The high concentrations of Malondialdehyde (MDA) have been reflected the lipid peroxidation expression that indicates cell damage [20].

This aim of the present chapter was to describe the previous studies focused on the determination of pollutant concentrations in bio-indicator organisms and to discuss the results of biological assessment of their toxicities in order to provide clearer and more informative data concerning the Bizerte lagoon coastal ecosystem state.

#### **2. Persistent organic pollutants (POP)**

#### **2.1 Definition**

Persistent organic pollutants (POPs) are organic chemical compounds having physical and chemical properties that result in their widespread dissemination, persistence and accumulation. They are known by their characteristics: persistence, bioaccumulation, transport, and toxicity effects. POPs have a lifespan that can reach many years, because of their resistance to chemical and biological degradation. They are lipophilic and tend to disperse widely across ecosystems; they accumulate to high concentrations in the wildlife tissues biomagnifying up the food chain. They are semi-volatile, allowing for long-range transport in air or absorbed to particulate matter in environmental matrices such as air or water.

These chemicals are organic substances of natural or anthropogenic origin, once released into the environment; they remain intact for exceptionally long periods of time [21]. The major intentional man-made sources of POPs are industrial sources like incinerating plants, power stations, agricultural sprays, heating stations and evaporation from water surfaces, soil, or from the landfills.

#### **3. Sampling location**

#### **3.1 Bizerte lagoon**

The Bizerte Lagoon is located in the Tunisia northern littoral (Mediterranean Sea) (latitude, 37°80′–37°14′ N; longitude, 9°46′–9°56′ E). It spread 128 km2 of

*Persistent Organic Pollutants in the Bizerte Lagoon Ecosystem: Occurrence, Distribution, and… DOI: http://dx.doi.org/10.5772/intechopen.100549*

**Figure 1.** *Location of the Bizerte lagoon (northern Tunisia).*

area and 7 m of mean depth. In the north it communicates by a 7-km-long canal with the Mediterranean Sea, nonetheless, in the South with the Ichkeul Lake by the River of Tinja. This lagoon is under tributaries of many rivers such as Gueniche, Ben Hassine, Mrezig, Garek, and Tinja has a high importance in terms of seafood products exports; it show an economically significant body of water due to a variety of peach and aquaculture activities (oysters and mussels farming).

Due the population growth and technological development around the lagoon, this ecosystem is compelling to the influence of divers' physical factors that highly fluctuate during the year. It is severely affected by human activities on its shores, including the four harbors presence, urban development (bounded by the cities of Bizerte, Zarzouna, Menzel Jemil, Menzel Abderrahmane, Menzel Bourguiba and Tinja), and receives urban effluents, agricultural inputs and industrial discharges from many industrial units related to various fields of activity such as cement, steel plant, a metal factory, and a refinery (**Figure 1**) [22].

In the lagoon of Bizerte, the water column is under the influence of the winter contributions of rains and fresh water in provenance of the wadis, particularly that of Ichkeul [23]. On the other hand, the water column is under the influence of the winter contributions of rains and fresh water in provenance of the wadis, particularly that of Ichkeul [23]. On the other hand, the water exchange between the Mediterranean and the Ichkeul make the temperature and salinity of the lagoon vary between (10–29°C) and (32.5–38.5‰), respectively.

#### **4. POPs in the Bizerte lagoon**

Due to their persistent nature and ability to undergo long-range transport, POPs are ubiquitous in the environment. They are found in soil, lake and river sediment, benthos, water columns and the atmosphere. They enter and accumulate in the food chain through various pathways, such as deposition onto crops, soil ingestion by grazing animals, and bioaccumulation up through trophic levels. At concentrations typically found in food, the adverse health effects caused by POPs are almost entirely chronic (including cancer), disruption of the endocrine system, neurotoxicity, and development damage.

The occurrence of POPs is reported major in the aquatic environment pollution of which released into water bodies from a large variety of anthropogenic sources such as agricultural and municipal waste effluents, industrial coastal activities, atmospheric deposition, maritime transport and accidental spill [24, 25]. Many studies have been carried out for monitoring the organic pollutants and their impact on different aquatic ecosystems around the Bizerte lagoon. The most identified POP chemicals in Bizerte lagoon ecosystems are belonging the families of brominated (and chlorinated) aromatics, including brominated flame retardants (BFR) such as polybrominated diphenyl ethers (PBDEs) and their methoxylated analogues (MeO-PBDEs) [4–9]; polychlorinated biphenyls (PCBs) [10–12]; polycyclic aromatic hydrocarbons PAHs [13, 14]; and organochlorine pesticides (OCPs) [11, 12, 15].

#### **4.1 Brominated flame retardants (BFRs)**

Brominated flame retardants (BFRs) are brominated organic compounds used as additive flame retardants in plastics, paints, textiles, electronics, and vehicles. PBDEs and analogs (such as the methoxylated PBDEs (MeO-PBDEs)), have been found in many environmental matrices such as, aquatic organism, water, sludge, and sediment [26], and in fish and marine mammals [27], respectively. Contamination by BFRs is of environmental concern due to their persistence, potential for endocrine disruption and bioaccumulation, and long-range transport. It has been reported their adverse effect, such as morphological, immunological, and behavioral modifications, and enzyme induction [28, 29].

The levels of the BFRs in two species like in sea bass (*Dicentrarchus labrax*) and mullet (*Mugil cephalus*) collected from the Bizerte Lagoon were examined. The PBDE mean concentrations in fish were 45.3 and 96.2 ng/g lw respectively in mullet and sea bass. MeO PBDE concentrations in mullet and sea bass were 6.46–286 ng/g lw and 49.4–798 ng/g lw, respectively [4]. In other study, polybrominated diphenylethers (PBDEs) and methoxylated polybrominated diphenyl ethers (MeO-PBDEs) were determined in *solea solea* muscle. Mean levels of these compounds were 279 ng/g lw in sole [5]. In addition, sediment and urchins were used for the levels analysis of halogenated flame retardants (HFRs) and the methoxylated-PBDEs. The sediment concentrations were found between nd – 51.8 ng/g dry weight (dw) for the HFRs, and not detected for rest compounds. However, the levels of polybrominated diphenyl ethers (PBDEs), halogenated norbornenes (HN) and methoxylated PBDEs (MeO-PBDEs) were 3.67 to 56.9 ng/g lipid weight (lw), 4.52 to 116 ng/g lw and nd - 364 ng/g lw, respectively, in sea urchins [8].

Ameur et al. determined the concentrations of HFRs and MeO-PBDEs in *Hexaplex trunculus* collected in the same ecosystem and found that the mean tissue levels vary between 187 and 264 ng/g lw [6]. Recently, Mekni et al. measured the presence of the Brominated flame-retardants in sediment and fish eel (*Anguilla Anguilla*) samples [9]. Sediment HFR (halogenated flame-retardants) levels were 3.30–28.5 ng/g dry weight (dw), while OPFR (organophosphate flame retardants) levels were 9.77–164 ng/g dw [9]. However, the levels were 4.72–151 ng/g lipid weight (lw) and 19.7–2154 ng/g lw were estimated in fish, respectively.

#### **4.2 Polychlorinated biphenyls (PCBs)**

PCBs are produced and used in a wide range of industrial applications, such as inks, coatings, electrical transformers, and paints. Despite their prohibitions, they can also be found in various compartments such as sediments [4, 10, 30], in biota [31, 32], in human blood, breast milk and serum [33, 34].

*Persistent Organic Pollutants in the Bizerte Lagoon Ecosystem: Occurrence, Distribution, and… DOI: http://dx.doi.org/10.5772/intechopen.100549*

Concentrations of polychlorinated biphenyls (PCBs) were determined in sediments at range between 0.89 and 6.63 ng/g dw [10], and in fish species collected from Bizerte Lagoon were found 164 to 336 ng/g lw and from 282 to 642 ng/g lw, respectively in mullet (*Mugil cephalus*) and sea bass (*Dicentrarchus labrax*). PCB-118, PCB-138, PCB-153 and PCB-180 were dominant contaminants in the studied fish species, accounting respectively for 9.00%, 14.0%, 28.5% and 23.6% of total PCBs [11]. Polychlorinated biphenyls (PCBs) were measured in fish (*solea solea*) muscle and the mean levels were 1417 and 315 ng/g lipid weight (lw) [5]. Recently, the concentration of polychlorinated biphenyls (PCBs) identified in water samples were ranged between 3 and 10.4 ng L−1. PCB-28 (0.2–1.4 ng L−1) and PCB-52 (0.8– 3.5 ng L−1) were the predominant PCB congeners [35].

#### **4.3 Polycyclic aromatic hydrocarbons (PAHs)**

Polycyclic aromatic hydrocarbons (PAHs) are an important class of persistent organic pollutants derived from anthropogenic or natural sources. They are formed as a consequence of incomplete combustion of organic matter; they are also found in the crude oil components and derived products. Natural sources of PAHs include forest-fires and post-depositional transformation of biogenic precursors [36]. PAHs reach the marine environment via such sources as effluent discharges, urban runoff, atmospheric transport, and agricultural runoff. Because of their low water solubility and their hydrophobicity, PAHs in the aquatic environment rapidly become associated with inorganic and organic suspended particles [37, 38] and subsequent deposition in sediment. PAHs accumulate in aquatic sediments after adsorption to particles due to their high hydrophobicity and low solubility [38, 39].

Total PAHs identified in sediments from 10 stations around the Bizerte lagoon (Menzel Abderahmen, Menzel Jemil, Oued Guenniche, Oued Garek, Oued Ben Hasssin, Mnzel Bourguiba…) were ranged between 83.3–447 (ng/g dry wt) [13] with a mean value of 218 ng/g dry wt. Only anthracene, acenaphthene, and phenanthrene are found in all stations. Recently, Barhoumi et al. have been showed that form many station (n = 18), the total concentration varied between 16.9 to 394.1 (ng/g dry wt) with a mean concentration of 85.5 ng/g dry wt [14]. The maximum levels of PAHs were found along the channel and were ranged between 160.2–394.1 (ng/g dw), followed by station at the nearness of the mouth of the Tinja River which is affected by agricultural inputs (102.8 ng/g dw). The same researchers have showed the presence of PAHs levels in two different species, mussels (*Mytilus galloprovincialis*) (107.4 to 430.7 ng/g dw) and fish (*Anguilla anguilla*) (114.5–133.7 ng/g dw). In addition, Naphthalene was the most important hydrocarbon identified in mussels (31.5–272.6 ng/g dw) and fish (57.9–68.6 ng/g dw) [40]. More recently, 16 PAHs were measured in 40 surface sediment samples from the Bizerte lagoon and the high concentrations of total PAHs were 122–19600 ng/g [15].

#### **4.4 Organochlorine pesticides (OCPs)**

The organochlorine pesticides (OCPs) represent one of the most known persistent organic pollutants (POPs), which have caused great concern all over the world as a result of their persistence for many years, high lipophilicity, long range transportation, chronic and acute toxicities and bioaccumulation [41]. The OCPs have been produced and used for agricultural and industrial purposes for a long time and on a large scale. Organochlorine pesticides had been used throughout the world thanks to exceptional insecticidal and fungicidal properties [42].

The OCPs concentrations were measured in fish such as *Mugil cephalus* and *Dicentrarchus labrax* and the mean levels were found from 52.9 to 157 ng/g lw and from 158 to 265 ng/g lw, respectively [11]. In other study, Barhoumi et al. have found that the total sediments concentrations of OCPs from 1.1 to 14.0 ng/g dw (average value, 3.3 ng/g dw) [12]. Among the OCPs, the concentrations range of DDTs (dichlorodiphenyltrichloroethane) in addition with its metabolites and HCB (hexachlorobenzene) were 0.3–11.5 ng/g dw (1.9 ng/g dw) and 0.6–2.5 ng/g dw (1.4 ng/g dw), respectively. Lately, sediment has examined for study the vertical distribution of organochlorine pesticides (OCPs). The OCPs concentrations ranged from 26.98 ± 0.04 ng/g found at 3 cm depth and 10.23 ± 0.02 ng/g at 6 cm depth in site SC1 (located in the mouth of the channel connecting the lagoon to the Mediterranean Sea). However, in station SC2 (located in the metallurgy of Menzel bourguiba), the concentrations are of the order of 11.77 ± 0.11 ng/g (9 cm deep) and 1.47 ± 0.02 ng/g (20 cm deep) [43].

#### **5. POPs impact on marine organism**

The occurrence of chemicals compounds in aquatic environment has led researchers to examine biological effects on aquatic organisms. Persistent organic pollutants (POPs) are toxic chemicals that adversely affect organism health and the environment. Many effects are associated to POPs exposure such as immune alteration, reproductive disorder, endocrine disruption, and neurological disorders. In the Bizerte lagoon, the use of marine organisms as bio-indicators of contaminant loads (and as models of laboratory toxicity studies), has enabled the gathering of information on the state of the ecosystem under consideration.

Laboratory research has documented many effects of POPs in a wide range of aquatic organism. Toxicity of these compounds are associated to various biochemical and population effect in benthic nematodes, marine phytoplankton, and bivalves. Among these studies, Louati et al. studied the response of microbial communities following sediment enrichment in anthracene as well as the monitoring of their biomass, activity and composition following the bioremediation [44]. A significant reduction of bacterial abundance, a strongly oxygen consumption inhibition, as well as, a microbial structure modification in comparison with the control were registered. In other study, marine nematodes communities were subjected to a 100 ppm polycyclic aromatic hydrocarbons mixture of phenanthrene, fluoranthene, and pyrene during 30 days. Abundance, diversity, and taxonomic structure were changed. *Spirinia parasitifera* became the resistant species to PAHs contamination while *Oncholaimus campylocercoides* and *Neochromadora peocilosoma* were strongly inhibited [45].

Othman et al. studied the toxicity of PAHs mixtures on natural phytoplankton communities [46]. Results show that PAHs decreased the photosynthetic potential with a dramatic change in phytoplankton community composition (size classes and chlorophyll a) were strongly affected. The diatom *Entomoneis paludosa* appeared favored under PAH exposure as evidenced by increase in cell density, whereas autotrophic flagellates and dinophytes were strongly reduced. In other study, the anthracene toxicity was studied on the benthic bivalves *Ruditapes decussatus.* A change of the siphon movement and decreasing filtration rate were recorded. In addition, the oxidative stress status of the gills was affected with the modification of proteins [47]. Exposure of the pelagic organism (*Mytilus galloprovincialis*) to anthracene induced the enzymatic activities such as, acetylcholinesterase (AChE), glutathione (GSH), and malondialdehyde (MDA) in digestive gland. A reduction of the filtration rates and an increase in lipid peroxidation in digestive gland were registered. Finally, the GSH content as well as AChE activity have been reduced in digestive gland [48]. Study examined the Benzo[a] Pyrene (B[a]P) toxicity in

*Persistent Organic Pollutants in the Bizerte Lagoon Ecosystem: Occurrence, Distribution, and… DOI: http://dx.doi.org/10.5772/intechopen.100549*

*Mytilus galloprovincialis* and *Ruditapes decussatus* showed an increase of the total oxyradical scavenging capacity (TOSC) in the digestive gland after exposure to 100 and 300 μg/L concentrations of B[a]P. The superoxide dismutase (SOD), glutathione S-transferase (GST), and catalase (CAT) activities in gills and digestive gland were significantly induced. A significant increase of the GST activity and a decrease of AChE activity in digestive glands and gills were recorded in the two species. Interestingly, an increase of MDA level in the gills and digestive gland only in *Ruditapes decussatus* species [49].

Recently, Nasri et al. investigated the meiobenthic response to the polybrominated diphenyl ether (BDE-47) exposure in laboratory during one month [50]. BDE-47 caused a decrease in the bacterial abundance, and the nematodes taxonomic diversity as well as a change of all functional traits abundance, especially, the feeding group, amphid shape, and adult length were the most affected. In other study, Nasri et al. determined the taxonomic and trophic response of marine nematodes with the same concentrations of BDE-47 [51]. Species abundance and all univariate indices were significantly affected. BDE-47 treatment caused the microvores group represented by two species of *Terschellingia* to be replaced by the more resistant trophic groups such as epigrowth feeders (*Paracomesoma dubium*) and facultative predators (*Metoncholaimus pristiurus*).

#### **6. Conclusions**

The uses of chemical and biological analyses are promising tools for determining the quality of the Bizerte lagoon ecosystem and to survey the effects of the contaminants from anthropological activities. As a result, several studies have shown the effectiveness of the use in addition of sediments, associated organisms, such as marine bivalves, to monitor pollution [52]. They are specific indicators of different matrices in the ecosystem in relation to their position, and they show different rates of biotransformation and bioaccumulation compared to xenobiotics [53].

#### **Conflict of interest**

The authors declare no conflict of interest.

#### **Author details**

Ahmed Nasri1 \*, Takoua Mhadhbi1 , Mohamed Allouche1 , Amel Hannachi1 , Hary Demey2 , Patricia Aïssa<sup>2</sup> , Hamouda Beyrem1 and Ezzeddine Mahmoudi1

1 Faculty of Sciences of Bizerte (FSB), Laboratory of Environment Biomonitoring, University of Carthage, Zarzouna, Bizerta, Tunisia

2 Department of Chemical Engineering, ETSEIB, Universitat Politècnica de Catalunya, Barcelona, Spain

\*Address all correspondence to: a7mednas@gmail.com; ahmed.nasri@fsb.u-carthage.tn

© 2021 The Author(s). Licensee IntechOpen. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/ by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

*Persistent Organic Pollutants in the Bizerte Lagoon Ecosystem: Occurrence, Distribution, and… DOI: http://dx.doi.org/10.5772/intechopen.100549*

#### **References**

[1] Nasri A et al. Ethinylestradiol (EE2) residues from birth control pills impair nervous system development and swimming behavior of zebrafish larvae. Science of The Total Environment. 2021; 770, 145272.

[2] ANPE 1990. Preliminary diagnosis for the Study of the Ecological Balance of LakeDiagnostic préliminaire pour l'Etude de l'Equilibre Ecologique du lac de Bizerte. GIC-NNEA-TECI, ANPE, TunisiaTunisie

[3] Stegeman JJ. Cytochrome P450 gene diversity and function in marine animals: past, present, and future. Marine Environmental Research. 2000; 50, 61-62.

[4] Ameur WB et al. Polybrominated diphenyl ethers and their methoxylated analogs in mullet (*Mugil cephalus*) and sea bass (*Dicentrarchus labrax*) from Bizerte Lagoon, Tunisia. Marine environmental research. 2011; 72(5), 258-264.

[5] Ameur, WB et al. Organochlorine and organobromine compounds in a benthic fish (*Solea solea*) from Bizerte Lagoon (northern Tunisia): Implications for human exposure. Ecotoxicology and environmental safety, 2013; 88, 55-64.

[6] Ameur, WB et al. Legacy and Emerging Brominated Flame Retardants in Bizerte Lagoon Murex (*Hexaplex Trunculus*): Levels and Human Health Risk Assessment. Archives of environmental contamination and toxicology, 2020; 1-13.

[7] El Megdiche Y et al. Anthropogenic (PBDE) and naturally-produced (MeO-PBDE) brominated compound levels in Bizerte Lagoon clams (*Ruditapes decussatus*): Levels and human health risk assessment. Marine pollution bulletin, 2017; 125(1-2), 176-185.

[8] Mekni S et al. Occurrence of halogenated flame retardants in sediments and sea urchins (*Paracentrotus lividus*) from a North African Mediterranean coastal lagoon (Bizerte, Tunisia). Science of The Total Environment, 2019; 654, 1316-1325.

[9] Mekni S et al. Occurrence of halogenated and organophosphate flame retardants in sediments and eels (*Anguilla anguilla*) from Bizerte Lagoon, Tunisia. Frontiers in Environmental Science, 2020; 8, 67.

[10] Derouiche A et al. Polychlorinated biphenyls in sediments from Bizerte lagoon, Tunisia. Bulletin of environmental contamination and toxicology, 2004; 73(5), 810-817.

[11] Ameur WB et al. Concentration of polychlorinated biphenyls and organochlorine pesticides in mullet (*Mugil cephalus*) and sea bass (*Dicentrarchus labrax*) from Bizerte Lagoon (Northern Tunisia). Chemosphere, 2013; 90(9), 2372-2380.

[12] Barhoumi B et al. Distribution and ecological risk of polychlorinated biphenyls (PCBs) and organochlorine pesticides (OCPs) in surface sediments from the Bizerte lagoon, Tunisia. Environmental Science and Pollution Research, 2014; 21(10), 6290-6302.

[13] Trabelsi S, Driss MR. Polycyclic aromatic hydrocarbons in superficial coastal sediments from Bizerte Lagoon, Tunisia. Marine Pollution Bulletin, 2005; 50(3), 344-348.

[14] Barhoumi B et al. Polycyclic aromatic hydrocarbons (PAHs) in surface sediments from the Bizerte Lagoon, Tunisia: levels, sources, and toxicological significance. Environmental monitoring and assessment, 2014; 186(5), 2653-2669. [15] Salem FB et al. Distribution of organic contamination of sediments from Ichkeul Lake and Bizerte Lagoon, Tunisia. Marine pollution bulletin, 2017; 123(1-2), 329-338.

[16] Viarengo A et al. The use of biomarkers in biomonitoring: a 2-tier approach assessing the level of pollutant induced stress syndrome in sentinel organisms. Comparative Biochemistry and Physiology Part C Toxicology & Pharmacology. 2007; 146, 281-300.

[17] De Lafontaine Y et al. Biomarkers in zebra mussels (*Dreissena polymorpha*) for the assessment and monitoring of water quality of the St Lawrence river (Canada). Aquatic Toxicololgy. 2000; 50, 51-71.

[18] Mosleh YY et al. Effects of chitosan on oxidative stress and metallothioneins in aquatic worm *Tubifex tubifex* (Oligochaeta, Tubificidae). Chemosphere, 2007; 67, 167-175.

[19] Oliveira MM et al. Brain acetylcholinesterase as a marine pesticide biomarker using Brazilian fishes. Marine Environmental Research. 2007; 63, 303-312.

[20] Sunderman Jr. Biochemical indices of lipid peroxidation in occupational and environmental medicine. In: Foa, V., Emmett, E.A., Maroni, M., Colomi, A. (Eds.), Occupational and Environmental Chemical Hazards: Cellular and Biochemical Indices for Monitoring Toxicity. Wiley, Interscience, New York, pp. 1987; 151-158.

[21] Wong M, Poon B. Sources, fates and effects of persistent organic pollutants in China, with emphasis on the Pearl River Delta, in: H. Fiedler (Ed.), The Hand Book of Environmental Chemistry, Persistent Organic Pollutants, 3rd edn, Springer-Verlag, Berlin/Heidelberg, 2003, Chapter 13.

[22] Essid N, Aissa P. Quantitative study of free nematodes in the North and East sectors of the Bizerte lagoon (Tunisia). Etude quantitative des nématodes libres des secteurs Nord et Est de la lagune de Bizerte (Tunisie). 2002.

[23] Sakka Hlaili A et al. Winter-summer variation of the phytoplankton community of theVariation hivernoestivale de la communauté phytoplanctonique de la lagune de Bizerte lagoon in natural environments and fertilized with nutrients. Review of the Faculty ofen milieux naturel et fertilisé en nutriments. Revue de la Faculté des Sciences ofde Bizerte, Tunisia, 2003; 2, 37-49.

[24] Tolosa I, Mesa-Albernas M, Alonso-Hernandez CM. Organochlorine contamination PCBs, DDTs, HCB, HCHs in sediments from Cienfuegos bay, Cuba. Marine Pollution Bulletin, 2010; 60, 1619-1624.

[25] Montuori P et al. Polychlorinated biphenyls and organochlorine pesticides in Tiber River and estuary: occurrence, distribution and ecological risk. Science of the Total Environment, 2016, 571, 1001-1016.

[26] Aznar-Alemany Ò et al. Occurrence of halogenated flame retardants in commercial seafood species available in European markets. Food and Chemical Toxicology, 2017;104:35-47.

[27] Weijs L et al. Levels and profiles of persistent organic polluants in several tissues of harbor porpoises (*Phoconea phoconea*) from the Black Sea. Organohalogen Compound, 2009; 71:432-436

[28] Darnerud PO. Toxic effects of brominated flame retardants in man and wildlife. Environment International, 2003; 29, 841-853.

[29] Gill U et al. Polybrominated diphenyl ethers: human tissue levels and *Persistent Organic Pollutants in the Bizerte Lagoon Ecosystem: Occurrence, Distribution, and… DOI: http://dx.doi.org/10.5772/intechopen.100549*

toxicology. Reviews of Environmental Contamination and Toxicology, 2004. 183, 55e97.

[30] Cheikh M, Derouiche A, Driss MR. Determination byDétermination par (CPG ECD) of organochlorine pesticide residues in thedes résidus de pesticides organochlorés dans les sédiments de la lagune de Bizerte lagoon sediments.. Bulletin of thede l'Institut National Institute of Science and Technology of the Seades Sciences et Technologies de la Mer, 2002; 7, 160-163.

[31] Covaci A et al. Levels and distribution of organochlorine pesticides, polychlorinated biphenyls and polybrominated diphenyl ethers in sediments and biota from the Danube Delta, Romania. Environmental Pollution, 2006; 140, 136-149.

[32] de Mora S et al. Distribution of petroleum hydrocarbons and organochlorinated contaminants in marine biota and coastal sediments from the ROPME Sea Area during 2005. Marine Pollution Bulletin, 2010; 60, 2323-2349.

[33] Ennaceur S, Gandoura N, Driss MR. Distribution of polychlorinated biphenyls and organochlorine pesticides in human breast milk from various locations in Tunisia: levels of contamination, influencing factors, and infant risk assessment. Environmental Research, 2008; 108, 86-93.

[34] Ennaceur S, Driss MR. Serum organochlorine pesticide and polychlorinated biphenyl levels measured in delivering women from different locations in Tunisia. International Journal of Environmental Analytical Chemistry, 2010; 90, 821-828.

[35] Necibi M et al. Distributions of organochlorine pesticides and polychlorinated biphenyl in surface water from Bizerte Lagoon, Tunisia. Desalination and Water Treatment, 2015; 56(10), 2663-2671.

[36] Young LY, Cerniglia CE. Microbial Transformation and Degradation of Toxic Organic Chemicals. Wiley, New York. 1995.

[37] Gearing PJ et al. Partitioning of No. 2 fuel oil in controlled estuarine ecosystem, sediments and suspend particulate matter. Environmental Science and Technology, 1980; 14, 1129-1135.

[38] Chiou CT, McGroddy SE, Kile DE. Partition characteristics of polycyclic aromatic hydrocarbons on soils and sediments. Environmental Science and Technology, 1998; 32, 264-269.

[39] Karickhoff S. Semiempirical estimation of sorption of hydrophobic pollutants on natural sediments and soils. Chemosphere, 1981; 10, 833-846.

[40] Barhoumi B et al. Occurrence of polycyclic aromatic hydrocarbons (PAHs) in mussel (*Mytilus galloprovincialis*) and eel (*Anguilla anguilla*) from Bizerte lagoon, Tunisia, and associated human health risk assessment. Continental Shelf Research, 2016; 124, 104-116.

[41] Wang W et al. Depth-distribution, possible sources, and toxic risk assessment of organochlorine pesticides (OCPs) in different river sediment cores affected by urbanization and reclamation in a Chinese delta. Environmental Pollution, 2017; 230, 1062-1072.

[42] Xu X et al. Associations of serum concentrations of organochlorine pesticides with breast cancer and prostate cancer in U.S. adults. Environmental Health Perspectives, 2010; 118, 60-66.

[43] Necibi M, Mzoughi N. Distribution of organochlorine pesticides in sediment cores from the Bizerte Lagoon (Tunisia). International Journal of Environmental Analytical Chemistry, 2020; 100(10), 1118-1132.

[44] Louati H et al. The roles of biological interactions and pollutant contamination in shaping microbial benthic community structure. Chemosphere, 2013; 93(10), 2535-2546.

[45] Louati H et al. Responses of a free-living benthic marine nematode community to bioremediation of a PAH mixture. Environmental Science and Pollution Research, 2015; 22(20), 15307-15318.

[46] Othman HB et al. Structural and functional responses of coastal marine phytoplankton communities to PAH mixtures. Chemosphere, 2018; 209, 908-919.

[47] Sellami B et al. Effects of anthracene on filtration rates, antioxidant defense system, and redox proteomics in the Mediterranean clam *Ruditapes decussatus* (Mollusca: Bivalvia). Environmental Science and Pollution Research, 2015; 22(14), 10956-10968.

[48] Sellami B et al. The effects of anthracene on biochemical responses of Mediterranean mussels *Mytilus galloprovincialis*. Chemistry and Ecology, 2017; 33(4), 309-324.

[49] Dellali M et al. Multi-biomarker approach in *Mytilus galloprovincialis* and *Ruditapes decussatus* as a predictor of pelago-benthic responses after exposure to Benzo [a] Pyrene. Comparative Biochemistry and Physiology Part C: Toxicology & Pharmacology, 2021; 249, 109141.

[50] Nasri A et al. Using meiobenthic taxa, nematofauna biological traits, and bacterial abundance to assess the effects of the polybrominated diphenyl ethers compound: Case study of tetrabromo

diphenyl ether BDE-47. Science of The Total Environment, 2021; 770, 145251.

[51] Nasri A et al. Ecotoxicity of polybrominated diphenyl ether (BDE-47) on a meiobenthic community with special emphasis on nematodes: Taxonomic and trophic diversity assessment. Environmental Pollution, 2021; 277, 116727.

[52] Bodin N et al. PCB contamination in fish community from the Gironde estuary (France): blast from the past. Chemosphere, 2014; 98, 66-72.

[53] Scaps P. A review of the biology, ecology and potential use of the common ragworm *Hediste diversicolor* (O.F. Muller) (Annelida: *Polychaeta*). Hydrobiologia, 2002; 470,203-218.

Section 2
