Fungicides

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This entry is adapted from the peer-reviewed paper 10.3390/su14074056

Fungicides are considered a vital tool for agriculture, protecting crops against fungal diseases and therefore securing high agricultural productivity. The evolution of technology can provide novel chemical fungicides, such as nanofungicides and chiral fungicides, to address plant resistance development. The extensive use of chemical fungicides leads to a risk for public health, natural waters, and non-target organisms.

constructed wetlands fungicides plant physiological responses microbiome

1. Environmental Fate and Risk of Chemical Fungicides

1.1. Inorganic Fungicides

The most known and used inorganic fungicides are copper and sulfur. Copper and sulfur are presented in numerous forms and are applied in various crops to control important foliar fruit diseases. Various copper and sulfur formulations for organic farming have been authorized by the European Union. Following their biogeochemical cycle, inorganic fungicides can run off to surface water and be adsorbed in the soil and sediments. Copper and sulfur are important trace metals for organisms’ fundamental functions, yet large concentrations can be harmful ^[1].

When copper oxide is dissolved in water, the dominant and biocidal oxidation ion is Cu²⁺ ^[2]. Acidity and salinity play an important role in aquatic biota toxicity. High pH values result in a reduction of available hydrogen ions, which leads to copper toxicity. Therefore, copper ions can be attached at aquatic organisms’ cells. Other physicochemical factors that influence the toxicity levels are the dissolved organic matter and dissolved water organic carbon content ^[1]^[3]. As Beck and Saundo-Wilhelmy ^[4] have reported, the tendency of sediments to adsorb heavy metals is high, and thus sediments can facilitate the availability of toxic chemicals in the water and aquatic organisms. Some species have a high level of sensitivity to copper, whereas others can efficiently overcome it. Copper is bioaccumulated in fish, decapod crustaceans, and algae and stored in bivalves, barnacles, and aquatic insects. The most sensitive species to copper exposure are cyanobacteria, while coccolithophores and dinoflagellates have a lower sensitivity to copper, and diatoms present resistance to copper ^[5].

Copper cannot be degraded in soil but can be accumulated through copper-based degradation compounds occurring in different forms. Copper’s mobility in the soil profile is characterized as medium to low. It was reported that the high concentrations of copper in vineyards soils and groundwater was caused mainly by copper-based fungicide use, negatively affecting water quality and food safety ^[6]. Copper residues in soil could cause toxic effects on macro- and microorganisms, adversely influencing the various beneficial interactions in soil, such as pesticide biodegradation, soil structure, nutrition availability for plants, and pathogen resistance ^[7]. Element copper is able to cause toxicity to beneficial bacteria and fungi in the environment ^[8]. Diaz-Ravina et al. ^[9] reported that the microbiocidal activity in vineyard soil can be significantly reduced by high application rates and prolonged use of copper-based pesticides. For instance, high-dose application of Cu-based pesticides can have negative effects on arbuscular mycorrhiza fungi (AMF) ^[10]. Schoffer et al. ^[11] reported that copper soil pollution is more common in countries or regions that have not enacted regulations for copper-based pesticides applications, which consequently follow only commercial formulation guidelines.

Sulfur can be found in various forms in the environment, such as gas (for example, SO₃) and salt (for example, MgSO₄), which are created through bacterial physiological processes. In sediments and soils, sulfur can be found as a trace element or in an inorganic form. Sulfur can cause toxicity to bacteria and fungi that are beneficial to the environment, which are not considered as crop pests. In addition, it was indicated that sulfur can be phytotoxic to some plants, such as cucurbits, apricots, and raspberries ^[12]^[13]. Kuklinska et al. ^[8] also reported that Vibrio fischeri is sensitive to sulfur exposure. The available information regarding sulfur interaction with organisms, its toxicity threshold, and its environmental fate is limited compared to copper.

1.2. Organic Fungicides

The environmental fate of organic fungicides depends on various physicochemical parameters, such as ionization (pk_a), water solubility, volatility, K_ow, and half-life in soil and water (DT₅₀). Soil texture, organic carbon content, pH, clay mineral type, dissolved organic matter, and cation exchange capacity also play an essential role in the environmental fate of fungicides, defining processes such as run-off to surface water, adsorption, or leaching. In addition, rainfall, irrigation, microbiological degradation, hydrolysis, photolysis, and application rate could affect fungicide fate ^[14]^[15].

Fungicide residues in surface water (for example, streams, lakes, rivers) and groundwater have been detected by many monitoring studies worldwide. The majority of these studies were focused on a few fungicides of local importance. The extensive (multiple applications and high doses) use of fungicides in specific crops (for example, vineyards, horticulture, orchards and so on) can lead to pollution of nearby natural waters from fungicide residues. Hence, the spatial and temporal distribution of fungicide residues in surface waters varies throughout the year and amongst agroecosystem compartments. Usually, the highest concentrations of curative fungicides are detected during growing or preharvest seasons, whereas preventative fungicides are found at early plant growth stages and during the winter period. Regions planted with grape and tree crops have received high application doses of fungicides, resulting in high detection frequency and high concentrations of fungicides in the ecosystems ^[16]^[17].

The presence of various pesticides has been investigated in vineyard groundwater bodies in northern Italy. The environmental quality standard set by the EU (0.1 μg/L) was exceeded by five fungicides (metalaxyl–M, fluopicolide, penconazole, tetraconazole, and dimetomorph), presenting significantly high concentrations. Τhe maximum concentrations of metalaxyl–M and penconazole were 8.015 μg/L and 18.72 μg/L, respectively ^[15]. In addition, a similar monitoring study was conducted in Spanish vineyards, where the detection frequency of metalaxyl, dimethomorph, and penconazole reached 50%. Moreover, the highest concentration was observed for same fungicides (metalaxyl and penconazole) ^[18]. These results indicate that the extensive use of fungicides in vineyards can cause an essential surface and groundwater pollution.

Papadakis et al. ^[14] conducted a pesticide monitoring study in two river basins in North Greece, with corn, cotton and cereals as the main crops, over a two-year period. Twenty-nine fungicides were detected at least once, while multiple detections (7 to 10 times) of boscalid, diphenylamine, etridiazole, and hexachlorobenzene were also observed. Extremely high concentrations for seven fungicides (azoxystrobin, diphenylamine, etridiazole, propiconazole, tebuconazole, quintozene, and difenoconazole), ranging from 0.153 to 0.819 μg/L, were identified. In the worst-case scenario, the risk quotient index was higher than one for four fungicides. It was showed that fungicides contribute to ecotoxicological risk for river basins.

The presence of 24 fungicides was investigated in the surface water and sediments of a horticulture area in Australia. The agricultural activity of the studied area included tree fruits, bulbs, vineyards, vegetables, and herbs. Although the authors reported that the individual fungicide residues did not pose environmental risks, due to low ecotoxicological endpoints, several fungicides were detected in concentrations above 0.2 μg/L (iprodione, myclobutanil, pyrimethanil, cyproconazole, trifloxystrobin, and fenarimol) and others had a detection frequency ranging between 18 and 36% (myclobutanil, trifloxystrobin, pyrimethanil, difenoconazole, and metalaxyl). The temporal distribution of residues was affected by the chemical class of the fungicides. Preventing fungicides were detected across the whole season, and curative fungicides mostly in March or October ^[19]. Although agricultural activity is the main source of pollution, urban and industrial activities can pollute the environment as well. Merel et al. ^[20] confirmed that the presence of carbendazim in the Rhine river (west Germany) originated from industrial wastewaters.

Recently, new ecotoxicological endpoints have been introduced for many fungicides due to their secondary side effects. The majority of studies for toxicological effects on non-target organisms have been conducted on a laboratory scale, using model organisms such as Lemna spp., Daphnia spp., and Dario spp. ^[16]. Dario rerio is an essential organism for toxicological studies, as Dario is sensitive to the exposure of toxic compounds. Endocrine dysfunction, oxidative stress, and immune system disorders were observed when zebrafish were exposed to carbendazim during larval and fetal stages at concentrations above 4000 ng/L ^[21]. Apart from carbendazim, tebuconazole caused adverse effects on the congenital system of Zebrafish, limiting locomotion at concentrations 4 and 6 mg/L ^[22]. The acute and chronic toxicity of strobilurins kresoxim-methyl, pyraclostrobin, and trifloxystrobin were investigated in D. magna neonates and embryos by Cui et al. ^[23]. The results showed that Daphnia embryos are more sensitive to fungicide exposure than neonates, presenting 157.3 µg/L, 3.9 µg/L, and 1.7 µg/L 48-h EC₅₀ for kresoxim-methyl, pyraclostrobin, and trifloxystrobin, respectively. In addition, the lowest-observed-effect concentrations were similar to the environmental concentrations, and thus the authors reported that the studied fungicides were very toxic for D. magna.

1.3. Chiral Fungicides

Many fungicides have an asymmetric center, which can provide two types of stereoisomers: enantiomers and diastereomers. Enantiomers have identical physicochemical properties but behave differently in asymmetric environments, such as in their biochemical processes. Enantiomers also show different biological activity, environmental fate, and toxicological profile. Diastereomers may not have identical physicochemical properties, and their biological activity usually varies ^[24].

Stereoselective fungicides differ in terms of toxicity, bioaccumulation, and bioavailability on non-target organisms ^[24]. Deng et al. ^[25] investigated the toxicity of four stereoisomers of metconazole to the aquatic algae Chorella pyrenoidosa. The results showed that the 96 h EC₅₀ values were different, following the pattern cis-1S,5R-Z > trans-1S,5S > trans-1R,5R > cis-1R. In addition, the photosynthesis dysfunction, the generation of reactive oxygen species (ROS), and the antioxidant response were induced more drastically by 1S,5S. In a similar study, three enantiomers of epoxiconazole were tested for their toxicological impact on the green alga Scenedesmus obliquus. The EC₅₀ values followed the order (+)-epoxiconazole > (−)-epoxiconazole > rac-epoxiconazole, whilst different effects on the determined chlorophyll contents, malondialdehyde contents, and antioxidant enzyme activities of algae cells were observed ^[26].

The environmental behavior of chiral fungicides has mainly been studied by evaluating their half-lives in crops and soils. The half-life values can provide interesting information about the potential biodegradation and the persistence of the studied compounds ^[27]. While fungicide enantiomers may present different half-life values from the racemic mixture, similar values were observed in other cases. The half-life of penconazole enantiomers in plant tissues and soil was determined in a field experiment. The results showed that the penconazole enantiomer (−) was degraded significantly faster than its (+) isomer in grapes and soil ^[28]. On the contrary, propiconazole stereoisomers were studied under aerobic, anaerobic, and sterile conditions by incubating the stereoisomers in three different types of soil, with the study investigating the dissipation process. The results showed no significant stereoselectivity under anaerobic and sterile conditions in all tested soil after 200 days of incubation, which is in contrast with the aerobic conditions where significant stereoselectivity was identified ^[29].

1.4. Nanofungicides

A fungicide is classed as nanofungicide if the size of the active ingredients ranges between 10 and 100 nm. The use of nanoparticles in fungi disease management can be divided in two categories: nanoparticles as protectants (alone) and as carriers for organic fungicides ^[30]. The main advantages of nanoparticles as carriers for organic fungicides are the improvement of persistence and activity of the active ingredient, the increased ability for translocation within plants, the confrontation of the low-water-solubility problems, and the achievement of slow release.

Preventative nanofungicides can be applied directly to roots, leaves or seeds. Copper, silver, zinc oxide, and titanium dioxide are amongst the most studied nanofungicides. The effectiveness of ZnO, Ag, CuO, and Cu nanoparticles was compared with a commercial formulation containing Cu(OH)₂ against common fungi strains such as Bacillus cinerea, Alternaria alternata and so on. The comparison evaluated mycelial growth, colony formation, seed germination, and the hyphal and spore morphology of the fungi. Mycelial growth of fungi strains was inhibited in vitro by all the nanoparticles, but the most effective were Cu and ZnO. In addition, the nanoparticles were more lethal at the spore germination stage ^[31]. Shyla et al. ^[32] investigated the antifungal activity of zinc oxide, silver, and titanium dioxide particles against Macrophomina phaseolina, which can infest oilseed and pulse crops. At lower concentrations, Ag nanoparticles were more effective than ZnO and TiO₂ against target fungi. In general, silver presented the highest antifungal activity from the other metals. Silver ions can cause dysfunction in thiol groups of fungal cell walls. As a result, the electron transport chain, energy metabolism, and transmembrane function are disrupted. Furthermore, fungal DNA can be mutated, respiratory chain dissociated, membrane permeability decreased, and cell lysis affected by silver-based fungicides.

Another popular nanoparticle fungicide with low toxicity risk to human health and the environment is chitosan. Chitosan can block nutrient supply, prevent the biosynthesis of mRNA and proteins, disrupt the cell membrane, and inhibit H+-ATPase of fungi. Some of the fungi that can be managed by chitosan are Fusarium crown rot, root rot in tomato, gray mold of grapes, and rice blast disease in rice ^[33]. Chitosan–lactide copolymer nanoparticles were used as carriers for pyraclostrobin, a low-water-soluble fungicide. The carrier was tested against Colletotrichum gossypii at different concentrations and compared to a commercial formulation of pyraclostrobin. The results showed that nanofungicide effectiveness did not exceed the commercial one at three and five days post-application. Nevertheless, the nanofungicide antifungal activity was increased at day 7 post-application ^[34]. Janatova et al. ^[35] achieved a successful Aspergillus niger control by formulating five individual essential oils with silica material MCM-41 in nanocapsules. Their effectiveness was reported to be higher than commercial oils at 14 days post-Aspergillus niger infection.

Despite the advantages of nanofungicides, these compounds can enter natural waters through leaching, run-off, or spray-drift. Soil properties such as surface charge, cation species, and the type of soil can define the mobility of nanoparticles in the soil. In addition, the nanoparticles can modify the soil sorption capacity of pesticides, resulting in the fluctuations of their toxicity severity. As a result, toxicity effects have been observed in humans, plants, microorganisms, and vertebrates due to their exposure to nanofungicides ^[36]. According to Ameen et al. ^[37], the exposure of nanopesticides can have a different impact on plant growth depending on application conditions such as application rate and size and type of nanomaterial. Nanoparticles can be taken up and cross the plasma membrane through various processes, such as endocytosis, pore formation, and carrier proteins ^[38]. Plants can activate defense systems and overcome stress parameters (including nanoparticles). However, there is the possibility for plants to fail to overcome toxicity effects, showing symptoms such as damaged DNA, reduced rate of transpiration, and others ^[39]. For instance, decreased content of leaf photosynthetic pigment (chlorophyll a and b) and reduced biomass (17–20%) have been observed in maize after application of the nanofungicide Cu(OH)₂ ^[40]. In addition, nanofungicides have the potential to harm beneficial soil bacteria and fungus. Abd-alla et al. ^[41] reported that high concentrations of Ag-nanoparticles reduced mycorrhizal colonization, glomalin content, and mycorrhizal responsiveness of AMF Glomus aggregatum. Hence, the nitrogen-fixing Azotobacter vinelandii presented various toxicity symptoms under Ag-nanoparticle presence, such as deduced cell number, apoptosis structural damage, inhibition of biological nitrogen fixation, and ROS generation ^[42].

Furthermore, various aquatic organisms have been investigated for their responses to nanoparticle exposure. Mortality, hatching delays, and various developmental malformations were shown when zebrafish embryos were exposed to nanoparticles ^[43].

1.5. Chemical Plant Defense Activators

Chemical plant defense activators constitute another group of fungicides with novel mode of action. Acibenzolar s-methyl (ASM) is a compound that can induce plants’ defense mechanisms, such as systemic acquired resistance, with salicylic acid taking part in the process. ASM can be used as an alternative solution to common bactericides and fungicides. Many researchers report that ASM is able to effectively manage various diseases in different crops, for example, onion and tomato ^[44]. In addition, ASM induces the production of enzymes and phytoalexins when the plant is chemically, biologically, or physically stressed. However, ASM has been identified as phytotoxic, has been linked to production losses, and in some situations, may exacerbate other pest assaults ^[45]. Potassium phosphate (PP) is a salt that is applied in cultures as a foliar fertilizer. Plant defense function is also stimulated by PP use. PP is characterized by high mobility and solubility. As a result, harmful oomycetes are successfully controlled in different cultures, such as papaya, tomato, and potato ^[45]^[46]. Another salt, fosetyl-Al, is a worldwide broad-spectrum fungicide that is commonly used in horticulture. Its action is based on preventing the development of fungi spores and mycelium, inhibiting the pathogen penetrating into the plant. It was also documented that fosetyl-Al plays a role in plant defense mechanism activation ^[47].

The environmental fate and toxicological effects of chemical plant activators to non-target organisms have not been extensively studied. The effects of fosetyl-Al on model species Danio rerio in water and Enchytraeus crypticus in soil were evaluated at a laboratory scale. The ecotoxicological assessment showed that fosetyl-Al is not considered safe for D. rerio and E. crypticus for concentrations higher than the PECs, which are 1.067 mg/kg for soil and 0.06496 mg/L for surface water ^[48]. In the case of ASM, Guziejewski et al. ^[49] reported that it is moderately to highly toxic to fish, moderately toxic to invertebrates. and highly toxic to aquatic plants.

2. The Role of Vegetation and Microbial Communities on Fungicide Removal in CWs and Other Phytoremediation Systems

Phytoremediation is a technology that uses plants and microorganisms located in the rhizosphere in order to remove, mitigate, break down, and retain pollutants such as pesticides in soils, surface water, and groundwater. However, phytoremediation processes can be adversely affected by pesticides, which can cause phytotoxicity, as plants are exposed to stressful conditions. CWs are a type of phytoremediation system is based on wetland plant species, which usually grow in soil or gravel substrate ^[50].

Among other factors, the vegetation and its related functions are crucial for pesticide removal in CW systems. Vegetation provides pollutant uptake, phytoaccumulation, degradation, and sorption through rhizosphere, flocculation, sedimentation, and suitable conditions for enhanced microbial activity ^[51] Therefore, an efficient phytoremediation process in CWs depends on two main factors: the tolerance of plants to pollutants and the presence of favorable conditions for microorganism growth in the rhizosphere, contributing to contaminant degradation ^[52]. In most studies, phytoaccumulation and plant uptake are associated with K_ow of each compound. High or low LogK_ow values may facilitate uptake or translocation, respectively. However, the optimum LogK_ow values for uptake, translocation, and accumulation range between 3.0 and 4.0 ^[53]. Furthermore, researchers ^[54] have reported that macrophytes such as Typha latifiola, Phragmitis australis, and others have high potential to absorb various pesticides, accumulating them in roots, stems, and leaves. In addition, the phytoavailability of a pesticide is determined by its molecular size (weight). A tiny molecule pesticide (Mr 500 Da) is often absorbed passively through diffusion, whereas a pesticide with a molecular weight more than 500 usually requires ATP hydrolysis to drive absorption ^[55].

Plants have established complex mechanisms for degrading xenobiotics like pesticides into detoxified compounds. Glycosyltransferase in plants catalyze the conjugation of sugars with endogenous metabolites or exogenous compounds. The purpose of glycosylation is to make substrates more water soluble, making them easier to degrade ^[55]. Many GTs s have been reported to protect plants from the toxicity of fungicides by O–glycosylation. The O–glycosylation products of their hydroxylated metabolites have been found in grape and strawberry for the fungicides thiabendazole, pyrimethanil, and cyprodinil ^[56].

Microbial diversity and richness have a dominant role in the stability and maintenance of CW treatment efficacy. Fungicide-degrading microorganisms (beneficial for plant growth) can be found either in filter bed material (porous media) or in plant roots. In addition, epiphytic or endophytic microorganisms can colonize wetland plants. The pesticide biotransformation, through microbial communities, is mainly conducted at the rhizosphere, which is favored by the release of oxygen and organic exudates of plants ^[57]. The combination of phytoremediation with bioaugmentation improves treatment efficiency compared to individual approaches. Therefore, researchers have recently focused on microcosmos functions and the potential amplification of microbiota activity ^[58].

The contribution of vegetation and the associated microorganisms to fungicide removal through CWs and other plant-based remediation systems is descried below, presenting various related studies.

2.1. Storbilurins

The action of strobilurins against fungal mitochondrial respiration is based on the electron transport block in the cytochrome bc₁ complex (complex III), between cytochrome b and cytochrome c₁, at the Q_o site. Therefore, ATP synthesis (energy supply) is blocked, causing oxidative stress and fungus cell death ^[59]. Pyraclostrobin (LogK_ow = 3.99) and azoxystrobin (LogK_ow = 2.50) are the best sellers among strobilurin fungicides and are mainly applied to soybean, grape, wheat, and corn crops. Other commonly used strobilurin fungicides are tryfloxystrobin, picoxystrobin, kresoxim-methyl, and fluaxastrobin, with log K_ow values of 4.5, 3.6, 3.4, and 2.86, respectively, and their water solubility inversely proportional to LogK_ow values ^[56]^[60].

Maillar et al. ^[61] indicated that a stormwater wetland achieved 100% removal of kresoxim methyl and 93% of azoxystrobin. P. Australis (70–80%) was the main wetland vegetation during the monitoring period. Other plants tested were T. latifiola and Choenoplectus lacustris. Although plant uptake was not quantified, vegetation was assumed to contribute to fungicide removal due to the optimum Log K_ow values of the tested compounds.

In addition, the phytoaccumulation ability of three plants (Juncus effusus, Pontederia cordata and Sagittaria latifolia) against azoxystrobin was assessed in order to be used in wetland remediation systems. Plants were planted in sandy soil, with azoxystrobin being applied to every two months. Researchers quantified the azoxystrobin residues in plants and soil during the experiment. P. cordata achieved the highest removal rate (51.7%). Particularly, the root of P. cordata contained up to 39.8% of the total absorbed azoxystrobin ^[62]. Furthermore, Elsaesser et al. ^[63] supported that the presence of the submerged macrophyte Elodea nuttallii (Planch) increased the removal efficiency of a vegetated flow-through stream system in comparison with an unplanted one, achieving more than 90% removal of the initial concentration of trifloxystrobin.

The complex between cytochrome b and cytochrome c₁, which is unsettled by strobilurins, appears in all eukaryotic organisms, including plants. Thus, the exposure of CWs plants to strobilurins may cause reduced plant growth and thus decreased phytoaccumulation ability. Pedersen et al. ^[64] tested the response of isolated mitochondria from wheat and showed that the electron transport in the cytochrome bc₁ was blocked.

2.2. Demethylation Inhibitors (DMI)

The DMI fungicide group consists of various chemical classes, such as triazoles and imidazoles, which have a common mode of action. DMIs inhibit the cytochrome P450, which is responsible for ergosterol production and the synthesis of the cell wall of fungi ^[65]. Difenoconazole, tebuconazole, imazalil, epoxiconazole, and cyproconazole are considered as commonly used members of the DMI group, and their respective LogK_ow values are 4.36, 4.1, 3.89, 3.3, and 3.09 ^[51].

Lv et al. ^[66] studied the removal efficiency of five different CWs planted with T. latifolia (cattail), P. australis (common reed), I. pseudacorus (yellow flag), J. effusus (soft rush), and B. erecta (water parsnip) against imazalil and tebuconzole. Planted CWs presented much higher removal efficiency than unplanted ones during the summer period. Despite the low phytoaccumulation of fungicides, the plants contributed to fungicide biodegradation, though the active microbial community hosted in their rhizosphere. Both fungicides were mainly phytoaccumulated by B. erecta, while J. effuses has shown significantly greater plant uptake in the winter. During the experiment, the CW vegetation accumulated only 5–6% of the initial applied fungicides (nominal concentrations). Another study indicated that the tebuconazole uptake by plants was quite low, ranging between 2.1 and 12.5%. However, the role of plant metabolism in tebuconazole dissipation was remarkable ^[58]. Furthermore, the role of three macrophytes in difenoconazole removal was investigated. Researchers suggested that the two main factors contributing to difenoconazole removal were the physicochemical properties of the tested compound and the biomass of the macrophytes ^[67].

2.3. Other Chemical Fungicides

Boscalid is a relatively new succinate dehydrogenase inhibitor (SDHI), with high effectiveness and a broad spectrum. Its mode of action is based on the inhibition of fungal respiration by blocking the ubiquinone binding site ^[68]. Papaenvagelou et al. ^[69] demonstrated that the moderately lipophilic boscalid was bioaccumulated by T. latifolia and P. australis, increasing the total removal efficacy of CWs. In addition, the lower uptake ability of plants during the winter period was attributed to a lower activity of plants when compared with the summer period.

The environmental fate studies of chloronitriles are focused on the fungicide chlorothalonil. Chlorothalonil can bind and deplete cellular glutathione (GSH) and can inhibit glycolysis by binding with glyceraldehyde 3-phosphate dehydrogenase (GAPDH), causing cell death ^[70]. One related to chlorothalonil removal using CWs showed that the fungicide removal was correlated with the presence of bacteria Pseudomonas spp., which were favored by P. australis ^[71].

As mentioned before, high copper residues in vineyard soil could have toxicological side effects. Hyperaccumulators can absorb high amounts of copper, especially in their shoots. The Brassicaceae family contains several plants able to accumulate significant amounts of copper. However, hyperaccumulators present low plant growth and thus low plant biomass, which is strongly correlated with copper plant uptake. Therefore, the choice of an appropriate plant is mainly based on the shoots’ ability to absorb high copper amounts ^[72].

Anilinopyrimidine fungicides might inhibit methionine biosynthesis. In addition, several researchers have shown that anilinopyrimidine fungicides can prevent the secretion of fungal hydrolytic enzymes such as laccases, lipases, and proteases ^[73]. Dosnon-Olette et al. ^[74] tested the potential toxicity effects of pyrimethanil on five macrophytes, with their removal performance also investigated. Among the tested species, L. minor achieved the higher phytoaccumulation of pyrimethanil (17.1%). Moreover, L. minor presented high tolerance to pyrimethanil exposure. A similar study was conducted using the macrophyte species L. minor and S. polyrhiza. The target compound was the morpholine fungicide dimethomorph. Both plants were not affected by fungicide exposure, with their removal rate reaching up to 18% ^[75].

Ag nanoparticles are possible carriers of fungicides, and thus their fate was investigated in CW systems. The effects of Ag nanoparticles on microbial communities in the CW substrates and the contribution to nanoparticle removal of CW vegetation were examined. The results showed that Ag nanoparticles caused changes in microbial community and also in the abundance of the key functional bacteria, decreasing the removal efficiency of the system. In addition, the plants of CW (Iris wilsonii) absorbed low amounts of Ag nanoparticles. However, the removal efficiency of Ag nanoparticles was evaluated at 95.72%, which indicates that the CW could effectively remove Ag nanoparticles ^[76]. Similarly, Cao et al. ^[77] investigated the removal performance of Ag nanoparticles though CWs. I. pseudacorus was planted in CWs and took part in the removal of Ag nanoparticles. Their residues were detected in the roots and leaves of plants. Roots could accumulate significantly higher amounts than leaves, with an overall removal performance higher than 98%.

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