The Morphophysiological, Histological, and Biochemical Response of Some Nontarget Organisms to the Stress Induced by the Pesticides in the Environment

Abstract

Ferns, amphibians, and fish are groups of nontarget organisms affected by many types of pesticides that end up in the environment. This chapter aims to approach the following themes: the influence of different pesticides on the spore germination process and on the differentiation of their gametophyte; aspects regarding the impact of some pesticides on breathing in fish (physiology and histopathology at the branchial level), as well as a series of effects at the hematological and biochemical levels; and changes of some hematological, biochemical, and structural parameters in amphibians. Species that are not directly targeted by the action of the pesticide in the environment, ferns can be used in their gametophyte stage, young or mature sporophyte in different biotests to evaluate the risk associated with these substances. The biochemical, hemathological, and histopathological changes recorded in both fish and amphibians can be considered biomarkers of pesticide pollution.

1. Introduction

Contamination of the natural environment, due to natural factors and human action, has been subject of numerous studies. In many ecosystems, human activity has led to changes of natural biogeochemical cycles, resulting in the accumulation of some substances.

Chemical contaminants are found everywhere in nature, and ecologists are the ones who assess their impact on natural communities of organisms. Among these, pesticides are the most common type of aquatic ecosystem contaminant.

Considering the physical and chemical properties (vapor expansion, volatility, evaporation capacity from water—codistillation) that transform pesticides, their bioaccumulative stability and capacity as well as the publication of ecotoxicological data on the action of different pesticides have become indispensable to continually monitor changes in the environment.

Different research on morphology, physiology, and biochemistry of gametophyte and sporophyte of fern species has shown that pesticides affect these processes, depending on species sensitivity, concentration, and exposure time.

Fish assimilates pesticides through gills or contaminated food. Gills are the main channel of pesticide penetration, which is why any disease at this level will have a great influence on the adaptive changes in the fish body. The effects of pesticides on fish are numerous and varied: they cause mortality both directly and indirectly, by starvation (destroying the organisms they feed on), affects hatching, growth rate, can lead to malformations, affects reproduction rate, modifies enzyme activity, and cause histopathological changes in organs, genetic effects, etc. Although under experimental conditions fish can survive pesticides in different concentrations, under natural conditions they are more vulnerable to disease, predators, no longer competitive, and no longer dealing with stress caused by changing seasonal temperature, reproduction season or temporary starvation.

Amphibians are organisms that populate aquatic ecosystems, being involved in aquatic trophic chains both by eating food and being food for predators. Over the last 20 years, scientists have reported the global decline of amphibian populations. Therefore, in 1996, the International Union for Conservation of Nature registered 156 species of amphibians in the Red List, and recent data show that 1856 species (about 32.5% of all amphibian species) are currently registered.

2. Influence of pesticides on spore germination and gametophyte differentiation in ferns

2.1 Gametophyte of ferns

Ferns and lycophytes (pteridophytes) represent 4% of Terra’s vascular plants [1], numbering about 11,000 species in 40 families and 300 genres [2]. Different species of pteridophytes are used as medicinal, food, horticultural, and agricultural plants [1] and in the last decade as organisms in acute [3] and chronic phytotoxicity tests [4].

The life cycle of pteridophytes is characterized by an alternation of generations between a well-developed sporophyte and a reduced gametophyte, independent of the sporophyte [1]. The gametophytic generation of ferns begins with the formation of spores (meiospores), from which, a gametophyte or multilcellular prothallium is differentiated by germination. The gametophyte of leptosporangiate ferns is above ground, photosynthetic, short lived, and heart shaped [1]. From the spore to the mature gametophyte stage with gamentagia, the differentiation of the gametophyte involves the passage through the prothallium filament and spatula-shaped prothallium lamella [5].

The spores germinate on the ground, the first division resulting in a smaller base cell (the initial cell of the rhizoid) and a larger apical cell (the initial cell of the prothallium). The initial cell of the rhizoid elongates and forms a hyaline rhizoid having no chloroplasts. The initial cell of the prothallium will form a multicellular prothallium filament, consisting of a variable number of cells with numerous chloroplasts. One or more rhizoids are formed on the cells at the base of the prothallium filament. Lamella formation (prothallium plate) is initiated by the longitudinal division of the end cell of the prothallium filament. One of the daughter cells divides through a slanted wall resulting in an apical (initial) wedge-shaped cell. The fore side of the prothallium then extends to a width of 3–4 cells and becomes spatula shaped. When the apex of the prothallium is heart shaped, the initial cell is replaced by a flat multicellular meristem. The activity of the lateral zones of this multicellular meristem helps the formation of the prothallium wings. The cells behind the apical meristem divide through a plan parallel to the underlayer, leading to the formation of a multi-layered medial crest. On the underside of the prothallium, there are antheridia, archegonia, and rhizoids. Prothallium differentiates different types of prothallium trichomes in some species.

The processes of spore germination and differentiation of gametophyte in ferns are influenced by a series of endogenous and exogenous factors, with environmental pesticides in the latter category.

2.2 Changes in the fern gametophyte caused by pesticides

Over the time, many researchers have highlighted the advantages of using spores and gametophytes of ferns in their experimental research [6] and ecotoxicology in recent years [3, 7]. The most advantages include: (1) the formation of a large number of spores/individual, (2) spores can be preserved in the laboratory and can thus be available throughout a year (3) are small and light, requiring reduced storage space, (4) spore germination and gametophyte differentiation may be obtained and monitored in vitro in small recipients, in simple laboratory conditions, and (5) conclusions are relevant for higher plants.

The spore germination process can be completely inhibited or delayed by pesticides in the environment. The glyphosate herbicide in concentrations of 0.48–19.20 mg/l significantly inhibited the macrospores germination process of Regnellidium diphyllum aquatic fern. From a total of 1050 megaspores used in the experiment, only 744 germinated [8].

A fungicide containing 50% metal copper applied in a concentration of 0.1–0.3 g/100 ml Knop solution significantly inhibited the spore germination process in Athyrium filix-femina and Polypodium vulgare ferns. The highest concentration of fungicide in the first species inhibited germination completely [9]. After 50 days of exposure to the fungicide, gametophyte differentiation was delayed compared to the control variant, which was in young cordate prothallium stage. Except for the lowest concentration of fungicide applied to P. vulgare in which the gametophyte was in the stage of prothallium lamella, at all the others, the stage of differentiation was prothallium filament. In addition, the filament cells had necroses, and the rhizoids were unelongated. Also P. vulgare species had more sinuous and more intensely colored cell walls at the level of rhizoids, [9] compared to the control variant.

A bifenthrin-based insecticide applied in a concentration of 0.01–0.04 ml/100 ml of Knop solution to the two above-mentioned species resulted in a decrease in spore germination rates, and at the highest concentration in both species, the gametophyte was in the form of a three-dimensional cell mass, whereas the gametophyte in the control variant was in cordate prothallium stage [9].

Asplenium scolopendrium spores which were grown in vitro on 20% copper metal fungicide containing culture medium showed a decrease in the germination percentage by up to 37.66% compared to the control variant. The species proved more sensitive than Athyrium filix-femina when the highest fungicide concentration was applied [10]. Gametophyte differentiation of both species was affected, so after 3 weeks of exposure, the gametophyte was in the stage of prothallium lamella formation for A. scolopendrium control variant, respectively, in the stage of young cordate prothallium for A. filix-femina control variant. The gametophyte was in the form of germinated spores when the highest concentration of fungicide (0.7%) was applied. Necrosis of prothallic chlorocytes, inhibition of rhizoid elongation, and even the absence of rhizoids have been observed in variants exposed to fungicide.

Acetamiprid-based insecticides (20%) dramatically affect spore germination in Asplenium scolopendrium and Athyrium filix-femina, the process being reduced by 63.34–100% in the first species, respectively, 41.34–100% in the second. In A. scolopendrium, in the variant exposed to 0.02% insecticide with acetamiprid, the gametophyte did not differentiate, the spores remaining ungerminated, while in A. filix-femina, after 6 weeks, the gametophyte was in the stage of three-dimensional cell masses and after 14 weeks in the cordate prothallium stage with archegonia [11].

The evaluation of glyphosate herbicide impact on Blechnum appendiculatum, Macrothelypteris torresiana, and Thelypteris dentata was made by applying it in concentrations of 0.33, 0.65, 2.72, and 10.89 g (active ingredient) l⁻¹ at different stages of the life cycle: spore, gametophyte, and sporophyte [12]. The results indicated almost complete inhibition of the spore germination process, discoloration of the prothallic chlorocytes, as well as the chloroplastic tissues of the young sporophyte. The authors state that the herbicide has a negative impact on spore banks in the soil, resulting in a mortality of 31–50% in all green stages of their life cycle [12].

Also, the fern sporophytes are affected by pesticides. The relative growth rate, the amount of chlorophyll pigments, and the photosynthetic activity of Azolla microphylla aquatic fern were also significantly affected by treatment based on endosulfan, an insecticide applied in a concentration of 0–600 ppm. The authors consider the photosynthetic activity and the amount of chlorophyll pigments in order to explain how pesticides act on the photosynthetic mechanism in Azolla [13].

Saturn herbicide in Azolla pinnata caused the decrease of nitrogen, phosphorus, and potassium uptake in concentrations between 0 and 0.004 ppm applied for 5–25 days, while furadan insecticide in concentrations of 0.001 and 0.002 ppm stimulated uptake of nitrogen, phosphorus, and potassium after 20 and 25 incubation days on the culture medium [14]. Saturn also affected the fresh and dry mass of Azolla pinnata.

The paraquat herbicide applied in the culture medium of Azolla microphylla, in a concentration of 2–6 μM, resulted in overproduction of reactive oxygen species (ROS) which determined the induction of antioxidant enzyme activity: superoxide dismutase, catalase, guaiacol peroxidase, and peroxidase ascorbate, the species tolerating herbicide toxicity for 72 h [15]. In higher concentrations, this antioxidant defense mechanism is no longer working, and Azolla does not survive. Fragmentation and browning of fronds were also observed. In addition to the abovementioned changes, the presence of paraquat in the medium determines the decrease of both chlorophyll and protein amount [15].

The mature sporophyte of Asplenium scolopendrium, Asplenium trichomanes, Athyrium filix-femina, Blechnum spicant, Dryopteris dilatata, Phegopteris connectilis, Polystichum aculeatum, and Woodsia ilvensis species subjected in situ to treatment with asulam herbicide showed important damage, the maximum being recorded 1 year after application [16].

3. Influence of pesticides on histological and biochemical parameters in fish

As a result of penetration into the aquatic environment, pesticides affect a wide range of nontarget organisms such as invertebrates and fish [17]. Aquatic ecosystems are the “final destination” of pesticides used in agriculture [18]. Early life stages are the most sensitive. Many toxicity tests are focused on the study of hatching and the occurrence of deformities in hatched larvae, decreased mobility or even immobility, lack of coordination in swimming movements; for example, deltamethrin in concentrations >0.005 ppb decreases hatching rate in common carp [19]; bifenthrin in concentrations of 50–200 ppb determines the lack of coordination in swimming movements, malformations of axial development in Danio rerio species [20].

Acute exposure to pesticide resulted in reduced fish populations and increased mortality [21, 22]. Chronic exposure to small amounts of pesticide increased the incidence of disease, stress, and behavioral disorders [23]. Pesticide bioaccumulation causes a major danger—bioaccumulation factor of cypermethrin in fish is 1200× [24]. Acute tests of toxicity are used to evaluate pesticide toxicity and get rapid consistent results at least within certain limits, regarding the concentration-response relationship. These tests allow the determination of LC50 values for different periods of time, NOEC values, providing a very rich database on the toxic effect of many pesticides on different fish species.

Considering the large amount of data in the specialized studies on the effects of pesticides on fish, the data in this paper refer in particular to the pesticide groups on which we have conducted our own experiments. Many authors have established concentrations or lethal doses for different pesticides and species. For example, the toxicological values for chlorpyrifos are between 1.3 μg l⁻¹ (96-h LC50) for Lepomis macrochirus and 2600 μg l⁻¹ (72-h LC50 72) for Gambusia affinis [25]; LC₅₀ in fish is less than 30 μg bifenthrin/l water [26, 27]; for most species, 96-h LC₅₀ values for bifenthrin determined by the static method are between 2 and 5 μg l⁻¹, and between 0.5 and 4 μg l⁻¹ by the continuous flow method [28]; propiconazole is moderate to slightly toxic for most aquatic organisms—96-h LC50 values for Cyprinus carpio 18.9 mg/l, Oncorhynchus mykiss 19 mg/l, and 21 mg/l for Lepomis macrochirus [29]; 48-h LC50 was 9 mg/l in Carassius auratus, 2.2 mg/l in Oncorhynchus mykiss, and 4 mg/l in Cyprinus carpio; the toxicity of chlorpyrifos in fish is generally between 0.01 and 1 mg/l (96-h LC50): 0.41 mg/l in Oncorhynchus mykiss and 0.015 mg/l in Salmo gairdneri [30].

As regards their action, pesticides block a certain metabolic process; the practical way in which this is done is sometimes difficult and in many cases is unknown or only partially clarified. The use of high sensitivity physiological indices is a tendency of recent years in aquatic toxicology. The most used physiological method is to determine pesticide action on fish breathing. Specialists can determine the concentrations in which breathing disorders begin to occur by recording the frequency of breathing movements and determining the oxygen consumption of fish kept in sublethal solutions of toxic substances (in state of rest or activity). The fish alters their energy metabolism in terms of spending a greater amount of energy to alleviate toxic stress [31], which results in improved oxygen use under hypoxia and anoxia conditions [32]. Determination and monitoring of oxygen consumption in aquatic organisms can be considered a better method to assess a substance toxicity than doing acute toxicological tests, because it has good results even in low concentrations of toxic substance [33]. One of the early symptoms of fish poisoning is breathing difficulties. Therefore, decreases in oxygen consumption in the gills were found by Cebrián et al. [34], even 24 h after exposure of Procambarus clarkii to chlorpyrifos, in the corresponding 96-h LC50 concentration. The frequency of opercular movements initially appears to support physiological activities in the polluted environment and is followed by the decrease in the breathing rate, which may be the result of gill diseases, as seen in the rainbow trout exposed to the action of fenvalerate and cypermethrin by Bradbury et al. [35].

Hematological parameters reflect fish state faster than determining other parameters because they change extremely rapidly under modified environmental conditions. For this reason, they are widely used to describe the state of fish health. Fish-nucleated red cells can be an ideal model to understand the harmful action of pesticides in different cell compartments; mitochondria and the nucleus of fish red blood cells are sensors in the programmed cell death mechanism [36].

Svobodova et al. reported significantly lower values of erythrocyte and hemoglobin in carp after acute exposure to deltamethrin (as a result of changes in hematopoiesis process), while the number of leukocytes did not change significantly [37]. Research carried out by Sopinska and Guz on permethrin poisoned carp revealed a decrease in the number of white blood cells, granulocytes in particular [38]. Anemia resulting from carp exposure to cypermethrin was also reported by Doruncu and Girgin [39].

Jayaprakash and Shettu evaluate the changes in some hematological parameters of the freshwater fish Channa punctatus exposed to sublethal concentrations of deltamethrin [40]. This study revealed: decrease in the hemoglobin content, total erythrocyte count, packed cell volume, mean corpuscular volume, mean corpuscular hemoglobin concentration, and a significant increase in the total leukocyte count.

Decrease in the number of erythrocytes in carp poisoned with pyrethroid insecticides (permethrin and cypermethrin), due to dysfunction of hematopoiesis [37, 39]. Exposure to Talstar EC 10 in a concentration of 57.5 μg l⁻¹ had no effect on the number of erytrocytes in common carp (Cyprinus carpio), with not significantly different values compared to those in the control group [41]. Atamanalp et al. [42] reported a significant increase in the number of erythrocytes in Oncorhynchus mykiss exposed to cypermethrin. Ahmad et al. observed significant decreases of hemoglobin in trout exposed to a sublethal dose of mancozeb, resulting in decreasing the amount of oxygen in the tissues and a decrease in the energy production of the animals [43]. Similarly, hemoglobin decreased percentually in Heteropterus fossilis after 30 days of exposure to deltamethrin [43].

Galloway and Handy studied the effect of organophosphorus pesticides (parathion, chlorpyrifos, malathion, and diazinon) on the immune system in invertebrates, fish, and other vertebrates [44]. Studies on fish have demonstrated an immunosuppressive effect to the action of these pesticides [45].

Physiological stress indicators, such as plasma cortisol and glucose level are commonly used to assess whether fish is stressed and whether the hypothalamo-pituitary-interrenal system works properly [46]. The intensive use of lactic acid in gill oxidation during stress caused by exposure to pesticides is mentioned in the literature [47] and is due to the increase in their energy demand [48] which is only covered by exogenous glucose. Glycogen mobilization is linked to the increase of energy demand; during pesticide-induced stress, a large amount of glycogen can be synthesized and stored in the muscles [49] which may later be in the form of glucose into blood tissues [50].

Due to the fact that fish live in a carbohydrate-deficient but protein-rich environment, they can use lipids and proteins as sources of energy more efficiently than carbohydrates [51]. Anaerobic glycolysis, recycling, and use of lactic acid preserve the energy potential of fish, thus ensuring better adaptation of fish to the polluted environment. Increases in the amount of energy (glucose) associated with lactate production have been reported in fish poisoned with pyrethroids [51].

High levels of lactic acid and strongly stimulated LDH activity in various tissues suggest the great importance of anaerobic glycolysis in adapting fish to stress caused by pyrethroid pesticides [52]. Velisek et al. examined the biochemical profile of carp (Cyprinus carpio) after 96 h exposure to bifenthrin in a concentration of 57.5 μg l⁻¹ and found significant increases in glucose level [41]. Bálint et al. observed an increase of blood glucose in carp after exposure to deltamethrin [53].

Velisek et al. showed an increase in blood glucose in trout and prussian carp as a result of metabolic stress induced by the action of Talstar 10 EC insecticide [41]. Jee et al. recorded an increase of blood glucose level in Sebastes schlegeli fish due to their exposure to cypermethrin (pyrethroid insecticide); researchers also reported a decrease in cholesterol and plasma proteins [54]. Datta and Kaviraj recorded increases in plasma glucose and decrease of liver glycogen in Clarias gariepinus exposed to a concentration of 0.005 mg cypermethrin/l [55]. Decrease in blood glucose levels following exposure to thyram was reported in Salmo gairdneri species by Van Leeuwen et al. [56].

Acetylcholinesterase is a highly used biomarker in aquatic ecotoxicology research [57], sensitive to low concentrations of organophosphates. Determination of acetylcholinesterase activity in erythrocytes is used to assess the degree of intoxication with acetylcholinesterase inhibitors. Recovery of acetylcholinesterase activity affected by exposure to organophosphorus pesticides can be carried out by dephosphorilation or AchE synthesis, which is a very slow process in fish; the disappearance of the inhibitory effect on acetylcholinesterase in Gambusia affinis exposed to oraganophosphorus compounds occurs within 45–60 days after exposure [58].

Costin et al. found a decrease of the specific activity of catalase, glutathione peroxidase, and glutathione reductase in prussian carps exposed to a sublethal concentration of deltamethrin (2 μg/l water), while glutathione-S-transferase activity increased [59]. Examination of the biochemical profile after 96 h of carp exposure to bifenthrin in a concentration of 57.5 μg l⁻¹ resulted in significant increases in glucose, ammonia, aspartate aminotransferase, creatine kinase, and monocytes [41].

Bifenthrin inhibits the production of ATPase [60], which explains the stronger effect of bifenthrin on aquatic organisms compared to terrestrial ones (maintaining the critical concentration of ions against the concentration gradient, in the much diluted aquatic environment, requires intensive ionic transport processes, the necessary energy being supplied by ATPase; the decrease in production of this enzyme leads to the death of organisms). Li et al., following the study of propiconazole action on Oncorhynchus mykiss, suggested Na/K-ATPase activity in the fish brain as potential biomarkers for intoxication [61].

Fish gills are the main place of ion exchange with the environment and, at the same time, the main channel of pesticide penetration. They are in constant contact with water, and any change in its composition may affect the gills’ permeability and their osmoregulatory functions. Cengiz found histopathological changes in the carp gills following acute exposure to deltamethrin in concentrations of 0.029 and 0.041 mg/l⁻¹ (exfoliation, necrosis, edema, hyperplasia, fusion of secondary lamellae, etc.) [62]; similar changes caused by deltamethrin action were identified in Gambusia affinis by Cengiz and Unlu [63]. Costin et al. found morphological changes in the carp gills exposed to sublethal concentrations of deltamethrin (2 μg/l water), 48 h after the exposure to 2 μg deltamethrin/l water, which accentuated after 14 days of exposure (longest interval); the author reported hyperemia, fusion of secondary lamellae, epithelial layer damage, and chlorogenic cell proliferation [59].

Fusion of secondary lamellae as a result of exposure to pesticides appears to have a protective role in diminishing the affected gill surface; this response slows down the penetration of toxic and may result in fish choking [64, 65]. The toxic effect of pyrethroids on fish increases with increasing liposolubility [66]; because of their strong lipophilic character, the pyrethroids are well absorbed by the gills, even in low-pesticide waters, thus decreasing the availability of oxygen to the tissues of the internal organs [67].

Histological changes due to exposure to pesticides also occur in other organs. Melo et al. observed hepatic tissue alterations in Rhamdia quelen (loss of cell form, piknotic nuclei, necrosis, and increase of hepatocytes loaded with by gallbladder pigment) following the exposure of fish to organophosphates (0.01 ml/l Folidol for 4–72 h) [68]. Examination of liver tissue samples after 96 h of carp exposure to bifenthrin in concentration of 57.5 μg l⁻¹ revealed hepatocyte degradation [41]. Cengiz and Unlu found histopathological changes caused by deltamethrin on the liver of Gambusia affinis [63]; cypermethrin causes hyperplasia and necrosis of hepatocytes in Labeo rohita species [69]. Ali et al. found DNA damage in erythrocytes and gill cells in Channa punctatus (Bloch), 5 days after exposure to chlorpyrifos in a concentration of 203 μg/l [70].

Much of the research on toxicological aspects of pesticide poisoning in fish refers to oxidative stress, causing the activity of antioxidant enzymes, proposed as biomarkers of pesticide intoxications [71, 72, 73]. It has been demonstrated that pyrethroid metabolism generates oxygen reactive species in poisoned fish [74]. The occurrence of oxidative stress in different fish tissues as a result of exposure to deltamethrin is supposed to be the main cause of product toxicity [75]. Chlorpyrifos toxicity is also manifested by the induction of oxidative stress [76]. It seems that oxidative stress contributes to the development and severity of syndromes in acute intoxication with these insecticides [77]. Mancozeb stimulated oxidative stress, while the amount of antioxidant serum enzymes decreased significantly; Fabra et al. demonstrated that this fungicide caused significant biochemical changes in cell membranes [78].

Exposure to pesticides may bring about behavioral changes such as: hyperactivity, loss of balance, and impossibility of maintaining the normal position in trout exposed to cypermethrin [79]. Symptoms of fish exposed to pyrethroid pesticides are: swimming near water surface, hyperactivity, loss of balance, increase of branchial mucus secretion, etc. Levine et al. found a decrease of swimming intensity in Danio rerio due to exposure to chlorpyrifos in a concentration of only 100 mg/l [80].

Toxicity of different pesticides on fish and other aquatic organisms can be influenced by various factors such as: pH, temperature, dissolved oxygen, water hardness, etc. Mauck et al. found that bifenthrin is more toxic at low temperatures, and its toxicity is only slightly influenced by water hardness [81]. Macek et al. recorded 96-h LC50 values of 7.1, 15, and 51 μg l⁻¹ in Oncorhynchus mykiss at temperatures of 12.7, 7.2, and 1.6°C [82].

Although the data on histochemical, biochemical, hematological, and behavioral changes in poisoned fish are quite numerous, they are little related to Carassius gibelio, Perca fluviatilis, and Alburnus alburnus species which are not commonly used in toxicological tests. In toxicology research, the emphasis is placed on biochemical, chemical, structural, and ultrastructural changes, disregarding the physiological processes which are part of the primary response of organisms to stressors.

As a contribution to the study of pesticide effect on fish, we will briefly present some of our results for these species exposed to the action of six pesticides—3 insecticides (Talstar One—in concentrations of 0.05, 0.1, 0.2 and 0.4 μl bifenthrin/l water, Reldan 40 EC—0.4, 0.8, 1.6, and 3.2 mg chlorpyrifos/l water, and Actara 25 WG—0.016, 0.032, 0.064 and 0.128 mg thiamethoxam/water) and 3 fungicides (Tilt 250 EC—1.06, 2.12, 4.24, and 3.2 mg chlorpyrifos/l water, Dithane M-45—2, 4, 8, and 16 mg mancozeb/l water, and Tiradin 70 PUS—0.007, 0.014, 0.028, and 0.56 mg thiram/l water).

The analyzed pesticides reduce energy metabolism and breathing rhythm in crucian carp, bass, and bleak [83, 84, 85, 86, 87, 88, 89, 90, 91, 92]. In many of the variants, the immediate response to the toxic action is a breathing rate stimulation of variable duration, in inverse ratio to the pesticide concentration, with high decrease in oxygen consumption, followed by the “stabilization phase” to the new conditions, with small fluctuations of the mentioned indices around an average value [85, 86, 87, 88, 89, 90, 91, 92]. The advantage of determining energy metabolism and breathing rate is the rapidity of this response to the action of the stressor (pesticide) on the one hand and their noninvasive character on the other.

The number of erythrocytes increases after exposure to Reldan 40 EC, Tilt 250 EC, and Tiradin 70 PUS [86, 87, 88, 89, 92]; decreases of this hematologic parameter are recorded under the action of Actara 25 WG and Talstar One [84, 90, 92]. Blood glucose levels in fish increase for 2 weeks under the action of Talstar One, Actara 25 WG, and Tilt 250 EC [84, 85, 86, 90, 92]. Dithane M-45 fungicide and Reldan 40 EC insecticide have a hypoglycemic effect [83, 86, 92]. Tiradin does not change blood glucose level in crucian carp and bleak intoxicated with a concentration of 0.02 ml/l for 2 weeks [91, 92].

Behavioral changes occur in crucian carp and perch exposed to pesticides: hyperactivity in the first 24–48 h, disordered movements, frequent rising at the surface of water in the first hours after exposure to toxic, and apathy under the action of insecticides and fungicides [85, 86, 87, 88, 89, 90, 91, 92]. Increases in mucus production occur especially in crucian carp under the action of fungicides [83, 87, 88, 89, 91, 92].

There are different results as regards the sensitivity of the species: the prussian carp was more sensitive under the action of Reldan 40 EC and Actara 25 WG [86, 92] as well as Tiradin 70 PUS and Tilt 250 EC fungicides [87, 88, 91, 92]; the bleak was more sensitive to Talstar One and Dithane M-45 [92]. In some experimental variants, there were no significant differences in the reactivity of the species under the action of the six pesticides.

Among the physiological parameters, blood glucose values and the number of red blood cells give a clear signal of fish stress, which is why we recommend using these parameters as biomarkers for pesticide-induced toxic stress. The study of oxygen consumption and breathing rate combined with the number of erythrocytes and blood glucose as well as the main behavioral changes allow the formation of a symptomatological picture of fish poisoned with pesticides.

4. Influence of pesticides on histological and biochemical parameters in amphibians

The impact of pollutants on humans in their environment is particularly complex and explains the lack of action or late and often confusing reactions regarding the protection measures. The degree of exposure depends on the simultaneous presence of some essential factors (nature and concentration of pollutants) and modifiers that ultimately influence the absorbed quantity.

Amphibians are some of the most representative species of vertebrates in aquatic and agricultural ecosystems, because they are the natural enemies of many pests of crop plants. Due to their high sensitivity to the changes occurring in their natural environment and because their larval development occurs exclusively in the aquatic environment, amphibians have been considered bioindicators for the aquatic and agricultural ecosystems [93, 94, 95, 96] and have been used as test specimens to study the action of different chemicals in these ecosystems.

Bridges [97] conducted numerous studies on the effects of long-term exposure to pesticides. The presence of contaminants in the environment may alter the predator-prey interaction among aquatic species by changing the predator or prey levels of activity or behavioral change of the predator. The effect of predator-prey meeting may be dependent on the fact that both species are exposed simultaneously to a contaminant/pollutant or exposure occurs only to one of the two species.

Due to the short period of exposure to nonpersistent pollutants, it is important to examine the long-term effects of short-term exposure on amphibians and the sensitivities at a certain stage in their development cycle. Any delay in metamorphosis or any decrease in size during metamorphosis may have an impact on the evolution of the amphibian populations, leading to its decline or local extinction.

Sampath et al. [98] studied the toxic effect of two pesticides (carbaryl and methyl-parathion) on the excretory system of Rana tigrina tadpoles and showed an increase in the rate of N-urea and NH3-N elimination in intoxicated specimens depending on the concentration of pesticide. Short-term exposure (96 h) to action in low doses of endosulfan organochlorine insecticide caused changes in the growth of Litoria freycinetti tadpoles either immediately or on a long term [99]. Tadpoles exposed to toxic action grew more slowly, and those who survived were more easily captured by predators in their natural environment.

Tomizawa and Casida [100, 101] demonstrated that neonicotinoid insecticides acted selectively on the nicotine-acetylcholine receptor in insects and mammals, making them the safest insecticides (as a mode of action). They described the mechanism of selective toxicity of neonicotinoid insecticides.

Gendron et al. [102] published the conclusions of a study on the effects of the leopard frogs’ exposure to a mixture of pesticides. The hypothesis shows that the exposure of Rana pipiens leopard frogs to agricultural pesticides can affect the dynamics of the parasitic worm infections, Rhabdias ranae. The pesticide treatment did not influence the growth of the parasites, the results indicating that they got matured and reproduced earlier in frogs exposed to pesticides, compared to control specimens. Such alternations in developmental cycle characteristics that increase the transmission of parasites may lead to an increase in virulence. The results contribute to further discussion on the role that anthropogenic factors might have in the gradual death of amphibians due to complications arising from the disease observed in different parts of the world.

Greulich and Pflugmacher [103] suggested a number of factors for the recent decrease in amphibian populations, one of them being exposure to pesticides. Specialists studied the absorption and effects of the pyrethroid cypermethrin insecticide in relevant concentrations from the environment for two different amphibian species, Bombina variegata and Rana arvalis, with the observation that cypermethrin absorption caused deformities, abnormal behaviour, and mortality. These changes depended on the dose of pesticide.

Some studies show that glyphosate has a harmful effect on the environment, especially on amphibians. Howe et al. [104] studied the effect of glyphosate on four North American amphibian species: Rana clamitans, R. pipiens, R. sylvatica, and Bufo americanus, and observed the following aspects: decrease of the anteroposterior diameter of the body, increased time for metamorphosis, tail malformations in tadpoles, and gonadal disorders. These effects may partly appear as a result of changes in the hormone level. There was a high level of transcription for ARNm segments encoding β-receptor synthesis for thyroid hormones, following exposure of individuals to solutions containing glyphosate.

Chen et al. [105] conducted studies on the effect of glyphosate on fauna in wetlands, having the zooplankton, Simocephalus vetulus, and Rana pipiens tadpoles as research subject. Significant effects of the pesticide action on the two species were determined at concentrations lower than those expected in the environment. Increased pH values (7.5) also showed the toxic effects of the pesticide on the two species under study, although the reproduction rate for S. vetulus improved to a pH level above 5.5 in the absence of the pesticide. The stress caused by the lack of food, associated with a pH of 5.5, decreased the survival rate of S. vetulus.

Relya [106] studied how four commercial compounds (diazinon, carbaryl, malathion, and glyphosate) affected the survival and larval growth rate of five species of amphibians (Rana pipiens, R. clamitans, R. catesbeiana, Bufo americanus, and Hyla versicolo). The combination of pesticides has occasionally resulted in lower survival rates and development than those determined by each pesticide, but never lower than those caused by the worse of the two. This suggests that the combination of the four pesticides had the same effect as the estimated total concentration of pesticides on the ecosystem.

The toxicity of pesticides in general and their genotoxicity on nontarget organisms in particular was of special interest to researchers. Shaolong et al. [107] studied the toxicity and genotoxicity of two pesticides (imidacloprid and RH-5849), used in China since 1992, for two species of amphibians. RH-5849 insecticide did not prove to be toxic to tadpoles even if they were kept for 96 h in a saturated solution. Two techniques were used to study the genotoxicity of the two insecticides: micronucleus test and single cell gel electrophoresis. These tests showed significant changes in the DNA of amphibian erythrocytes.

Relya [108] studied six North American species of amphibian larvae (Rana sylvatica, R. pipiens, R. clamitans, R. catesbeiana, Bufo americanus, and Hyla versicolor) to investigate the long-term effects of Roundup, many of the existing studies focusing on short-term effects (1–4 days). He also studied the effects of Roundup’s association with other sources of stress, such as predators, on the tadpoles’ survival rate during 16 days with and without chemical signals from predatory salamanders (Notophthalmus viridescens).

The values range from 0.55 to 2.52 mg of active ingredient (IA)/L, considerably lower than those used in previous studies. Stress increased by predators made Roundup two times more toxic to one of the six species (R. sylvatica). This finding suggests that the synergistic action of pesticides and predators may be a general phenomenon for amphibians (the range of pesticides can be very wide). Based on this research, pesticides such as Roundup certainly play a significant role in the decline of amphibian populations around the globe.

Relyea et al. [109] also studied the effect of pesticides (glyphosate and malation) on the natural environment of amphibians in the presence of zooplankton and phytoplankton (algae). Three species of amphibian larvae (Hyla versicolor, Bufo americanus, and Rana pipiens) were studied and combined with their predators (no predators with Notophthalmus viridescens and Dytiscus sp. Larvae) and pesticides (no pesticides, with malathion insecticide and Roundup herbicide). Roundup proved to have a negative effect on tadpoles, reducing their biomass by 40%, with no indirect effects on the amphibian community through predators or abundance of algae. Malathion in a concentration of 0.3 mg/l caused the number of tadpoles to decrease. This insecticide associated with the tritone in the amphibian environment does not have significant effects; the presence of beetle larvae associated with malathion has a positive effect, which determines the increase in the survival rate of the tadpoles in parallel with the decrease in the number of predators. While high concentrations of malathion can cause the death of amphibian larvae, the small concentrations may have positive effects as beetle larvae, and the predators of amphibian larvae are killed. These data lead us to the conclusion that pesticides can have direct and indirect effects on natural amphibian communities.

Toxicological investigations on amphibian larvae have mainly focused on the effects of heavy metals and pesticides on their growth, development, and behaviour, and only a few data refer to the bioaccumulation of toxins under natural conditions. Hofer et al. [110] studied the accumulation of inorganic (Pb and Cd) and organic (organochlorine and polyaromatic hydrocarbon pesticides) toxic in the body of Rana temporaria tadpoles at various altitudes in the Alps. They found an increase in the accumulation of Pb and Cd in the body of tadpoles in low pH waters, with a high concentration of metals. The amount of organochlorine substances was relatively low due to the age of the tadpoles (about 2 months) and the absence of lipid deposits necessary for the absorption of these substances.

Seifert and Stollberg [111] investigated the interaction of the neonicotinoid imidacloprid insecticide with the nicotine-acetylcholine receptor (nAcChR) in the frog embryonic muscle cells. The response of the muscle cell to the action of acetylcholine, nicotine, and imidacloprid was recorded as a contraction. The contractions of acetylcholine or nicotine are inhibited by α-bungarotoxin. Imidacloprid does not lead to the contraction of the muscle cell, but it can attenuate the contractions produced by acetylcholine or nicotine. They have found that imidacloprid is an antagonist of nAcChR receptor in the muscle cell and an agonist in its toxic action on insect nerve receptors.

Honda et al. [112] studied how to activate or inactivate 11 nicotinic insecticides in a two-step system of metabolic coupling and receptor binding. The authors incubated neonicotinoid insecticides with CYP3A4 and NADPH or AOX on a cosubstrate of N-methyl-nicotinamide for metabolism. They used ketoconazole or menadione for inhibition of subsequent conversions. They also used Drosophila nAChR receptor and [³H] imidacloprid or α4β2 and [³H](−)-nicotine receptor to determine changes in the action of neonicotinoid insecticides. Drosophila nAChR receptor coupled with CYP3A4 system activates imidacloprid and thiamethoxam, while the other nine insecticides do not undergo any change in their toxic potential. AOX system coupled with Drosophila nAChR receptor strongly inactivates clothianidin, dinotefuran, imidacloprid, desmethyl-thiamethoxam, and thiamethoxam has a low inactivation effect on nitenpiram and nithiazines but has no effect on the other four insecticides studied.

Neurotoxicity of pesticides has been studied by many researchers on the sciatic nerve of Rana ridibunda. Zafeiridou et al. [113] studied the influence of three herbicides: acetochlor, 2,4,5-trichlorophenoxyacetic acid (2,4,5-T), and 2,4-dichlorophenoxyacetic acid, on the value of the action potential of the sciatic nerve in Rana ridibunda. Since the action of 2,4-D is pH-dependent, the toxicity of the three chemical compounds has been studied at a pH of 3.3. It has been found that a low pH increases the toxicity of 2,4,5-T and 2,4-D pesticides and decreases acetochlor toxicity.

Casco et al. [114] demonstrated high toxicity of cypermethrin in Bufo arenarum larvae. They showed that low doses of cypermethrin, lower than those used in agriculture, can cause massive apoptosis in the central nervous system of the tadpoles through a neurotoxic mechanism.

In 2006, Relya [115] published an article on the effects of pesticides, pH, and stress due to predation on amphibians. It was shown that pH and stress due to predation and a single application of the insecticide (carbaryl) affected the survival and growth of Rana catesbein larvae and Rana clamitans green frogs under natural conditions. Decreasing the pH had no effect on the survival rate but caused a faster growth of tadpoles. Low concentrations of carbaryl did not affect the two species, but higher concentrations resulted in a lower survival rate and higher increases in individuals of Rana catesbein species. The stress caused by predation and low pH did not increase herbicide lethality due to its rapid rate of degradation under natural environmental conditions. Even if stress, pH, or predators could make carbaryl more lethal under laboratory conditions, after repeated insecticide applications, these stress factors did not interact under natural conditions even if a single carbaryl application was used.

The application of auxin herbicides of 2,4-dichlorophenoxyacetic acid (2,4-D) type in deciduous forests is mainly used in reafforested areas. The absorption of this herbicide in the target plants is reduced due to its low solubility in water. Because of this, the acid is esterified to increase its liposolubility and to diffuse into the plant-bearing vessels. But the esterification of auxin herbicides increases their absorption in other organisms, such as amphibians [116].

Numerous researches have shown that the gelatinous shell of amphibian embryos has a protective role against various environmental pollutants. Edginton et al. [117] studied the rate of absorption, metabolism, and excretion of butoxyethyl ester for 2,4-dichlorophenoxyacetic acid (2,4-D BEE) in Xenopus laevis embryos. They found that embryos with a gelatinous coating had a low absorption of toxic compared to embryos without the coating after 8 h from exposure. Metabolism of 2,4-D BEE lasted between 35 and 42 min, and the accumulation of radioactive residue was 35% located in the gelatinous shell and 65% in the embryo.

Vonesh and Buck [118] have shown that pesticides influence the process of laying eggs in amphibians. Four experimental batches were used to examine the effect of the pesticide marketed under the name of Sevin^® and the active substance, carbaryl, on the method of laying eggs in Hyla chrysoscelis species. Their results have shown that unpolluted ponds are preferred by amphibians for laying eggs, while the presence of Sevin^® and carbaryl in the environment reduces the process, while volatile chemicals have no influence at all.

Gurushankara et al. [119] investigated the effect of malathion in various concentrations on the survival, growth, and food consumption in Limnonectus limnocharis. Exposure of tadpoles to malathion changed all parameters. Thus, there was a decrease in survival rate, body size in parallel with the increase in malathion concentration in the environment as well as a reduction in the amount of food consumed.

Bioaccumulation of malathion and its impact on Ambystoma tigrinum was studied by Henson-Ramsey et al. [120]. The toxic was administered by two methods: soil contamination with malathion in a concentration of 50 and 100 μg/cm² and food—the administration of earthworms that lived on a malathion-contaminated soil in a concentration of 200 μg/cm².

The amount of xenobiotic substances was determined by gas chromatography in liver, muscle, skin, and bones. Malathion was higher in muscle, skin, and bones after 1 day of exposure and in viscera, after 2 days of exposure. To determine the biological impact of malathion on Ambystoma tigrinum, cholinesterase activity was measured in the brain, with a reduction in its activity. Animals exposed to toxic action did not show clinical signs of toxicosis.

Aquatic organisms are exposed to repeated pesticide applications over time. Repeated and long-term (79 days) administration of low concentrations of malathion (10–250 μg/l)—the most widely used insecticide, affects the entire aquatic community of zooplankton, phytoplankton, periphyton, and amphibian larvae [121]. This caused the decline of zooplankton followed by an increase in the biomass of the phytoplankton and the decline of periphyton after repeated treatments. The decline of the periphyton has had little influence on the development of Rana sylvatica species because it has a short time of metamorphosis, but has strongly influenced Rana pipiens larvae which have a longer time of metamorphosis with a high mortality level. In conclusion, malathion did not directly kill amphibian larvae but caused changes in the trophic chain and caused their mortality in an indirect way.

Toxicity caused by endosulfan (an organochlorine chemical compound used as an insecticide in vegetable, fruit, cotton, and coffee) was studied on Bombina orientalis, by Kang et al. [122]. They found a decrease in the survival rate of the tadpoles in parallel with an increase in the toxic concentration. The surviving tadpoles showed morphological changes consisting of tail dysplasia, presence of vesicles in the pectoral and ventral side, tail bending, and cephalic dysplasia; the higher the toxic concentration, the more significant the morphological changes.

Similar changes have also been reported in intoxication of larvae and Bombina orientalis tadpoles with molinate, a thiocarbamate used as an herbicide [123]. The change in survival rate was not significant until the blastula stage. A significant decrease of this parameter was observed in the tadpole stage even in low doses of molinate. The surviving tadpoles showed many malformations: trunk and tail bending, tail and eye dysplasia, and cephalic dysplasia. All these data suggest that the molinate herbicide influenced the larval development of this species.

Relyea and Jones [124] set LC50 after 96 h for glyphosate (Roundup) for various North American amphibian species. There were used nine species of larval anura Groser 25 belonging to Raniidae, Bufonidae, and Hylidae families, as well as four larval urodeles from two families: Salamandridae and Ambystomatidae. LC50 was between 0.8 and 2 mg acid/l for the nine anura species after 96 h and ranged from 2.7 to 3.2 mg acid/l for the four larval urodele. This work brings new data on the sensitivity of amphibians to the action of glyphosate.

Endosulfan insecticide appears to be highly toxic in low doses for amphibian populations and also exhibits long-term toxic effects. Toxicological studies conducted by Jones et al. [125] established LC50 in 4 days (LC504-d) for endosulfan insecticide on tadpoles of nine amphibian species belonging to three families (Bufonidae, Hylidae, and Ranidae). LC504-d value was between 1.3 and 120 ppb, which places it in the category of highly toxic pesticides. At the end of the treatment, experimental animals were kept in clean water for 4 more days. There has been a significant increase in mortality depending on the family. Bufonides were the least sensitive, hylides were moderately sensitive, and ranidae were the most sensitive.

Atrazine is a widely used herbicide for weed control in corn crops. Its residues could be detected in the soil 1 year after application. Numerous studies have tracked the effect of atrazine on embryonic development of amphibians in the environment and reported its ability to induce an enzyme—aromatase, which appears to convert testosterone to estrogen [126, 127, 128, 129, 130, 131, 132].

Some studies followed the effect of atrazine on juvenile and adult amphibians. Storrs et al. [133] followed the effect of atrazine administered directly on the soil on the behaviour of Bufo americanus species. They also studied various ways to absorb, accumulate, and eliminate this toxic from the body using ¹⁴C-labeled atrazine. Their results have shown that amphibians do not avoid atrazine-enriched soils. This pesticide is rapidly absorbed through the skin, and accumulated in the intestine, especially in the gallbladder.

Small quantities of herbicides used in agriculture often affect the surface of water and can become a stress factor for aquatic organisms. Williams and Semlitsch [134] studied the effect of 4 herbicides (the active substance atrazine, S-metolachlor and glyphosate) on the amphibian larvae of 3 species: Bufo americanus, Pseudacris triseriata and Hyla versicolor. The glyphosate under the trade name Roundup WeatherMax (containing 572 ppb equivalent glyphosate acid) determines the increase in mortality of Pseudacris triseriata tadpoles, while Roundup OriginalMax (containing 572 ppb equivalent glyphosate acid) determines the extension of the larval period in Bufo americanus species.

Chronic exposure to atrazine and S-metolachlor has no side effects on survival, metamorphosis, or larval period in the organisms. These results show the importance of adjuvants in the preparation of pesticides and their different effects on organisms even in the case of similar products.

Budischak et al. [135] followed how some parasites of amphibians can influence the toxicity of some pesticides. They used two batches of Rana palustris: tadpoles infected with Echinostoma trivolvis trematode, then intoxicated with malathion and uninfected tadpoles, and intoxicated with the same dose of malathion. It has been shown that the parasite does not influence susceptibility to pesticide action.

The influence of six pesticides (Reldan 40EC, Actara 25WG, Tilt 250EC, Champion 50WG, Fusilade Forte, and Dual Gold 960EC) on physiological indices of Pelophylax ridibundus was studied by Paunescu et al. [136]. The chronic pesticides exposure can lead to alteration of various indices (increased hepatosomatic index, decrease in erythrocytes number, leucopenia, hyperglycaemia, and decrease in cholesterol levels), as well as to hepatic lesion.

The fact that some species of amphibians survive in agroecosystems is due to their special plasticity. Thus, irrigation canals represent new habitats that amphibians colonize, and in the absence of predators, which are less mobile (fish), they become metapopulations of altered landscapes [137, 138, 139].

5. Conclusions

As nontarget species for pesticides in the environment, ferns can be used in their gametophyte stage, young or mature sporophyte in different biotests to evaluate the risk associated with these substances. The biochemical, hemathological, and histopathological changes recorded in both fish and amphibians can be considered biomarkers of pesticide pollution.

Acknowledgments

This work was supported by a grant of the Romanian Ministery of Research and Innovation, CCCDI-UEFISCDI, project number PN-III-P1-1.2-PCCDI-2017-0332”/Project 3 “Development of plant extracts and innovative phytosynthesized nanostructured mixtures with phytotherapeutical applications aimed at eliminating the biocenotic stress in horticultural crops”, within PNCDI III.