Open access

Relationship Between Biomarkers and Pesticide Exposure in Fishes: A Review

Written By

Nédia de Castilhos Ghisi

Submitted: December 7th, 2011 Published: July 25th, 2012

DOI: 10.5772/48604

Chapter metrics overview

2,800 Chapter Downloads

View Full Metrics

1. Introduction

1.1. Human population growth and sustainability

In the modern world, small livelihood-based farms that grow multiple crops have been mostly replaced by large, agricultural conglomerates, in which food is grown in monocultures. This shift has allowed for the expansion of suburbs and new cities, but it has also made these places isolated, dirty, degraded, abandoned, and depersonalized. Owing in no small part to pesticide use, one of the legacies of the twentieth century is the potential for severe anthropogenic ecological damage. It has been recognized since the 1970s that byproducts and waste from technical-industrial development degrade the biosphere and threaten to irretrievably poison the environment to which we, as humans, also belong [1].

The term “sustainable” means that an activity can be continued and repeated indefinitely and predictably into the future. It is concerning that, in large part, human activities are logically unsustainable: global human population cannot continue to increase in size forever. We cannot continue to take fish out of sea faster than their populations can recover if we want to have fish to eat in the future. We cannot continue to develop crops in forests if soil quantity and quality deteriorates and water supplies become eutrophic or toxic. We cannot continue to use the same pesticides if an increasing number of pests and pathogens become resistant to them. We cannot maintain nature’s diversity if we continue to cause species extinction [2].

The source of many environmental problems, not to say of all, is simply our current level of rapid human population growth [Figure 1]. More people means more demand for energy, more consumption of non-renewable resources such as oil and minerals, more pressure on the renewable ones like forests and fisheries, more need for crops and food production, and so on. Surely, this cannot continue sustainably [2]. In addition to—or as a consequence of—this population growth, environmental pollution is also increasing.

Figure 1.

Representation of human population growth – before agricultural revolution [18th century] human population took around 1000 years to double in size. Currently, total population has been doubling in the last decade. Source: [2]

While world population [more than seven billion people] has doubled in the last decade, the demand for water supply increased by a factor of six according to United Nations data for the year 2000. Water demand grows faster than the population, mainly due to the introduction of more hygienic habits globally and the omnipresent desire to increase the productivity of food and industrialized products, the latter of which are generally also intended to maximize agricultural production. However, this increase in per-capita consumption exacerbates the imbalance between the distribution of water on Earth and its centers of human population density. Thus, the cheapest and most viable way to supply a growing world population is to learn how to use the available water in a more efficient manner [3].

1.2. Water sustainability as a resource

Water, when taken on the whole, can be considered an abundant resource on the planet. Total reserves sum up around of 1,265,000 trillion m3, distributed among solid [ice], liquid [rivers, lakes, oceans, water tables] and gaseous [atmosphere] phases [Figure 2].

However, out of the proportions shown in Figure 2, the relatively small parcel of water present in freshwater courses stands out. These waterways constitute the main source of water supply and are also the most common recipients of sewage discharge. Total world water demand is only around 11% of the mean discharge of rivers. Of this, 70% is used in agricultural activities, 20% in industrial activities and 10% goes to domestic and municipal use [3]. Therefore the actual crisis is not in terms of global water scarcity, but in its heterogeneous distribution. This is aggravated by the disorderly growth of local demand and, above all, by the fact that water degradation has reached unprecedented levels, not just in urban areas, but it rural areas as well [3].

In the same manner, water scarcity is not an issue that is exclusive to arid and semi-arid regions. Many areas with abundant water supplies, although insufficient to attend excessively high demands, have also experienced conflicts regarding water use and suffered consumption restrictions that have affected economic development and quality of life [4]. In addition, there has been severe extraction pressure on many aquifers to the point where many have been degraded nearly to extinction, especially near big cities or metropolises.

Figure 2.

Distribution of the world’s water. Source: Courtesy “Earth Update” CD-ROM, Rice University and the Houston Museum of Natural Science; used with permission.

The physical expansion of cities toward wellsprings has been causing serious damage, often focing their relocation.

The demand for a clean and safe water supply for human consumption, agriculture, and recreational purposes has been rising rapidly in the last few years. Water has become a limiting factor for agricultural, urban and industrial development. Recipient waterways, such as lakes, rivers and coastal areas receive great amounts of industrial, agricultural and urban waste directly via inputs and conveyance as well as indirectly through atmospheric deposition of aerial emissions. A complex mixture of toxic substances with an increasing number of contaminants has been deposited in these waters, posing a threat to aquatic ecosystems as much as to the health and well-being of human populations [5].

Regrettably, the waterways that are becoming contaminated are otherwise especially valuable resources. Unfortunately, contamination is very easy, but decontamination is often very costly and in some cases impossible to achieve [2]. In contrast with waste disposed in terrestrial environments that has more-or-less local effects, toxic waste in aquatic environments can be easily transported by currents and dispersed over large areas. Toxic chemicals in water, even in low amounts, can be concentrated to lethal levels by filtering aquatic organisms and top predators [6]. Pesticides, herbicides, oil waste and leakages, heavy metals [such as mercury, zinc and lead], detergents, and industrial waste can harm and kill organisms that live in or use this contaminated water. The potential risks of contamination in aquatic biota and humans can to be evaluated through biomonitoring programs. The relevance of these studies rises with the growth of urban, industrial, and agricultural activities around water sources [i.e., rivers, lakes and reservoirs] associated with frequently inadequate or insufficient water treatment.This is reflected in attempts to establish rules concerning the safety of water supplies in reservoirs and possible risks to environmental and human exposure [7]. These kinds of studies fits in a discipline called ecotoxicology.


2. Ecotoxicology

The term ecotoxicology was introduced by Truhaut in 1969 and was derived from the root words ecology and toxicology. The introduction of this term reflected a growing interest in the effects of chemicals in other, non-human species. Truhaut identified a field of study that was interested in the harmful effects of these substances within the concept of ecology. Ecotoxicology can be defined as the study of the harmful effects of chemicals on ecosystems, including effects on individuals as well as consequences in populations and higher levels of organization[8].

Despite the definition above, many of the first works in ecotoxicology had little to do with ecology or ecotoxicology. At the time, principal importance was placed on the detection and determination of pollutants in animal and plant samples, although the analytical results could sometimes be related to the effect on populations and communities. Analytical techniques such as chromatography, thin layer chromatography and atomic absorption facilitated the detection of very low concentrations of chemicals in living organisms, but establishing the biological significance of their presence or the organismal response to a specific dose of these substances remained difficult [8].

A substance is considered a pollutant when it is detected in levels above those that would normally occur in a particular environment. This immediately brings up the question “what level is considered to be normal?” For most synthetic organic chemicals, such as pesticides, the answer is simple: no detectable level is “normal,” because these substances do not exist in the environment until they are introduced by humans. On the other hand, substances such as metals, sulfur dioxide, nitrogen oxides, polycyclic aromatic hydrocarbons [PAHs] and methyl mercury naturally occur and their presence in the environment pre-dates humans. Naturally, there is variation in the concentration of these compounds across different sites and through time. This complicates the assessment of what is normal [8].

There is a conventional distinction between the definitions of the terms “pollutant” and “contaminant”: a pollutant is a substance that causes real environmental damage, while the term “contaminant” does not necessarily imply that the chemical is harmful. Still, it is difficult to deal with this distinction. First, there is the general toxicological principle that toxicity is related to dose. In this way, a pollutant can fit the description of pollutant in one situation [at high doses] but not in others [low concentrations]. Second, there is no general agreement on what constitutes environmental damage. Some scientists consider deleterious biochemical changes in organisms to be an environmental damage; others apply the term to population decrease. Third, the effects of the levels of chemicals measured in living organisms – or in their environment – are frequently unknown, even though the term pollutant is applied to them. This subject becomes even more complicated due to the possibility of toxicity increasing when organisms are exposed to a suite of environmental chemicals; such synergistic effects may result in chemicals having more a deleterious influence on organisms when in a mixture than they would otherwise in isolation [8].

Determining whether a contaminant is a pollutant also depends on its concentration in the environment, on the organisms to be considered, and on the possible damages of the contaminant to the organism. Thus, a compound can fit the description of a pollutant for one organism but not for another [8]. In order to minimize these problems of terminology, the term “pollutant” is used for environmental chemicals that exceed normal levels and cause damage. And environmental damage includes biochemical and physiological changes that adversely affect individual organisms, birth, growth and mortality rates[8], and reproduction.

An exciting aspect of ecotoxicology is that it represents an approach that extends from molecules to ecosystems, from genes to physiology [8]. This is further explored in the discussion of response levels in biomarkers.


3. Pesticides

When pesticides began to be developed on an industrial scale, manufacturers were not very concerned about the specificity of their products. These chemicals could damage anything, so long as they did not harm the crop, human beings, or their animals. A good example of this was when P. H. Müller received the Nobel Prize in Physiology or Medicine in 1948 “for his discovery of the high efficiency of Dichloride-Diphenil-Trichlorethane [DDT] as a contact poison against several arthropods” [9]. This insecticide was widely used after World War II to exterminate mosquitoes that caused malaria and typhus. It is cheap and very effective in the short run, but in the long run it is harmful to human health and possibly carcinogenic. It also interferes with animal life, leading to, for example, higher mortality in birds. In less than two decades DDT was banned in many countries, and today its use is forbidden almost worldwide.

The study of the impacts caused by pesticides gained attention from 1979 on, inspired by discoveries of pollution by nematicides in aquifers of many north-American states. Following this, many other cases of pesticide contamination of soil, water resources, animals and, more critically, human beings were diagnosed in the temperate regions, but little investigation was carried in tropical regions [10].

Many chemicals used to kill plagues have become important environmental pollutants. These pesticides are pulverized or released above plague areas, but only a small amount reaches the target, with most of it falling over resident crops or bare soil. Therefore, such pesticides are used in excessive quantities. This occurs especially with herbicides because they are cheaper than insecticides and fungicides [2]. The real problem emerges when the pesticide is toxic to species other than the target ones and, in particular, when they are transported outside the areas where they were applied and persist in the environment longer than expected.

Agricultural production is currently highly dependent on the use of pesticides, and the abandonment or reduction in their use would lead to a decrease in production, a rise in production costs, higher consumer prices, and, in some places, hunger and malnutrition[11].

The sale of pesticides involves billions of dollars a year. Current transgenic technology has increased the commercialization of certain kinds of pesticides, which can be deliberately used on resistant crops. With many of these substances reaching the environment, especially aquatic ecosystems where they are most concentrated, it is natural to think that some could be accumulating in individuals that compose food chains [12]. Due to their biological activity and to the huge quantity in which they are spread annually in the environment, pesticides can harm human health and the environment; for example, the induction of DNA damage can lead to adverse reproductive reactions, cancer and many other chronic diseases [13]. Recent evidence of negative effects of herbicides in amphibians, reptiles, fish and many other organisms continue to elucidate the fact that we are still discovering the extent to which populations can be affected by the current use of pesticides [14–18].

Therefore, the goal of this chapter is to discuss the methodologies and results of experiments and field surveys that analyze the effects of pesticides, primarily on aquatic communities and especially in fish.


4. Biomarkers

In ecotoxicology, there are many levels of response that can be evaluated.

The presence of a pesticide or other xenobiotic compounds in a portion of the aquatic environment does not, by itself, indicate a deleterious effect. Connections must be established between external levels of exposure, internal levels of tissue contamination and early adverse effects. The evaluation of these adverse effects—particularly if it is based upon only one level of response—can be affected by the ability of various pollutants [and their derivatives] to mutually affect toxicity, or even to act synergistically [19].

Deleterious effects on populations are often difficult to detect in feral organisms because many of these effects tend to be made manifest only after longer periods of time. When the effect finally becomes clear, the destructive process may have gone beyond the point where it could be reversed by remedial actions or risk reduction. In these sceneries the importance of early-warning signals, or biomarkers, that detect adverse biological responses towards anthropogenic environmental toxins become critical. A biomarker is any biological response to a chemical agent present in the environment that can be measured in the organism [or in its cells], in its metabolic products [urine, feces], or in hair, feathers, etc., that is indicative of some deviation in the standard pattern found in non affected organisms [20].

Pollutant stress generally triggers a cascade of biological responses, each of which may, in theory serve as a biomarker [21]. In the established reference response levels [22][Figure 3], biomarkers evaluate the precocious responses to pollutants. The responses in higher hierarchical levels are late response measures, frequently when the entire environment is already impacted. Biomarkers are important because they give us much more information on the biological effects of a certain pollutant than simply its quantification. Moreover, by the use of multiple biomarkers important information can be obtained. Biomarkers can be used after trophic, environmental, or occupational exposure, to elucidate the relation of cause-effect and dose-effect in health risk assessment, and in clinical diagnoses and for monitoring purposes [19].

Figure 3.

Schematic representation of the sequential order of responses to pollutant stress within a biological system. Modified from [22].

In developing a better understanding of the toxicity of contaminants, two kinds of studies can be carried out: bioassays, which are laboratory experiments, or biomonitoring, with direct field surveys. Although bioassays generate complementary data, it is important to note that experimental conditions do not always entirely reflect the natural environment [23].

On the other hand, we should not confound the term biomarker with bioindicator. A bioindicator is defined as an organism whose presence or absence, behavior, or some other characteristic gives information on the environmental conditions of its habitat. Fish species have attracted considerable interest as bioindicators in studies assessing the biological and biochemical responses to environmental contaminants. Fish can be found virtually everywhere in the aquatic environment and they play a major ecological role in the aquatic food-web because of their function as carriers of energy from lower to higher trophic levels [24]. Well like mammals, fish can suffer bioaccumulation, and have the advantage as bioindicators because they can respond to mutagenic agents in low concentrations by activating the P450 cytochrome enzymatic system, a system of monoxigenase enzymes with a heme group and different specificities per substrate. These enzymes play a fundamental role in the metabolism of xenobiotic substances and of endogen compounds [25].

Genetic, biochemical, and histopathological biomarkers are among the most common biomarkers in ecotoxicological studies with fish.

4.1. Genetic biomarkers

Genetic biomarkers evaluate the most precocious level of response: at the molecular level. DNA is a molecule that contains all the necessary information for the survival and perpetuation of an organism [26]. The exposure of an organism to genotoxic substances can lead to a sequence of events [27] that affect higher levels of response. Genetic ecotoxicology can be defined as the study of pollutant-induced changes in the genetic material of biota in nature and has two components: first, initially, the genotoxicity of pollutants, such as structural alterations in the DNA, and second, consequently, the procession and expression of DNA damage in mutant gene products, resulting in long-term heritable effects, such as changes in gene frequency within exposed populations, mutational events, etc. [28].

Many biomarkers have been used as tools for exposure detection and for the evaluation of the effects of genotoxic pollution. These biomarkers consist of tests such as the evaluation of chromosomal abnormalities, DNA adducts and breaks, the measurement of micronucleus frequency and other chromosomal anomalies, and the Comet Assay [29]. Here we will discuss surveys that used the Piscine Micronucleus Test [in conjunction with nuclear morphological alterations] and the Comet Assay.

Among many mutagenicity assays, piscine micronucleus and nuclear morphological alterations test [Figure 4] has been applied successfully because it is simple, safe, sensitive and it does not depend on the karyotypic characteristic of the study animal [30]. This last point is important because most fish have a relatively large number of small chromosomes, which are hard to visualize [31]. When fish erythrocytes are used there is also no excessive time consumption or animal suffering. Thus the micronucleus test in fish erythrocytes has been shown to be a promising technique in the investigation of environmentally-caused mutagenesis [32].

Micronuclei are small cytoplasmic chromatin masses present outside of the main nucleus of the cells that can originate from a chromosomal break as well as from a dysfunction in the mitotic spindle apparatus [33]. They are whole or partial chromosomes that were incorporated inside the nucleus of the daughter cell during cellular division and appear with a small dark round structure identical in appearance to the cellular nucleus [29]. Although there is a measurable basal level of spontaneous formation of micronuclei in most fish species [32], broad scale exposure to environmentally-relevant levels of clastogenic compounds in the laboratory [29,34,35] has been shown to elevate the frequency of micronuclei.

Figure 4.

Photos of a micronucleus and nuclear morphological alterations found in erythrocytes of neotropical fish Astyanax altiparanae. The first cell represents a normal nucleus, the second cell has a micronucleus and all other cells show nuclear morphological alterations.

In addition to the presence of micronuclei, nuclear morphological alterations can also occur, such as when the nucleus does not show a regular oval shape, but has a projection or an invagination of chromatin. In reference [36], they showed that these alterations are induced by well known genotoxic compounds, even when the micronucleus has not been formed. It is believed that these nuclear anomalies are due to problems with the nuclear lamina, because this structure confers the regular oval shape and stability on the nucleus [37].

Tests that directly evaluate breaks in DNA strands or chain alterations followed by DNA damage are commonly used to analyze the genotoxic impact in aquatic animals [38]. The Single Cell Gel Electrophoresis [SCGE], or Comet Assay, was firstly applied in ecotoxicology fifteen years ago, and has become one of the most popular tests for the detection of strand breaks in aquatic animals under in vitro, in vivo and in situ exposure[39].

The Comet Assay is a rapid, quantitative technique in which visual evidence of DNA damage in karyotic cells can be measured [Figure 5]. It is based on the quantification of denaturized DNA fragments that migrate out of the cell nucleus during electrophoresis. This method has been broadly used in many areas, including biomonitoring, genotoxicity, ecological monitoring, and also as a tool for DNA damage research or reparation in many kinds of cells in response to a variety of DNA-damaging agents [40].

There are many advantages to the Comet Assay: [a] genotoxic damage is detected at the individual cell level; [b] most eukaryotic organisms can be used in the Comet Assay; [c] a small number of cells is required; [d] it is usually easier to perform and more sensitive than other methods for the evaluation of strand breaks; [e] DNA strand breaks form quickly after a genotoxic exposure, so the essay provides an early evaluation of biota’s response [38].

The Comet Assay is usually done with erythrocytes because they are easily obtained through non-destructive methods and do not require the additional cellular isolation step. However, other tissues have also been tested for the genotoxic effects of contaminants because genotoxic effects can be tissue-specific [41].

Figure 5.

Pictures of five different damage rates in the Comet Essay through an immersion lens. a. zero damage; b. damage one; c. damage two; d. damage three; e. damage four [possibly in apoptosis]. Source: the author [2009].

4.2. Biochemical biomarkers

According to the Central Dogma of Molecular Biology [DNA mRNAprotein], DNA is indirectly responsible for protein production [26]. Therefore, DNA alterations can lead to damage in proteins [and the resulting enzymes]. These alterations can be quantified through biochemical biomarkers.

The most sensitive biomarkers are usually changes in the level and activity of biotransformation enzymes. Biotransformation is the conversion, catalyzed by enzymes, of a xenobiotic compound into a more water-soluble form, facilitating its excretion [19]. The enzymes responsible for biotransformation reactions are found throughout an organism [blood, kidneys, lungs, skin, nervous tissue, small intestine, and liver], but the liver is undoubtedly the organ in which they are most concentrated [42]. The biochemical biomarkers to be highlighted in this chapter are the activities of GST [Glutathione S-transferase], CAT [Catalase], lipoperoxidation [LPO], and Acetylcholinesterase [AchE].

The enzyme Glutathione S-transferase [GST] belongs to phase II of metabolism, and is responsible for the conjugation of electrophilic components or those that come from phase I with GST. The conjugation reaction started by GST is important to cells because it acts in the hydrolysis of lipophilic substances, which can then be excreted as inert substances in the organism. This super family of enzymes occurs in prokaryotes, plants, mollusks, crustaceans, insects, amphibians, reptiles, fish, and mammals [19].

Catalases are intra-cellular enzymes located in the peroxisomes that facilitate the removal of hydrogen peroxide, which is transformed into molecular oxygen and water [43]. Catalases are also cited as detoxication enzymes on some substrates, such as phenols, alcohols, formic acid and formaldehyde [44,45].

The lipid peroxidation or oxidation of polyunsaturated fatty acids is a regular physiological process that is important in cellular maturation [46–48] and lipid mobilization [49,50]. Some classes of contaminants, however, can have detrimental effects on this process [51,52] and can lead to damage in cellular function [52,53] and malfunction of cellular membranes and essential organelles, in addition to potentially affecting transporting processes, metabolites and ion gradient maintenance, and receptor-mediated signals transduction [54], with subsequent structural modification of the lipoproteinic complexes in cellular membranes [55,56].

Lipoperoxidation has been used successfully as a xenobiotic-induced oxidative stress measurement in organisms exposed in vivo to myriad chemicals such as metals [iron, cadmium, mercury and lead], paraquat, malathion, deltamethrin, and glyphosate [19,57–61].

The term cholinesterase [AChE] usually refers to the sum of the activities of pseudo-cholinesterase, or butirilcholinesterase, and acetylcholinesterase, or real cholinesterase, both of which are present in muscles [62]. The measurement of the AChE activity is often used to diagnose the exposure to anticholinesterasic toxins in fish, and can be considered one of the most ancient biomarkers [62,63]. Some authors [19] indicate that fish exposed to pesticides can show a reduction in acetylcholinesterase activity that is proportional to concentration and exposure time. The enzymatic measurement of cholinesterase allow the detection of sub-lethal toxicological effects, mainly of organophosphate compounds and carbamates, even without the presence of clinical symptoms.

4.3. Histophatological biomarkers

We can also observe damage in higher, cellular and tissue, response levels, which are detected through histopathological techniques. Morphological techniques such as light microscopy have been used in toxicology because they allow an evaluation of the possible effects of xenobiotics on target organs and tissues. According to [64], the effects in cell and tissue structure are important parameters to be considered in the evaluation of the potential toxicity of contaminants in living organisms.

Some authors [65] report that, through morphology, it is possible to reveal the most-affected target organs as well as to detect an organism’s sensitivity to the toxicity level of the compounds to which it was exposed. Histopathology also permits the differentiation of injuries promulgated by disease from those caused by environmental factors, such as the exposure to pollutants [66].

The advantage of histopathology as a biomarker lies in its use at intermediate levels of biological organization. Histological changes appear as a medium-term response to sub-lethal stressors, and histology provides a rapid method for detecting effects from xenobiotic compounds, especially chronic ones, in various tissues and organs [67]. For example, fish exposure to chemical contaminants is likely to induce a number of lesions in different organs [68,69]. Gills [70], kidneys [71,72], liver [73,74] and skin [75] are suitable organs for histological examination in order to determine the effect of pollution.

The article [67] propose an index of histopathological tests for any given organ, which leads to standardized quantification and allows legitimate comparison between different studies and, with restrictions, between different organs. This tool leads to a better understanding of the significance of histological findings after exposure to contamination.

Certain organs are the primary markers for aquatic pollution. For example, gills and skin have large surfaces that are in direct and permanent contact with potential irritants, and both have mucous cells. The liver plays a key role in metabolism and subsequent excretion of xenobiotics and is also the site of vitellogene production. The kidneys are very important for maintenance of a stable internal environment with respect to water and salt, excretion, and partially for the metabolism of xenobiotics [67].


5. Relationship between biomarkers and pesticide exposure in fishes

Every week, new articles are published showing the detrimental effects of many pesticides. These effects can be seen at all response levels: molecular, cellular, histological, individual, or even at higher ecological levels such as population, community, or ecosystem.

It is important to evaluate the effect of pesticides at lower response levels for the purposes of early damage detection, before they affect higher levels and decimate an entire community or ecosystem. Hence, we will focus on lower level responses at the molecular, cellular and histological level.

Mutagenic chemicals have a high probability of inducing carcinogenic effects in various fish species. A majority of these chemicals have been found to cause tumors at specific or multiple sites in fish [76]. Herbicides and pesticides comprise a large group of mutagenic chemicals, but information on herbicidal genotoxicity is lacking. Penthachlorophenol [PCP] and 2,4-dichlorophenoxyacetic acid [2,4-D] are chlorinated phenols widely used in agriculture. Chlorinated phenols in general are noted for exhibiting strong biological effects. For example, 2,4-dinitrophenol decouples oxidative phosphorylation; intervening in the oxidative pathways of metabolism. A clinical manifestation of this effect is the very rapid onset of rigor mortis in victims of pentachlorophenol poisoning [77]. Another study with humans revealed a significant increase in chromosomal abnormalities observed in the lymphocytes of workers exposed to PCP, leading to possible carcinogenic effects [78]. [79] surveyed the mutagenic effect of these two pesticides [PCP and 2,4-D] in the fish Channa punctatus. Using the Piscine Micronucleus Test [PMT], which evaluates the rate of permanent DNA damage, they observed that an increase in the dose and exposure time to both pesticides increased the rate of mutation in fish erythrocytes. In this work, they concluded that PCB was more toxic than 2,4-D in terms of Micronucleus induction.

The same fish species was used to evaluate the acute genotoxic effect of the insecticide Endosulfan [80]. Endosulfan is one of the most abundant organochlorine pesticides in the global atmosphere and is capable of undergoing long range transport to remote locations such as the Arctic [81]. Using the Comet Assay with liver and gill tissue, the authors observed a dose-dependent response; that is the higher the dose, the higher the rate of DNA breakdown in C. punctatus. In this case, there is evidence that gill cells are more sensitive than hepatic ones.

We reported on the case of an accidental spill of about 8000 liters of endosulfan in Paraıba do Sul River in the state of Rio de Janeiro State [Brazil], in November 2008 [7]. In this study, we analyzed the fish Pimelodus maculatus before [dry season] and after [rainy season] the spill in two affected drinking water reservoirs [Ilha dos Pombos and Santa Cecília] and in one that was unaffected [Santa Branca] These reservoirs are destined for human water provision. Fish from the Ilha dos Pombos reservoir [rainy season] that had been affected by the endosulfan spill showed several histopathological alterations in the gills and liver. Gill alteration index was similar in the fish from the three reservoirs in the dry season, but increased in the affected reservoirs during the rainy season, probably due to the endosulfan spill that occurred two months before this sampling. Figures 6 and 7 show the alterations found in the Paraíba do Sul River fishes. With regard to biochemical biomarkers, Cholinesterase activity in axial muscle was higher in P. maculatus from Ilha dos Pombos [33X] and Santa Branca reservoirs [11X] during the rainy season sampling after the endosulfan spill. Although we do not have normal values for acetylcholinesterase activity in this fish species, based on studies with other fish species the activity in the muscle showed a decrease in the dry season in all the reservoirs. Several pollutants such as organophosphates, carbamates [82], metals [83], hydrocarbons, and endosulfan [84] can decrease cholinesterase activity through inhibition or reduced expression, although an increase in activity has also been reported for fish muscle [85].

A similar study to [7] was also carried out in polluted and unpolluted areas in Estuarine Lakes at Santa Catarina Coast in southern Brazil, using the bioindicator fish Geophagus brasiliensis [86]. Estuaries are important sinks of pollutants derived from anthropogenic activities. The lakes in Santa Catarina are of great economic importance to the surrounding areas, enabling cultivation of rice crops and pig farming by various irrigation and drainage channels, as well as providing fish and shrimp to support ~10,000 artisanal fishermen. The results showed that both studied lakes are impacted by potential genotoxic substances. Severe lesions in the livers of G. brasiliensis were also observed. The inhibition of acetylcholinesterase activity suggested the presence of pesticides or metals in the study sites. The presence of large areas of rice crops around Santa Marta Lake [one of the studied lakes] may provide an explanation for the occurrence of substances with neurotoxic potential in the lakes. Three pesticides widely used on rice crops in southern Brazil [Clomazone, Quinclorac and Metsulfuron-methyl] have been shown to inhibit AChE activity in another species of fish, Rhamdia quelen [87]. The inhibition of AChE activity in fish can have adverse consequences for the animal itself, mainly by affecting its swimming ability and therefore its ability to find food and escape from predators [88].

Figure 6.

Histopathological findings from the liver of Pimelodus maculatus in Paraıba do Sul River. [A] Normal tissue. Arrows show vessels. [B] The presence of pancreatic tissue [large arrow] and the high incidence of melanomacrophage centers [small arrows]. [C] Occurrence of differentiated tissue [small arrows]. [D] Arrows show leukocyte infiltration. [E] Large necrosis area [arrows]. [F] Arrows show a large differentiated area of tissue. Scale bar=100 mm. Font: [7]

Figure 7.

Gills of Pimelodus maculatus from Paraíba do Sul River. A and B: Normal aspect of gills showing primary [small arrows] and secondary lamellas [large arrows]. Scale bar = 50 and 100μm respectively. C and D: arrows: fusion among secondary lamellas. Scale bar = 20 and 50μm respectively. E and F: neoplasia. Scale bar = 20 and 100μm. G and H: arrow= ectoparasite. Scale bar = 50 and 10μm. I: Observe the epithelial cells alterations [arrows]. Scale bar = 10μm. Font: [7].

Another field study compared two areas in southern California with records of chlorinated hydrocarbon [DDTs and PCBs] contamination to one less contaminated site. The frequency of micronuclei in circulating erythrocytes of two sea fishes was much higher in the contaminated areas. The DNA damage rate was up to four times lower in the uncontaminated site [89]. Organochlorine compounds such as in the DDT family, used as pesticides in agriculture, and polychlorinated biphenyls or PCB, which are important industrial chemicals and are used as non-flammable oils in many commercial products, are extremely persistent and difficult to degrade. Despite the fact that these compounds have been forbidden in many developed countries and their worldwide production and use have drastically decreased in recent years [90], at present they are widespread and have become ubiquitous contaminants of natural systems. PCBs are currently the most abundant chlorinated aromatic contaminants in the environment.

It was not until after DDT use had become widespread that the impacts of pesticides started to gain world’s attention and an environmental revolution began. This happened in 1962, with the release of the famous book Silent Spring, by Rachel Carson [91]. She described the process know as biomagnification, through which DDT and other organochlorine insecticides become more concentrated in higher levels of the food chain, being detected in the breast milk of women around the world and in the fatty tissues of Eskimos, inhabitants of isolated lands in Arctic. DDT is responsible for making bird’s egg shells thinner, particularly in birds of prey; this compound nearly drove the Peregrine falcon to extinction. DDT blocks calcium absorption, which makes the eggs easily broken and interrupts incubation, consequently undermining reproduction.

Currently, the pesticides with the highest sales rates worldwide are those based on glyphosate. Their sales have risen 20% a year, mainly due to the advent of biotechnology, which has provided plants that are resistant to this herbicide. Described by the manufacturers as pesticides low in toxicity and with good environmental compatibility, the glyphosate-based herbicides can seem like a silver bullet to those dealing with unwanted vegetation. However, there is public interest in the ecological, safety, and health concerns that may arise through the use of products from transgenic harvests [92].

There is some literature on the undesirable effects of glyphosate. Laboratory studies have detected adverse effects in every toxicological test category: medium-term toxicity [salivary gland lesions], long-term toxicity [inflammations of the mucous membranes of the stomach], genetic damage [human blood cells], reproductive effects [reduction in the number of spermatozoa in mice; higher frequency of abnormal spermatozoa in rabbits], and carcinogenicity [higher frequency of liver tumors in male mice and thyroid cancer in female mice][93].

The author [94] cites many positive results for the mutagenicity of glyphosate for a variety of test systems [e.g. Salmonella typhimurium – reverse mutation test, Drosophila melanogaster - induced sex-related lethal recessive mutations, and chromosomal aberrations in Allium cepa and cultures of human lymphocytes].

The most popular commercial product based on glyphosate is Roundup®. Its active ingredient is the 48% acid equivalent of the isopropylamine salt of N-[phosphonomethyl] glycine [C3H8NO5P; Monsanto Agricultural Co, St. Louis, MO, USA]. Roundup is a broad-spectrum, nonselective, postemergent herbicide that is used to kill unwanted plants in a wide variety of agricultural, lawn and garden, aquatic, and forestry situations. Despite its long and extensive use, the ecotoxicological data for Roundup are scarce.

A study by [95] evaluated the genotoxic potential of Roundup® in blood cells of the European eel [Anguilla anguilla]. In a bioassay, they subjected the fish to realistic exposure concentrations of 58 and 116 μg/L for 1-3 days, and also addressed the possible association with oxidative stress. Comet and erythrocytes’ nuclear abnormalities assays were used as genotoxic end points, reflecting different types of genetic damage. The authors showed higher rates of DNA damage in the contaminated fish than in the control group after 3 days of exposure [the same result was obtained in the Piscine MicronucleusTest]. The biochemical markers were assessed through enzymatic [catalase, glutathione-S-transferase, glutathione peroxidase and glutathione reductase] and non-enzymatic [total glutathione content] antioxidants, as well as by lipid peroxidation [LPO] measurements. Antioxidant defenses were unresponsive to Roundup. LPO levels increased only for the high concentration after the first day of exposure, indicating that oxidative stress in blood caused by this agrochemical was not severe. Overall results suggested that both DNA damaging effects induced by Roundup are not directly related with an increased pro-oxidant state.

Another study [96] showed different results. These authors evaluated the effects of Roundup Transorb®[RDT] on the Neotropical fish Prochilodus lineatus. Juvenile fish were acutely exposed [6, 24 and 96 h] to 1 mg/L of RDT, 5 mg/L of RDT, or only water [control]. They performed antioxidant analysis in the liver and acetylcholinesterase [AChE] determination in brain and muscle. After 6 h of exposure fish showed a transient reduction in superoxide dismutase and catalase activity. RDT also inhibited glutathione-S-transferase after 6 and 24 h of exposure. The reduction in these enzymes is probably related to the occurrence of lipid peroxidation [LPO] in fish exposed to the herbicide for 6 h. LPO returned to control levels after 24 and 96 h exposure to RDT, when fish showed an increased activity of glutathione peroxidase. The content of reduced glutathione also increased after 96 h exposure. Thus, after 24 and 96 h the antioxidant defenses were apparently enough to combat ROS, preventing the occurrence of oxidative damage. The exposure to RDT for 96 h led to an inhibition of AChE in brain and muscle at rates, which may not be considered a life-threatening situation.

The contradictory results of these studies warrant closer inspection. First, the concentration used in [96] was up to 86 times higher than the one used by [95]. Some studies show that biomarker responses are dose-dependent [19]. Second, the sensitivity of the fish must be taken into account. Not all fish have the same response to the same contaminant. The exposure time to the contaminant is also a factor that can be responsible for the differences in the results. Finally, the products used had different commercial names and different surfactants in their composition. Virtually all pesticides have other ingredients other than the active one, which actually has the exterminating action. Such ingredients are mistakenly called inert. Their purpose is to facilitate the use of the product or to make it more efficient. Usually the inert compounds are not identified in the pesticide’s label. In the case of glyphosate-based products, many “inert” ingredients were identified [93]. Differences in the test-organisms’ responses to glyphosate and to Roundup, its commercial formula, can be attributed to the toxicity of different compounds and surfactants in the commercial formula. Research has revealed that Roundup can be up to 30 times more toxic to fish than the pure glyphosate, due to the so-called inert compounds in the formula [94].

Some studies report pathological damage in fish exposed to glyphosate. The author [97] exposed Oreochromis niloticus to sub-lethal concentrations [5 and 15 mg/L] of Roundup for 3 months, and the organs exhibited varying degrees of histopathological change. In the gills, filament cell proliferation, lamellar cell hyperplasia, lamellar fusion, epithelial lifting, and aneurysm were observed. In the liver, vacuolation of hepatocytes and nuclear pyknosis occurred. Kidney lesions consisted of dilation of Bowman’s space and accumulation of hyaline droplets in the tubular epithelial cells. The results indicated that long-term exposure to glyphosate at sub-lethal concentrations had adverse effects stemming from histopathological and biochemical alterations in the fish. [98] has exposed Cyprinus carpio to immersion in Roundup [205 mg of glyphosate/L and 410 mg of glyphosate/L] in concentrations of 40 to 20-fold lower than those used in practice. Electron microscopy revealed that Roundup caused appearance of myelin-like structures in carp hepatocytes, swelling of mitochondria and disappearance of the internal mitochondrial membrane at both exposure doses. In this case, both studies, even though with different concentrations and species, confirmed that glyphosate can cause damages to fish tissues.

A study with the neotropical fish Corydoras paleatus contaminated with 3.20 μg/L glyphosate [6.67 µg/L Roundup®] showed that this pesticide might have genotoxic effects even at very small concentrations [99]. In this work, we performed PMT and Comet Assays with blood and liver cells, after the fish had been exposed to herbicide for 3, 6 and 9 d. A similar study [100], evaluated the sublethal effects of Roundup on the fish Astyanax sp. for 4 days. They tested two concentrations of Roundup: 3 µL/L and 6 µL/L. The PMT outcome was that only the highest dose showed any difference in response compared to the control group. Both works used the same commercial product, tested similar doses, and had similar responses, even though they were conducted with different fish.

The study [101] observed that Roundup® could affect cellular function [e.g., DNA] and that Roundup® and several glyphosate-based products interfered with cell-cycle regulation. In this work, the dose-response curves of the formulation products indicated a threshold for cell cycle induction even at very small concentrations, in agreement with other studies cited above. Failure in the cell cycle checkpoints leads to genomic instability and subsequent development of cancers from the affected cell [102,103]. Several lines of evidence have shown the highly conserved molecular basis of the cell cycle, from simple unicellular eucaryotes such as yeast to complex metazoans such as fishes or humans [104].

As discussed in the first pages of this chapter, a substance is considered harmful when it is detected in the environment at a higher concentration than it would normally occur. But what is the normal level for each substance? For many synthetic organic chemicals, such as pesticides, the answer is quite simple – no detectable level is normal because these compounds do not exist in the nature unless they are introduced by humans [8]. However, considering the current worldwide dependence on pesticides, it is impossible to avoid their entering natural environments, reaching animals, contaminating our food supplies and drinking water, etc. For this reason, countries try to establish a maximum tolerance limit for each pesticide in each component of the environment. One of the lowest limits is the one established by the European Union legislation, which is 0.10 μg/L [or 0.0001 mg/L] for all pesticides [individually] in water designated for human consumption [105]. Many studies have shown that this limit is safe [106]. In Brazil, Ministry of Health law 518 establishes the limits of some agrochemicals in drinking water, such as atrazine [0.002 mg/L or 2 µg/L], 2,4 D [0.03 mg/L], DDT [0.002 mg/L], Endosulfan [0.02 mg/L] and glyphosate [0.5 mg/L][107].

In 1974, the US Congress passed the Safe Drinking Water Act. This law requires the US Environmental Protection Agency [EPA] to determine the level of contaminants in drinking water at which no adverse health effects are likely to occur. These non-enforceable health goals, based solely on possible health risks and exposure over a lifetime with an adequate factor of safety, are called maximum contaminant level goals [MCLG]. Maximum contaminant levels [MCLs] are set as close to the health goals as possible, considering costs, benefits, and the ability of public water systems to detect and remove contaminants using suitable treatment technologies. The MCLG for glyphosate is 0.7 mg/L, or 700 ppb. EPA has set an enforceable MCL regulation for glyphosate at 0.7 mg/L, or 700 ppb. The MCLG for 2,4-D is 0.07 mg/L, or 70 ppb. For atrazine, the MCLG is 0.003mg/L; and for PCBs [Polychlorinated biphenyls] the MCLG is zero, and the MCL is 0.0005mg/L [108].

Canada has the Guidelines for Canadian Drinking Water Quality, which are intended to protect freshwater and marine life from anthropogenic stressors such as chemical inputs or changes in physical components. In this, the Maximum Acceptable Concentration [MAC] for atrazine and its metabolites is 0.005 mg/L; for 2,4-D is 0.1 mg/L, and for glyphosate the MAC is 0.28 mg/L [109].


6. Conclusions

In this chapter, we make explanations about some pesticides, and the effects of these on fishes, in field or laboratory assays. In addition to the pesticides cited above, many others are spread daily in the environment. However, little is known about the individual or synergistic effects that these products may have at the various levels of biological systems [in the short or long run]. Thus, many efforts have been made to explore to the deleterious effects of pesticides on non-target species, but there is still a lot to be done. These efforts are of great importance in understanding the impacts of pesticides in organisms and in the environment as well as in establishing of safe limits on the use of these products in the environment.



The author thanks the financial support of CAPES [PROEX], an entity of the Brazilian Government dedicated to the development of human resources.


  1. 1. orinE.Os Sete Saberes necessários à Educação do Futuro. São Paulo: editora Cortez; 2003
  2. 2. Towsend CR, Begon M, Harper JL. Fundamentos em Ecologia. 2nd ed. Porto Alegre: Artmed; 2006.
  3. 3. RebouçasA.daC.Água e desenvolvimento rural. Estudos Avançados (Internet). 2001154332744
  4. 4. SautchukC.FarinaH.HespanholI.OliveiraL. H.CostiL. O.IlhaM. S. O.GonçalvesO. M.MayS.BoniS. S. al.Conservação e reuso da água em edificações. São Paulo: Prol Editora Gráfica; 2005
  5. 5. PollackN.CunninghamA. R.RosenkranzH. S.Environmental persistence of chemicals and their carcinogenic risks to humans. Mutation Research. 20035288191
  6. 6. Primack RB, Rodrigues E. Biologia da Conservação. Londrina: Planta; 2001.
  7. 7. ArrudaA.FreireC.YamamotoF. Y.Silva deAssis. H. al.Monitoring water quality in reservoirs for human supply through multi-biomarker evaluation in tropical fish. Journal of environmental monitoring; 14261525
  8. 8. Walker CH., Hopkin SP., Sibly RM., Peakall DB. Principles of Ecotoxicology. 3rd ed. CRC Press; 2006.
  9. 9. (Internet). 2012 (cited 2011 Dec 20); Availablefrom:
  10. 10. RibeiroM. L.LourencettiC.PereiraS. Y.MaryR. M.Contaminação de águas subterrâneas por pesticidas: avaliação preliminar. Quimica Nova. 200730368894
  11. 11. Knutson RD.Economic impacts of reduced pesticide use in the united states: measurement of costs and benefits. Texas: 1999
  12. 12. Ferraro MVM.Avaliação do efeito mutagênico do tributilestanho (TBT) e do chumbo inorgânico (PbII) em Hoplias malabaricus (Pisces) através dos ensaios: Cometa, Micronúcleo e de Aberrações Cromossômicas. Universidade Federal do Paraná; 2003
  13. 13. Dimitrov BD, Gadeva PG, Benova DK, Bineva MV.Comparative genotoxicity of the herbicides Roundup, Stomp and Reglone in plant and mammalian test systems. Mutagenesis. 200621637582
  14. 14. Cavalcante DGSM, Martinez CBR, Sofia SH.Genotoxic effects of Roundup on the fish Prochilodus lineatus. Mutation research. 2008
  15. 15. Mohamed AH.Sublethal toxicity of Roundup to immunological and molecular aspects of Biomphalaria alexandrina to Schistosoma mansoni infection. Ecotoxicology and environmental safety. 2011May (cited 2011 Dec 13);74475460
  16. 16. PolettaG. L.Larrieraa.KleinsorgeE.MDMudryGenotoxicity of the herbicide formulation Roundup [glyphosate) in broad-snouted caiman (Caiman latirostris) evidenced by the Comet assay and the Micronucleus test. Mutation research (Internet). 2009Jan 31 (cited 2011 Sep 1);672295102
  17. 17. MañasF.PeraltaL.RavioloJ.OvandoH. al.Genotoxicity of glyphosate assessed by the comet assay and cytogenetic tests. Environmental toxicology and pharmacology. 2009Jul (cited 2012 Feb 28);2813741
  18. 18. Banks JE, Dick LK, Banks HT, Stark JD.Time-varying vital rates in ecotoxicology : Selective pesticides and aphid population dynamics. Ecological Modelling. 200821015560
  19. 19. van der OostR.BeyerJ.VermeulenN. P. E.Fish bioaccumulation and biomarkers in environmental risk assessment: a review. Environmental toxicology and pharmacology. 2003Feb;13257149
  20. 20. Van Gestel CA., Van Brummelen TC.Incorporation of the biomarker concept in ecotoxicology calls for a redefinition of terms. Ecotoxicology. 19965421725
  21. 21. Mc CarthyJ. F.HalbrookR. S.ShugartL. R.Conceptual Strategy for Design, Implementation, and Validation of a Biomarker-Based Biomonitoring Capability. Environmental Sciences Division; 1991
  22. 22. BayneB. L.BrownD. A.BurnsK.DixonD. R.IvanoviciA.LivingstoneD. al.The Effects of Stress and Pollution on Marine Animals. New York: Praeger; 1985
  23. 23. Ferraro MVM, Fenocchio AS, Mantovani MS, Ribeiro CDO, Cestari MM.Mutagenic effects of tributyltin and inorganic lead (Pb II) on the fish H. malabaricus as evaluated using the comet assay and the piscine micronucleus and chromosome aberration tests. Genetic and Molecular Biology. 20042711037
  24. 24. BeyerJ.SandvikbM.HyllandK.FjeldE.EgaasE.AasetE.alContaminant accumulation and biomarker responses in flounder ( Platichthys jlesus L.) and Atlantic cod (Gadus morhua L.) exposed by caging to polluted sediments in Sarrfjorden, Norway. Aquatic Toxicology. 1996367598
  25. 25. GoksøyrA.AnderssonT.BuhlerD. R.StegemanJ. J.WilliamsD. E.FörlinL.Immunochemical Cross-Reactivity of Beta-Naphthoflavone-Inducible Cytochrome 450P450ia) in Liver-Microsomes from Different Fish Species and Rat. Fish Physiology and Biochemistry. 1991
  26. 26. Griffiths AJF, Miller JH, Suzuki DT, Lewontin RC, Gelbart WM. Introdução à Genética. 7th ed. Rio de Janeiro: Guanabara Koogan; 2002.
  27. 27. ShugartL. R.BickhamJ.JackimG.Mc MahonG.RidleyW.SteinJ.etal. D. N. A.alterationsIn.HuggettR. J.KimerlyR. A.MehrleP. M.Jr BergmanH. L. .Eds.editorBiomarkers.BiochemicalPhysiological.HistologicalMarkers.ofAnthropogenic.StressChelsea: Lewis Publishers; 1992155210
  28. 28. ShugartL.Molecular markers to toxic agents. In: Newman MC, Jagoe CH, editors. Ecotoxicology: a Hierarchial Treatment. Boca Raton, USA: CRC Press; 1996133161
  29. 29. BombailV.AwD.GordonE.BattyJ.Application of the comet and micronucleus assays to butter ® sh ( Pholis gunnellus ) erythrocytes from the Firth of Forth, Scotland. Molecular Toxicology. 20014438392
  30. 30. MinissiS.CiccottiE.RizzoniM.Micronucleus test in erythrocytes of Barbus plebejus ( Teleostei, Pisces ) from two natural environments : a bioassay for the in situ detection of mutagens in freshwater Mignone river. Mutation Research. 199636724551
  31. 31. Hooftman RN, Vink GJ.Cytogenetic effects on the eastern mudminnow, Umbra pygmaea, exposed to ethyl methanesulfonate, benzo(a)pyrene, and river water. Ecotoxicology and environmental safety. 1981Sep;532619
  32. 32. Al-SabtiK.MetcalfeC. D.Fish micronuclei for assessing genotoxicity in water. Mutation research. 1995Jun;343(2-3):121-35.
  33. 33. HeddleJ. a. A.rapidin.vivotest.forchromosomal.damageMutation research. 1973May;18218790
  34. 34. Rodriguez-CeaA.AyllónF.Garcia-VazquezE.Micronucleus test in freshwater fish species: an evaluation of its sensitivity for application in field surveys. Ecotoxicology and Environmental Safety. 2003Nov (cited 2011 Dec 19);5634428
  35. 35. Grisolia CK, Starling FL.Micronuclei monitoring of fishes from Lake Paranoá, under influence of sewage treatment plant discharges. Mutation research. 2001Apr 5; 491(1-2):39-44.
  36. 36. AyllonF.Garcia-VazquezE.Induction of micronuclei and other nuclear abnormalities in European minnow Phoxinus phoxinus and mollie Poecilia latipinna: an assessment of the fish micronucleus test. Mutation research (Internet). 2000May 8;467217786
  37. 37. AlbertsB.BrayD.LewisJ.RaffM.RobertK.JDWatsonBiologia Molecular da Célula. 3rd ed. Porto Alegre: Artmed; 1997
  38. 38. FrenzilliG.NigroM.LyonsB. P.The Comet assay for the evaluation of genotoxic impact in aquatic environments. Mutation research. 2009cited 2012 Feb 29);68118092
  39. 39. OheT.WatanabeT.WakabayashiK.Mutagens in surface waters: a review. Mutation research. 2004Nov (cited 2011 Aug 23);567(2-3):109-49.
  40. 40. LiaoW.Mc NuttM. a.Zhu-GW.Thecomet.assaya.sensitivemethod.fordetecting. D. N. A.damagein.individualcells.Methods (San Diego, Calif.)2009May (cited 2012 Feb 29); 4814653
  41. 41. BelpaemeK.CooremanK.Kirsch-VoldersM.Development and validation of the in vivo alkaline comet assay for detecting genomic damage in marine flatfish. Mutation research. 1998Jul 31;415316784
  42. 42. Watkins PB.Drug metabolism by cytochromes 450in the liver and small bowel. Gastroenterology clinics of North America. 1992Sep;21(3):511-26.
  43. 43. StegemanJ.BrouwerM.Di GiulioR. T.Fö al.Molecular responses to environmental contamination: enzyme and protein systems as indicators of chemical exposure and effect. In: Huggett RJ, Kimerle RA, Mehrle PM, Bergman HL, editors. Biomarkers, Biochemical, Physiological, and Histological Markers of Anthropogenic Stress. USA: Lewis Publishers; 1992235335
  44. 44. AlbertsB.JohnsonA.LewisJ.RaffM.RobertsK.WalterP.Molecular Biology of The Cell. 4th ed. New York: Garland; 2002
  45. 45. NordbergJ.ArnérE. S. J.Reactiveoxygen.speciesantioxidants.themammalian.thioredoxinsystem.Free Radical Biology & Medicine. 200131111287312
  46. 46. ScheweT.RapoportS. M.KühnH.Enzymology and Physiology of Reticulocyte Lipoxygenase: Comparison with Other Lipoxygenases. Advances in Enzymology and Related Areas of Molecular Biology. 198658191272
  47. 47. MatsuiK.ShibataY.TatebaH.HatanakaA.KajiwaraT.Changes of Lipoxygenase and Fatty Acid Hydroperoxide Lyase Activities in Bell Pepper Fruits during Maturation. Bioscience, Biotechnology, and Biochemistry. 19916111199201
  48. 48. Van LeyenK.DuvoisinR. M.EngelhardtH.WiedmannM.Afunctionfor lipoxygenase inprogrammedorganelle degradation. Nature. 1998September):3925
  49. 49. FeussnerI.WasternackC.KindlH.KuhnH.Lipoxygenase-catalyzed oxygenation of storage lipids is implicated in lipid mobilization during germination. Proceeding of National Academy of Science of the United States of America. 1995December):1184953
  50. 50. FeussnerI.KühnH.WasternackC.Lipoxygenase-dependent degradation of storage lipids. Trends in Plant Science. 20016626873
  51. 51. BenzieI. F. F.Lipidperoxidation. A.reviewof.causesconsequences.measurementdietaryinfluences.International Journal of Food Sciences and Nutrition. 199647323361
  52. 52. SevanianA.UrsiniF.Lipid Peroxidation in Membranes and Low-Density Lipoproteins: Similarities and Diferences. Free Radical Biology & Medicine. 2000
  53. 53. KozarR. A.MckeoneB. J.PownallH. J.Free radical induced alterations in endothelial cell function. Journal of Surgical Research. 199456326
  54. 54. Meagher EA, FitzGerald GA.Indices of Peroxidation in Vivo: Strengths and Limitations. Free Radical Biology & Medicine. 20002812174550
  55. 55. Mason PR, Walter MF, Mason PE.Effect of oxidative stress on membrane structure: Small-Angle X-Ray diffraction analysis. Free Radical Biology & Medicine. 199723341925
  56. 56. Girotti AW. Serial Review: Regulatory and Cytoprotective Aspects of Lipid Hydroperoxide Metabolism.Free Radical Biology & Medicine. 2002
  57. 57. Kelly SA, Havrillal CM, Brady TC, Abramo KH, Levin ED.Oxidative Stress in Toxicology: Established Mammalian and Emerging Piscine Model Systems. Environmental Health Perspective. 1998106737584
  58. 58. SayeedI.ParvezS.PandeyS.Bin-hafeezB.HaqueR.RaisuddinS.Oxidative stress biomarkers of exposure to deltamethrin in freshwater fish, Channa punctatus Bloch. Ecotoxicology and Environmental Safety. 200356295301
  59. 59. ValavanidisA.VlahogianniT.DassenakisM.ScoullosM.Molecular biomarkers of oxidative stress in aquatic organisms in relation to toxic environmental pollutants. Ecotoxicology and Environmental Safety. 20066417889
  60. 60. HuculeciR.DinuD.StaicuA. C.MunteanuM. C.CostacheM.DinischiotuA.MalathionMalathion-Induced Alteration of the Antioxidant Defence System in Kidney, Gill, and Intestine of Carassius auratus gibelio. Environmental Toxicology. 20082452330
  61. 61. GlusczakL.MironS.BSMoraesSimões. R. R.SchetingerM. R. C.MorschV. al.Acute effects of glyphosate herbicide on metabolic and enzymatic parameters of silver catfish ( Rhamdia quelen ). Comparative Biochemistry and Physiology. 200714651924
  62. 62. Silva de Assis HC.Der Einsatz von Biomarkern zur summarischen Erfassung von Gewässerverschmutzungen. Technische Universität Berlin; 1998
  63. 63. SturmA.SilvaH. C.AssisD.HansenP.Cholinesterases of marine teleost fish : enzymological characterization and potential use in the monitoring of neurotoxic contamination. Marine Environmental Research. 19994738998
  64. 64. FentK.Ecotoxicology of Organotin Compounds. Critical Reviews in Toxicology. 19962613117
  65. 65. Wester PW, Canton JH.The usefulness of histopathology toxicity studies. Comparative Biochemistry and Physiology. 1991C(1/2):115-7.
  66. 66. SchwaigerJ.AdamS.PawertM.HonnenW.TriebskornR.The use of histopathological indicators to evaluate contaminant-related stress in fish. Journal of Aquatic Ecosystem Stress and Recovery. 199767586
  67. 67. BernetD.SchmidtH.MeierW.WahliT.Histopathologyin.fishproposal.fora.protocolto.assessaquatic.pollutionJournal of Fish Diseases. 1999222534
  68. 68. Sindermann CJ. Pollution- Associated Diseases and Abnormalities of Fish and Shellfish: a Review.Fisheries Bulletin. 197976471749
  69. 69. BuckeD.VethaakD.LangD.MellergraardS.Common diseases and parasites of fish in the North Atlantic: Training guide for identification. Copenhagen: Interanational council for the Exploration of the Sea; 1996
  70. 70. MallattJ.Fish Gills Structural Changes inducted by toxicants and other irritants: a statistical review. Canadian Journal of Fisheries and Aquatic Sciences. 198542463048
  71. 71. Oronsaye JA.Histological Changes in the Kidneys and Gills of the Stickleback, Gasterosteus aculeatus L, Exposed to Dissolved Cadmium in Hard Water. Ecotoxicology and Environmental Safety. 19891727990
  72. 72. BucherF.HoferR.The effects of treated domestic sewage on three organs (Gills, Kidney, Liver) of Brown trout (Salmo trutta). Water Research. 199327225561
  73. 73. Hinton DE, Laurén DJ.Liver structural alterations accompanying chronic toxicity in fishes: potential biomarkers of exposure. In: McCarthy JF, Shugart LR, editors. Biomarkers of environmental contamination. Boca Raton, USA: CRC Press; 19901757
  74. 74. Myers MS, Johnson LL, Olson OP, Stehr CM, Horness BH, Collier TK, et al.Toxicopathic Hepatic Lesions as Biomarkers of Chemical Contaminant Exposure and Effects in Marine Bottomfish Species from the Northeast and Pacific Coasts, USA. Marine Pollution Bulletin. 1998
  75. 75. VethaakD.The use of mesocosms to study disease in flounder (Platichthys flesus). In: Bylund G, Lönnström LG, editors. Diseases and parasites of flounder (Platichthys flesus) in the Baltic Sea. Turku: Baltic Marine Biologist Publication; 1994121129
  76. 76. Harshbarger JC, Clark JB.Epizootiology of neoplasms in bony fish of North America. The Science of the total environment. 1990May 1;94(1-2):1-32.
  77. 77. Seiler JP. Pentachlorophenol. Mutation research. 1991 Jan;257(1):27-47.
  78. 78. SchmidE.BauchingerM.DrespJ.Chromosome analyses of workers from a pentachlorophenol plant. Progress in clinical and biological research. 1982
  79. 79. MAFarahAteeq. B.AliM. N.AhmadW.Evaluation of genotoxicity of PCP and 2,4-D by micronucleus test in freshwater fish Channa punctatus. Ecotoxicology and Environmental Safety. 2003Jan (cited 2012 Mar 5); 541259
  80. 80. PandeyS.NagpureN. S.KumarR.SharmaS.SrivastavaS. K.MSVermaGenotoxicity evaluation of acute doses of endosulfan to freshwater teleost Channa punctatus (Bloch) by alkaline single-cell gel electrophoresis. Ecotoxicology and environmental safety. 2006Sep (cited 2011 Nov 19);6515661
  81. 81. WeberJ.HalsallC. al.Endosulfana.globalpesticide. a.reviewof.itsfate.inthe.environmentoccurrencein.theArctic.The Science of the total environment. 2010Jul 1 (cited 2011 Aug 12);40815296684
  82. 82. Oliveira Ribeiro CA, Silva de Assis HC.Recent Trends in the Acetylcholinesterase System, Acetylcholinesterase, Amsterdam: IOS Press; 2005
  83. 83. RabittoI. S.AlvesCosta. J. R. M.Silva deAssis. H. C.EEPelletierAkaishi. F. al.Effects of dietary Pb(II) and tributyltin on neotropical fish, Hoplias malabaricus: histopathological and biochemical findings. Ecotoxicology and environmental safety 2005Feb (cited 2011 Aug 23); 60214756
  84. 84. PayneJ. F.MathieuA.FanceyL. L.Acetylcholinesterasean.OldBiomarker.witha.NewFuture?.FieldTrials.inAssociation.withTwo.UrbanRivers.Papera.Millin.NewfoundlandScience. 199632222531
  85. 85. GillT. S.PandeJ.TewariH.Enzyme modulation by sublethal concentrations of aldicarb, phosphamidon, and endosulfan in fish tissues. Pesticide Biochemistry and Physiology. 1990Nov; 38323144
  86. 86. Benincá C, Ramsdorf W, Vicari T, de Oliveira Ribeiro C a, de Almeida MI, Silva de Assis HC, et al. Chronic genetic damages in Geophagus brasiliensis exposed to anthropic impact in Estuarine Lakes at Santa Catarina Coast-Southern of Brazil. Environmental monitoring and assessment. 2011May 15 (cited 2011 Oct 24
  87. 87. MironD.dosS.ShettingerM. R.MorschV. M.BaldisserottoB.MATiernoMoraes. al.Effects of the herbicides clomazone, quinclorac, and metsulfuron methyl on acetylcholinesterase activity in the silver catfish ( Rhamdia quelen ) (Heptapteridae). Ecotoxicology and Environmental Safety. 200561398403
  88. 88. BálintT.SzegletesT.SzegletesbZ.HalasyK.NemcsókJ.Biochemical and subcellular changes in carp exposed to the organophosphorus methidathion and the pyrethroid deltamethrin. Aquatic Toxicology. 19953327995
  89. 89. HoseJ. E.CrossJ. N.SmithS. G.DiehlD.Elevated Circulating Erythrocyte Micronuclei in Fishes from Contaminated Sites off Southern California. Marine Environmental Research. 19872216776
  90. 90. BrinkmanU. A. T.Productionproperties.usageof.polychlorinatedbiphenyls.In: Kimbrough RD, Jensen AA, editors. Halogenated biphenyls, terphenyls, naphthalenes, dibenzodioxins and related products. Amsterdam: Elsevier; 1989345
  91. 91. Carson R. Silent Spring. Houghton Mifflin Company; Anniversary edition (October 22, 2002); 1962.
  92. 92. Connera. J.JacobsJ. M.Genetic engineering of crops as potential source of genetic hazard in the human diet. Mutation research. 1999Jul 15;443(1-2):223-34.
  93. 93. CoxC.GlyphosateFactsheet.Journal of Pesticide Reform. 1998
  94. 94. Grisolia CK. Agrotóxicos: mutações, câncer e reprodução. Brasília: Editora Universidade de Brasília;2005
  95. 95. GuilhermeS.GaivãoI.SantosM. a.PachecoM.European eel (Anguilla anguilla) genotoxic and pro-oxidant responses following short-term exposure to Roundup--a glyphosate-based herbicide. Mutagenesis. 2010Sep (cited 2012 Feb 28); 25552330Available from:
  96. 96. Modesto KA, Martinez CBR.Effects of Roundup Transorb on fish: Hematology, antioxidant defenses and acetylcholinesterase activity. Chemosphere. 20108167817
  97. 97. JiraungkoorskulW.UpathamE. S.KruatrachueM.SahaphongS.PokethitiyookP.Biochemical and Histopathological Effects of Glyphosate Herbicide on Nile Tilapia (Oreochromis niloticus ). Environmental Toxicology. 200342607
  98. 98. SzarekJ.SiwickiA.AndrzejewskaA.Effects of the herbicide Roundup TM on the ultrastructural pattern of hepatocytes in carp (Cyprinus carpio). Marine Environmental Research. 2000502636
  99. 99. C.Avaliação do efeito Mutagênico do Herbicida Roundup ® em Bioensaio agudo com o bioindicador Corydoras paleatus (Pisces). Universidade Federal do Paraná; 2007
  100. 100. Rossi SC, Piancini LDS, Oliveira Ribeiro CA, Cestari MM, Silva de Assis HC.Sublethal Effects of Waterborne Herbicides in Tropical Freshwater Fish. Bulletin of environmental contamination and toxicology. 2011876037
  101. 101. MarcJ.Mulner-LorillonO.BelléR.Glyphosate-based pesticides affect cell cycle regulation. Biology of the cell. 2004Apr (cited 2012 Jan 5); 9632459
  102. 102. MolinariM.Cell cycle checkpoints and their inactivation in human cancer. Cell proliferation (Internet). 2000Oct;33526174
  103. 103. StewartZ. a.MDWestfallPietenpol. J. a.Cell-cycle dysregulation and anticancer therapy. Trends in Pharmacological Sciences. 2003Mar; 24313945
  104. 104. NurseP. A.longtwentieth.centuryof.thecell.cyclebeyondCell (Internet). 2000Jan 7; 1001718
  105. 105. Comunidade Econômica Européia. Directiva 80/778/CEE do Conselho, de 15 de Julho de 1980, relativa à qualidade das águas destinadas ao consumo humanoAvailable from:
  106. 106. RamsdorfC.FerraroW. AlmeidaM. V. M.RibeiroM. I. M.CAde CestariO.MMEvaluation of genotoxicity in Rhamdia quelen (Pisces, Siluriformes) after sub-chronic contamination with Fipronil. Environmental monitoring and assessment. 2011Sep;180(1-4):589-99.
  107. 107. Ministério da Saúde. Legislação Federal- Portaria MS2004
  108. 108. USEPA USEPA.National Primary Drinking Water Regulations. 2009Available from:
  109. 109. Federal-Provincial-Territorial Committee on Drinking Water, Federal-Provincial-Territorial Committee on Health and the Environment.Guidelines for Canadian Drinking Water Quality. Health [San Francisco). 2010

Written By

Nédia de Castilhos Ghisi

Submitted: December 7th, 2011 Published: July 25th, 2012