Effects Of Polychlorinated Biphenyls In Fish Biology Essay

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Polychlorinated biphenyls, PCBs, are organic, chlorinated chemicals that are synthetically produced from manufacturing companies. The high chemical properties and low reactivity of the compound allows the toxin to be useful in industrial applications such as coolant fluids and dielectrics. Although the toxin has been banned in the United States since 1976, its presence has been accumulated in the environment in areas such as air, water, sediments, and soils. PCBs are able to accumulate in lipid-rich tissues of organisms due to its chemical property allowing lipid solubility. The more chlorinated the compound, the higher the toxicity and isn't readily able to breakdown. PCBs are largely accumulated in fish that are located near manufacturing sites which leaked the toxins in water reservoirs. The question is whether the reproduction rates of the fish are affected and if the toxin is spread within fish species. The comparison is also made to rodents exposed to PCBs to see if similar effects can be observed between rodent and fish species. The results show the most PCB accumulation occurring in the liver for both species. Predator fish species show higher accumulation of the toxin. This answers the question as similar toxicity rates can been seen in two different species and the same organ, liver, is affected.


Polychlorinated biphenyl, known as PCB, is a synthetic, semi-volatile organic chlorine chemical molecule composed of two benzene rings with the chemical formula C12H10-xClx (Spaeth 2011). PCBs are considered a subcategory of Persistent Organic Pollutants, POPs. These toxins are unable to degrade in a chemical or physical state and are able to persist in the environment for decades (Spaeth 2011). The ability to be fat soluble allows the toxin to accumulate in the body at above average levels. Being semi-volatile allows the toxin to move to widely dispersed areas through wind and water currents, exposing the toxins the different species (Spaeth 2011).


Figure 1: Organic structure of Polychlorinated biphenyl. Chlorine may be placed on the numbers denoted on benzene rings; it can substitute any of the hydrogen molecules.

In a physical state, PCBs are not known to have a distinct smell or taste and may range from colorless to light yellow (Spaeth 2011). They may be found in three states: oily liquids, solids or vapor (Spaeth 2011). In commercial PCB mixtures, the lower chlorinated mixtures tend to be colorless, mobile and oil-like. High chlorinated mixtures are more likely immobile, viscous liquid and an amorphous solid. PCBs are thermally stable and have a long half-life of 8-15 years (Beyer, Bizuik 2009). There are 209 individual congeners of PCBs that can form depending on the number and location of the Chlorine atoms. Planer congeners contain two benzene rings in the same plane and are usually more toxic. Non-planer congeners have a 90 degree separation of the benzene rings (Spaeth 2011). Depending on the arrangement chlorine, the congener may act as estrogenic, anti- estrogenic, neurotoxic, or dioxin-like. This can lead to an abnormal production of estrogen, become toxic to neurons and become highly toxic with environmental pollutants (Spaeth 2011).

Polychlorinated biphenyl was manufactured in the United States from 1929 to 1976 (Beyer, Bizuik 2009). The widespread commercial use of PCBs unmerged due to its physical property transforming from liquid to solid state, overall chemical stability, inflammability and form homogenous solutions with organic solvents (Beyer, Bizuik 2009). PCBs were commercially used for dielectric and insulating coolant fluids in capacitors, electric motors and transformers, significantly in the production of fluorescent light fittings and electrical transformers (Spaeth 2011). Up until mid-1970s, PCBs were commonly used as plasticizers in cements, paints and rubber products in residential homes in the United States (Spaeth 2011). PCBs are able to enter the environment through the manufacturing site with the use and disposal of congeners (Spaeth 2011). Though PCBs have been banned, their presence in the environment can be seen via hazardous waste sites, improper and/or illegal disposal of wastes, leakages from old electrical transformers and the use of incinerators to burn wastes (Brazova, et al 2011). These contaminants are present in all aspects of the ecosystem such as air, fish, human adipose tissue, milk, serum, water and wildlife and human adipose tissues (Beyer, Bizuik 2009) Highly chlorinated PCBs are relatively stable, thus are widely distributed and transported through the environment (Spaeth 2011).


Different types of chemicals stimulate toxic effects but usually require their metabolism into activated intermediates; these intermediates remove hydrocarbons from an alkane of macromolecules of critical cells (Spaeth 2011). In the metabolic activation process, many halogenated hydrocarbons undergo cytochrome P-450 dependent oxidation or the hydrocarbon is transformed into an arene oxide or radical intermediates (Safe 1992). These enzymes are used to catalyze the oxidation of organic substances; intermediates include lipid and steroidal hormones (Safe 1992). Arene oxides and radial intermediates are capable of forming strong covalent compounds with cellular components. With PCBs, P-450 enzyme catalyzed oxidation use these toxins as substrates which can transform into highly reactive arene oxide intermediates (Safe 1992). These intermediates are capable to arrange into phenolic metabolites, which can further be oxidized and/or conjugated which can become precursors for sulfur containing metabolites, or they can become further metabolized through different metabolic pathways (Safe 1992).

Methylsulfonyl PCB metabolites contribute to unusual biochemical properties as it's able to bind to uteroglobin, binds to progesterone in mammals (Safe 1992). These metabolites are able to bind to lung-binding proteins with high affinity and accumulate in the lung tissue. In 1968, mass contamination of PCBs occurred in rice bran oil in Kyushu, Japan (Safe 1992). As a result, many individuals have exceptionally high levels of methylsulfonyl metabolites. In humans, this poisoning has altered lung capacity and overall function. This incident brought up the toxic response of the methylsulfonyl metabolites in lungs (Safe 1992).

The rate of PCB metabolism and formation of arene oxide compounds are dependent on an array of factors such the number of chlorines attached to the ring, chlorine patterns, and levels/patterns of P-450 present in organ along with the presence of other drug metabolizing enzymes (5A/1A). For example, P-450 isozymes induce the metabolism of dicholorphenyls with diortho chorine substitution (Safe 1992). The substitution occurs on the carbon next to the chlorine. P-4501 isozymes metabolize diphenyls without substituting ortho chlorines (Safe 1992). Evidence shows PCBs may not be required to become oxidized in pathways that don't contain arene oxide intermediates (Beyer, Bizuik 2009).

From in vivo and in vitro results, PCBs form strong covalent compounds with macromolecules such as protein, RNA, DNA (Safe 1992). These covalent adducts induce the strands of DNA to break apart and repair. This process is onset once the metabolism of PCBs produces arene oxide intermediates and alkylates macromolecules (Safe 1992). This toxin is not noted as bacterial mutagens; molecules of PCBs that are able to metabolize easily tend to be less toxic. Even though PCBs are broken down to through arene oxides to form other products, this toxin undergoes a metabolism for detoxification purposes (Safe 1992). Hydrocarbons are the main source of production of congeners and PCB mixtures which tend to be harmful (Safe 1992).

PCBs and Health

Humans are exposed to PCBs on a regular basis but the question lies in the amount of exposure, rather than the exposure itself (Spaeth 2011). Those who are exposed to PCBs on a daily basis have a greater risk for health problems (Spaeth 2011). PCBs can be ingested through the primary route which includes ingesting or consuming the toxin itself (Safe 1992). Since the toxin is fat soluble, it's easily absorbed in the gastrointestinal tract, which typically increases with congener chlorination (Spaeth 2011). It can be readily stored in the adipose tissue or fat tissue, due to the solubility of the toxin (Safe 1992). This allows PCBs to bio accumulate in animal tissues, which are then consumed by humans (Beyer, Bizuik 2009). PCBs also alter hepatic metabolism and induce microsomal liver enzymes which can cause hepatogenic porphyria and increase of the degradation of steroids in the liver (Bourez 2013). PCBs can enter the body by inhalation, in which the toxin accumulates into the lungs or by dermal absorption via skin contact (Spaeth 2011). The mechanism of dermal absorption and inhalation are still unknown. PCB half-life in the human body can range from a few years up to 20+ years, depending on the congener and physiological breakdown of the human (Spaeth 2011).

Acute effects can be seen on the liver, kidney and central nervous system from oral exposures to PCBS in animal studies (Spaeth 2011). Oral exposure of PCBs in rats shows moderate acute toxicity. Chronic health effects from PCB exposure of the systemic system include: body weight, cardiovascular, dermal, endocrine, ocular and renal (Spaeth 2011). Other systems include neurological, reproductive, developmental and genotoxic systems. A major health effect from PCB exposure is cancer, as PCBs are carcinogens and act as cancer promoters. Some cancers have an increase risk with the exposure of PCBs such as brain, breast, gastrointestinal, liver, lung, Non-Hodgkin's lymphoma, thyroid, and prostate cancers (Spaeth 2011). PCBs are able to mimic estrogen hormones and cause breast cancer (Beyer, Bizuik 2009).

PCBs can lead to recurrent infections by suppressing the antibody and immune response (Beyer, Bizuik 2009). In human studies, PCB exposure has caused an increase in respiratory, skin, ear, and measles infections (Spaeth 2011). In men, PCBs can reduce sperm mobility and decrease testosterone levels by inhibiting the production of testosterone (Spaeth 2011). In females, PCBs can cause an increased menstrual cycle and cause the period to come sooner in young females (Spaeth 2011). In a study with 2,000 pregnant women, a dose response relationship can be seen with PCB levels and diabetes (Spaeth 2011). Dioxin exposure has also lead to elevated diabetes. Pregnant women with this exposure have a significant increase risk to give birth to low weight infants (Bourez 2013). Low birth weights are often associated with chronic diseases in adulthood such as hypertension and diabetes; the exposure potentially can cause long term effects in the child's life (Bourez 2013). In utero, neurobehavioral effects are seen in children after birth with PCB exposure (Bourez 2013). Effects such lower IQ, lower results on achievement tests, low muscle tone and depressed reflexes, along with poor performance on scales to measure emotions and behavioral disorders (Spaeth 2011). In adults, neurobehavioral effects from PCBs are linked to loss of memory and IQ (Spaeth 2011).

Hypothyroidism is common among the diseases in which PCB exposure and decreased thyroid function are linked with each other (Spaeth 2011). In animal studies, PCBs can interfere with the thyroid hormone at multiple sites. Hypertension and high blood pressure is also common with PCB exposure. PCB exposure will cause wide range of cardiovascular diseases and elevated serum lipids which lead to higher plasma triglyceride levels, higher cholesterol levels, and increased mortality of cardiovascular disease (Beyer, Bizuik 2009). Arthritis is a disease with usually occurs with advanced age, but PCB exposure can trigger to cause the joint disease and inflammation (Spaeth 2011).

If breast feeding, PCBs can be exposed to infants via breast milk as milk is lipid-rich. Polychlorinated biphenyls are able to cross the placenta and expose the embryo and fetus during pregnancy (Bourez 2013). Lower birth weights have been reported in five of the ten PCB studies conducted with frequent consumption of fish and marine mammals (Bourez 2013). Consumption of PCBs during pregnancy has been reported to decrease the duration of the pregnancy and decrease head circumference of newborns (Bourez 2013). Inuit newborns exposed to high amounts of PCBs during in utero showed shorter gestation periods. Shorter gestation periods lead to lower birth weight and shorter length and head circumference (Bourez 2013).

PCBs usually accumulate in tissues with an abundance of lipids due to the nature of the toxin, thus adipose tissue is a common reservoir for PCBs (Ayotte et al 2013). In recent epidemiological studies, the presence of PCBs in fat cells can cause changes to the regulation of lipid metabolism (Ayotte et al 2013). High concentrations of PCBs are shown during weight loss due to the increase in lipid mobilization (Ayotte et al 2013). This toxin is released to the blood stream since the toxin creates a tight association to plasma proteins, significantly with lipoproteins and serum albumin, in high concentrations and can cause adverse health effects (Ayotte et al 2013).

Animals and PCBs

Current Research Question

Polychlorinated biphenyls are known to cause harmful effects in humans, but questions remained as to the effects in rats and fish. Along with other toxins, PCB induced effects are dependent on different factors such as age, sex, species, route of administration, duration of exposure and chlorine content of PCB mixture (Beyer, Bizuik 2009). Carcinogenic and toxic responses have been tested and observed in lab animals and cell cultures of mammals. The question remains if rodent and fish species, which are metabolically very different from each other, can respond to this toxin and display some type of health behavior.

If PCBs are exposed to different species, will the two species display similar health effects? The two species researched were fish and rodents. These two species either introduced to the toxin or the toxin was much present in the aquatic environment. Much of the effects on humans were collected data from exposed human populations. It's important to understand how animals and fish species respond to such contaminants as these results can be compared to humans. Rodents are great models as their genetics and behaviors are closely resembled to humans (Safe 1992). If rats display certain behaviors after PCB toxins, then the similar behavior and toxic effect is likely to be seen in humans (Safe 1992).


Aroclor 1260 refers to the molecule as 12 carbons with 60% chlorine mass, manufactured from Monsanto Company from 1930-1977 (Safe 1992). 'Aroclor' was the trade name for PCBs in the market (Safe 1992). In a research experiment, Sprague-Dawley rats were used to monitor the effects of Aroclor 1260 for two years in both male and female rats (Norback, Weltman 1985). The rats were fed different amounts Aroclor for two years with their basal diets; 100 ppm for 16 months and 50 ppm for 8 months (Norback, Weltman 1985). In the female rats, 95% displayed malignant hematoma (liver cancer), or hepatocellular adenocarcinomas as large as six centimeters in diameter (Norback, Weltman 1985). Approximately 12% percent of their weight contributed from the liver, compared to 4% in the control rats. In the male rats, 15% displayed malignant hematoma (Norback, Weltman 1985).

At different stages of PCB treatments, hepatocellular lesions varied in rats as the exposure continued with PCBs (Safe 1992). In the first month, centrolobular cell hypertrophy was observed along with the central and middle lobular regions along with the development of foci of altered cells in the liver after three and nine months of PCB treatment (Norback, Weltman 1985). Neoplastic nodules occurred at 12 months with the formation of trabecular carcinoma at 15 months (Norback, Weltman 1985). In some mice, cystic cholangioma developed with 18 months of treatment, followed by adenofibrosis at 22 months (Norback, Weltman 1985). At the end of the two years, there was a prominent development of adenocarcinoma. With the size of the hepatocellular carcinomas, being large in size, the lobes of the liver were replaced and elevated the liver surface as the neoplastic nodules compressed the area surrounding parenchyma (Norback, Weltman 1985). Slowly, the hepatocellular carcinomas were produced in such rats that have low incidence of spontaneous hepatocellular carcinomas. The mortality of the rats didn't increase due to the slow progression of both the neoplastic process and hepatocellular carcinomas (Norback, Weltman 1985). The different effects of PCBs are seen due to sex-linked differences with activating enzymes and deactivating carcinogens. The presence of sex hormones, androgens and estrogens, also compete with the carcinogen for metabolism (Norback, Weltman 1985). This discovery proves there are different effects of PCBs follow different mechanisms in the body depending on the sex and hormones present (Norback, Weltman 1985).

Another study used Wister male rats to determine chronic effects in this species by introducing PCB mixtures Clophen A60 and Clophen A42 (Safe 1992). These rats were feed 100 ppm of these mixtures and the livers were observed 800 days later. Neoplastic nodules and foci alteration of cells were observed after day 500 (Safe 1992). Foci of hepatocellular cells were always present in all the time intervals whereas neoplastic nodules and hepatocellular carcinomas developed as time increased (Schaeffer, et al 1992). These results show 61% of the rats had hepatocellular carcinomas for Clophen A60, 3% for Clophen A30 and 2% for the control rats (Schaeffer, et al 1992). Hepatocarcinogenicity was more potent in mixtures that contained higher chlorination; Clophen A60 is 60% chlorine by weight, as Clophen A30 is approximately 42% chlorine by weight (Schaeffer, et al 1992). This study indicates sex-linked differences with male rats being susceptible to hepatocellular carcinoma (Schaeffer, et al 1992).

Adipocytes and Lipid Content

PCBs are known as lipophilic pollutants that can accumulate in adipose tissues, as it is fat soluble (Ayotte et al 2013). Three PCB congeners (2,4,4'- trichlorobiphenyl [PCB-28], 2,3',4,4',5-pentachlorobiphenyl [PCB 118] and 2,2',4,4',5,5'-hexachlorobiphenyl [PCB-153] were added in equal amounts to a medium of two in vitro models of adipocytes. Over a four hour period, the absorption of the toxins was observed in adipocytes (Ayotte et al 2013). The in victor models were mouse embryonic fibroblasts (MEF), which differentiate into adipocytes and 3T3-L1 pre-adipocyte cell line (Ayotte et al 2013).

Amount of PCB absorbed is highly correlated with the amounts of triglycerides present in fat cells. In both cell models, majority of the fatty acids were saturated, mono-unsaturated and poly-unsaturated (Ayotte et al 2013). In the 3T3-L1 cell line, odd numbers of fatty acid chains were also observed. Each PCB congener differed with PCB accumulation after four hour incubation (Ayotte et al 2013). PCB-28 accumulated in 23% of MEFs and 81% of 3T3-L1 adipocytes; PCB-118 showed 13% accumulation in MEFs and 82% in 3T3-L1 adipocytes and 13% and 77% accumulation in MEFs and 3T3-L1 respectively for PCB 153 (Ayotte et al 2013). In all three congeners, there was a significant increase in absorption in the 3T3-L1 adipocytes (Ayotte et al 2013). This was due to the amount of intracellular triglyceride contents present in the two cell models. The amounts of PCB accumulation occurred with the high presence of triglycerides, thus 3T3-L1 adipocytes contain higher amounts of triglycerides compared to MEFs (Ayotte et al 2013). This shows a direct role of intracellular lipids with the storing capacity of adipocytes. The fatty acid profiles of cells do not directly influence the amount of PCB accumulation within the adipocytes, even though the amount of fat cells present in adipocytes was causing the accumulation of PCBs (Ayotte et al 2013). The accumulation of PCBs in cells was not driven by the composition of the cellular triglycerides. The amount of triglycerides positively correlates with the accumulation capacity of PCBS (Ayotte et al 2013). The higher amounts of triglycerides present in adipocytes, the higher capacity of PCB accumulation to occur (Ayotte et al 2013).

Among the three congeners, PCB-28 was more readily able to enter the cells (Ayotte et al 2013). Even after one hour of incubation, PCB-28 had already entered 20% of MEFs compared to 6% and 5% for PCB-118 and PCB-153 (Ayotte et al 2013). The same can be seen in 3T3-L1 adipocytes; PCB-28 had accumulation rates of 49% after one hour of incubation, much higher compared to PCB-118 and PCB-153. The physicochemical properties of the chlorine substituent also play a role in the accumulation of adipocytes (Ayotte et al 2013). The molecular makeup of PCB-28, in regards to molecular volume and lipophilcity, allows the toxin to enter the cell faster and bio-accumulate. As this was only true within the first hour of incubation, PCB-118 and PCB-153 caught up with PCB-28 after four hours of incubation (Ayotte et al 2013).


Polychlorinated biphenyls have been significantly decreased in production over the past 35 years, but the chronic effects from mass pollution of this toxin can be seen in fish (Maceina, Sammons 2013). These compounds not only exist in the terrestrial environments, but are commonly found in water sediments. PCBs were released into the water for long periods of time, and have been able to survive in water due to its chemical properties of low solubility and low volatility (Brazova, et al 2011). Organic pollutants are able to accumulate in species at much higher levels compared to the level in the water (Maceina, Sammons 2013). Mortality in fish is relatively low as fish are able to survive normal life spans with high PCB concentrations (Maceina, Sammons 2013). Fish were first thought of having the inability to biotransform PCBs but recent studies show the ability of fish to biotansform these toxins and metabolize into hydroxylated PCBS (Beyer, Bizuik 2009). In the Great Lakes, a number of species were able to synthesize hydroxylated PCBS from the original toxin (Beyer, Bizuik 2009). There rises great concerns as new compounds may contain greater toxicity than the parent compound (Beyer, Bizuik 2009).

Polychlorinated biphenyls in fish are commonly due to PCB consumption as food in fish (Brazova, et al 2011). PCB particles are consumed by phytoplankton and zooplankton at the bottom of the food chain (Brazova, et al 2011). As the food pyramid rises, larger fish in the aquatic ecosystem tend to have higher levels of PCBs due to the consumption of fishes (Brazova, et al 2011). PCB accumulation in fish species differ due to the differences in metabolic processes, season, age, sex and the physiological condition of the animal (Brazova, et al 2011). Different fish species will react differently with the toxin and bioaccumulation in tissues (Brazova, et al 2011).

One of the most contaminated areas with PCBs is located in eastern Slovakia due to vast amounts of PCB compounds released to the water by chemical factories (Brazova, et al 2011). While treatment and decontamination efforts were never executed, prevented measures left the soil and underground water heavily contaminated (Brazova, et al 2011). Currently, there are at least 40,000 tons of sediments with PCB contamination in the water. From this water reservoir, accumulation of PCBs was species specific and varied with trophic habits, weight and position on the food pyramid (Brazova, et al 2011). Nine fish species displayed different concentrations of PCB in muscle tissue in the water reservoir in years 2004 and 2009 (Brazova, et al 2011). Predatory fish, top level of food pyramid, contained higher levels of toxin in both years. But non-predatory fish in freshwater beam and roach was also contained large number of concentrations (Brazova, et al 2011).

Bioaccumulation of PCBs in muscle tissue of fish was dramatically increased two times within five years (Brazova, et al 2011). Two factors play an important role in the bioaccumulation of PCBs: the lipid content of the fish and the position of the fish on the food chain. Two cyprinids, carp and goldfish, were the only species which demonstrated an opposite effect. PCBs were selectively stored in fish tissue (Brazova, et al 2011). The liver contained the largest accumulation of the toxin, largest accumulation seen in northern pike and freshwater beam (Brazova, et al 2011). European perch and goldfish showed greater accumulation in tissue rather than liver; accumulation was also seen in the kidney and brain but at a less extent (Brazova, et al 2011). Higher accumulation levels were found in predators due to the higher trophic levels along with bottom feeding fish, lowest trophic level, by consuming the toxin directly from the sediment with toxin (Brazova, et al 2011). Predators contain such high levels due to the consumption of fish with PCB contamination; the same levels can be seen in feeding fish due to the direct exposure of PCBs from the sediments (Brazova, et al 2011). As predatory fish contained more PCB accumulation in fish, non -predatory contained higher levels of accumulation in the brain. This may occur due to the direct exposure to the toxin versus toxin exposure through fish consumption (Brazova, et al 2011).

There is no correlation with PCB concentration and overall weight of fish, there are correlations seen individual PCBs and total amounts of the toxin in organs of fish, specifically the brain (Brazova, et al 2011). Certain congeners of the toxin can be seen in predatory and non-predatory fish. Congener 138 is commonly found in non-predators whereas congener 28 is commonly found in predators (Brazova, et al 2011). Decreased levels of PCB accumulation in brain tissue can be seen in growing fish due to increased metabolic rates or due to the dilution effect in fat tissues as the body weight of the fish increases (Brazova, et al 2011).

Hudson River

In the mid-1970s, PCBs were suspected to cause effects in fish reproduction and disruption in the endocrine system (Maceina, Sammons 2013). Fish in the Hudson River have been extensively studied to see the effects of General Electric plants discharging PCBs into the upper Hudson River from two of their manufacturing plants (Maceina, Sammons 2013). Recreational fishing has been banned from 1976-1995, catch-and-release fishing has been allowed since 1995 (Maceina, Sammons 2013). Different species of fish were collected annually from different PCB discharge sites and sites downstream from PCB discharge sites. Amounts of PCBs present were measured in adult fish with estimated abundance and the size of the offspring at the age of one (Maceina, Sammons 2013).

Perca flavescens (Yellow Perch), smallmouth bass, and largemouth and their reproduction behavior have been studied by measuring PCB concentrations in the Hudson River since the 1970s (Maceina, Sammons 2013). These fish have been collected from three major habitats: rock-riffle, river banks, and submersed and floating leaved aquatic vegetation. For five years (2004-2009), these fish were collected at night with direct-current electrofishing (Maceina, Sammons 2013). Total lengths and weights were recorded for all fish. PCB concentrations in the fish were measured by tissue samples with the use of gas chromatography in 960 fish; PCB samples included Aroclors 1016, 1221, 1232, 1242, 1248, 1254, and 1260 (Maceina, Sammons 2013). These PCB samples contained 10-12 carbons with 16-60% chlorine mass. PCBs in young fish were not measured but as these young fish were the products of adults' fish with PCB toxins and in the same habitat; these fish are reflective to this exposure in age one fish as the adult fish contained PCB concentrations (Maceina, Sammons 2013).

Wet-weight PCBs and lipid-based PCBs were much lower in adults fish were much lower in fish that were downstream from the initial General Electric plants with PCB discharge (Maceina, Sammons 2013). Yellow perch were the exception; these fish didn't show significant differences with PCB concentrations in areas upstream and downstream from the plant (Maceina, Sammons 2013). But all three species of fish showed different amounts of the toxin present between two time frames, 2004-2006 and 2008-2009. There was a significant decrease of PCB concentrations in 2008-2009 for these fish species (Maceina, Sammons 2013).

In age one fish, the length of these species were much smaller in this river compared to average sizes (Maceina, Sammons 2013). All three species had variances in length by years; lengths in some years were longer compared to other years (Maceina, Sammons 2013). In 2007, there were there lower abundances of age-1 in all three species and were associated with the daily flows from June 1st to September 30th in 2006 with an average 194 m3/ sec daily flow (Maceina, Sammons 2013). In the same range for 2007, there was an average flow of 68 m3/ sec with high catch rates for age one fish in 2008. In all, the three fish species the size of age one fish was independent with PCB exposure compared to the growth of the fish in control pools (Maceina, Sammons 2013). In fact, these fish were greater in size with PCB exposure compared to fish in the control pools. In laboratory fish, PCB exposure in eggs, larvae and juvenile fish showed abnormal development with poor growth and short life span (Maceina, Sammons 2013). Larger growth of these fish species may be related to an increase in nutrients and primary production in the Hudson River (Maceina, Sammons 2013). This river is eutrophic compared to oligomestrophic which is able to provide nutrients, despite PCB concentrations in the sediments (Maceina, Sammons 2013).

Polychlorinated biphenyls in sediments were directly associated with the toxin concentration in adult black bass and yellow perch (Maceina, Sammons 2013). Majority of the bioaccumulation occurs with dietary uptake of the toxin from bottom dwelling organisms. This research found no correlation with PCB concentrations in adult muscle tissues and reduced reproduction success in all the species at a population level (Maceina, Sammons 2013). A small correlation was seen in yellow perch fish with PCBs in adults and age one of yellow perch fish in vegetation sites, but there was a high abundance of these age one fish at the PCB site (Maceina, Sammons 2013). The opposite has occurred; there has been a positive correlation with maternal PCB concentration and reproductive success with age one white perch. In the Housatonic River, PCB concentrations were greater in adult fish compared to the Hudson River (Maceina, Sammons 2013). The toxin didn't show effect on the population level of largemouth bass with growth and reproduction levels. These fish species are still able to reproduce in highly concentrated areas in aquatic systems with PCBs (Maceina, Sammons 2013).

Significance and limitations of Research


The results of the research studies show a link with PCB exposure and toxicity within organs. In both studies, there has been a correlation with PCB accumulation within the liver. In rodents, the occurrence hepatocellular carcinomas are observed in Sprague-Dawley rats. As rats are quite similar to human in genetics and behaviors, the occurrence of liver cancer may also be seen in humans. Many predator fish contain high concentrations of PCBs, which are transferred from feeding fish. Predator fish often contained higher amounts of PCB accumulation in the liver, compared to feeder fish. This same concept can be duplicated in humans. Humans, who act as predators, may consume fish with high concentrations of PCBs. This allows one to know more about the transfer of the toxin. Even though this toxin hasn't been manufactured over the last few decades, it's able to transfer from one species to another. This study can prevent the toxin accumulation in humans, as they may also consume fish with high accumulations of PCBs. These toxins are directly transferred and can cause significance health defects.

The physiological factors of the toxin are now known after research was done in two tissues of mice. PCBs are known to accumulate in adipose tissue but the mechanism was unknown until research showed the accumulation of PCBs occurring in triglycerides of fat cells. The presence of fat cells didn't correlate with the PCB accumulation; the presence of triglycerides in fat cells contributed the accumulation of PCBs. With this finding, mechanisms can be developed to block the toxin by looking at the fatty acid chains in adipose tissue. Since the PCB binding occurs on a molecular level, there may be ways to block the PCB binding with triglycerides. If PCB binding occurs in triglycerides of adipose tissue in humans, then this method would prevent humans, who are highly exposed to the toxin, from developing cancer and accumulation. This may be highly useful in pregnant women who consume fish from PCB polluted sites and may reduce the spread of PCB across the placenta to the fetus.


Many of these studies focused on rodent and fish species. Humans and rodents may be genetically similar but they aren't 100% similar. Rodents may be genetically more confined to the toxin compared to humans. The effects of PCB accumulation is humans may not occur at the same rates in rodent and fish species. Rodents and fish are much smaller in size compared to humans. The accumulations in these species may occur in shorter time periods as their organs are smaller than those of humans. Rodents also show a high accumulation of the toxin in the liver and the effects were seen within a few years. Toxic effects in human may not be seen for years due to their size.

In fish, only a handful of species have been studied with PCB accumulation in fish. This study is only limited to the few fish species studied near PCB manufacturing sites. Majority of the research was done with yellow perch, smallmouth bass and largemouth species. These fish species are relatively small compared to sharks and whales. All fish species will not replicate the same absorption mechanism of this toxin. Some fish species may be able to block the toxin from binding to adipose tissue. Much of the research has also only focused on adipose tissue.

Much of the research was conducted near PCB manufacturing sites and samples were taken within a close proximity of the site. This limits the research as only water and fish species near the site are examined. PCBs are able to settle in sediments and transfer with water currents far from the original contamination sites, this limits the water samples studied. Also, much of the research done with water samples and animal species were studied in the 1990s and early 2000s. This makes some of the research old and needs to be updated with current samples of PCB in water sites and animal species.


Polychlorinated biphenyls are always going to be persistent in the environment due to the chemical nature of their molecular structure which allows the toxin to not become biodegradable. This means the toxin is always present in the ecosystem. This can lead to massive amounts of research on non-animal species such as trees near the contaminated water. Trees can also become affected by toxins which limit their growth and effects pollinations. It would be interesting to see if there are any effects of PCBs on different tree species. Future research on trees can determine maximum limitations before seeing adverse effects such as limited growth, reproduction, and leave counts.

This toxin is found in soil sediments in the aquatic systems, but the toxin may be found in non-aquatic soils near manufacturing sites. PCBs accumulations were extremely high in aquatic systems near the site and accumulated in fish species near the site. PCBs may have been able to leak into the ground soils and cause some type of adverse effect. Soil samples can be examined to determine if PCB samples are found and will allow researchers to look at the impact of the soil. Soils that are nutrient rich may start leeching with the presence of PCBs and start causing the vegetation in those soils to die. Future research can determine amounts of PCBs toxins that soils can tolerate before the soils start losing nutrients and kill plants in those soils. These toxins may have been able to travel deep enough and contaminate water pipes that aren't fully enclosed. Water pipes that have leaks may allow the toxin to enter the water lines and effect people who drink the water. The exposure may have been low, as the toxin would have to travel down great depths of ground soils, but it may be able to cause accumulation within the population. People near the contaminated sites may need to undergo lab studies to ensure they aren't accumulating this toxin from their water pipes.

Sharks have a highly resistant to cancer and are often studied in labs to determine their biological processes that allow them to be so resistant. PCBs are able to cause cancer in humans, but research can be done to see the effects on sharks. Much of the fish species studied in previous research are relatively small compared to sharks, which means their organs are much smaller. Smaller amounts of toxins may be needed to these small fish species in order to see PCB accumulation in these fish. But in sharks, as they are highly resistant to cancer, they may need extremely large amounts of PCB exposure in order for the toxin to accumulate. It would be interesting to see if sharks are able to resist PCB accumulation and not get cancer. If this occurs, sharks may be the perfect model to determine how PCBs may be blocked in their bodies and prevent the accumulation in humans.

Works Cited Page:

Beyer, Angelika, Bizuik, Marek. "Environmental Fate and Global Distribution of Polychlorianted Biphenyls." Reviews of Environmental Contamination and Toxicology 201 (2009): 136-59. Print.

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