Assessment Strategy for Evaluating the Environmental and Health Effects of Sanitary Sewer Overflows from Separate Sewer Systems




НазваниеAssessment Strategy for Evaluating the Environmental and Health Effects of Sanitary Sewer Overflows from Separate Sewer Systems
страница46/64
Дата конвертации28.10.2012
Размер2.21 Mb.
ТипДокументы
1   ...   42   43   44   45   46   47   48   49   ...   64

Fates of Heavy Metals

Arsenic


Callahan, et al. (1979) stated that all of the potential environmental fates, except for photolysis, can be important for arsenic. Bioaccumulation of arsenic, however, is limited because of its toxicity: the organism usually dies before the bioaccumulation of the material can reach high magnitudes. Arsenic can also be metabolized by organisms to form trivalent arsenicals. Arsenic can be adsorbed onto clays, iron oxides, inorganics and either remain suspended or accumulate in sediments. The EPA (1976) stated that compounds of arsenic are ubiquitous in nature, insoluble in water and occur mostly as arsenides and organic arsenopyrites. Arsenics exist in the trivalent (+3) and pentavalent (+5) states as either organic or inorganic compounds. The trivalent inorganic arsenicals are more toxic than the pentavalent forms both to mammals and aquatic species. Most arsenic forms, however, are toxic to humans. Phillips and Russo (1978) stated that arsenic may be bacterially methylated, much like mercury, to form highly toxic methylarsenic or dimethylarsenic. These methylated forms of arsenic are very volatile and are readily oxidized to less toxic forms.

In a survey of 130 natural receiving water quality monitoring stations, the EPA (1976) reported that the ranges of observed arsenic values was 5 to 336 g/L, with a mean value of 64 g/L. Durum (1974) reported that a survey of 728 USGS water samples resulted in a range of 10 to 1,100 g/L with a median value of less than 10 g/L. The maximum value was found in the southeastern region of the U.S. The median and minimum values were similar for all areas of the country. The southwestern and northwestern parts of the country had the lowest maximums observed (10 and 30 g/L respectively) while the New England region had a maximum value of 60 g/L and the central states had a maximum value of 140 g/L.

Phillips and Russo (1978) reported that arsenic is accumulated by fish from both water and food, but the concentration factors are quite low. Arsenic in fish tissue is concentrated in the fish fat. The muscle tissue also accumulates arsenic, but the biological half-time has been reported to be only 7 days in green sunfish. Shellfish, however, concentrate arsenic to a much greater extent than fish, Marine organisms contain more arsenic than fresh-water forms. Unfortunately, arsenate present in the tissues of consumable seafood is rapidly converted to arsenite following death, a much more poisonous form (Phillips and Russo 1978). The EPA (1976) reported that even though arsenic is accumulated in aquatic organisms, it is not progressively concentrated along a food chain.

Phillips and Russo (1978) reported that arsenic concentrations in fish collected from various Wisconsin waters was typically less than 1 g/g. Young and adult bluegills placed in ponds treated with sodium arsenite, a herbicide, contained arsenic levels similar to the concentration of arsenic in the pond (0.3 to 9 mg/L) after 16 weeks exposure. They also reported that arsenic concentrations in mature fish muscle were about 60 percent of whole fish arsenic concentrations. However, immature bluegills obtained arsenic concentrations almost twice the adult concentrations. The EPA (1978) reported that arsenic concentrations for freshwater fish are usually below 1 g/g wet weight, with concentrations of 0.5 g/g for bluegills and 0.07 to 0.15 g/g for trout. They also reported on a study that incubated rainbow trout eggs in water containing various concentrations of sodium arseniate or arsenic trioxide. A similar accumulation pattern was observed for both arsenic compounds. The embryos accumulated up to 2.5 g/g arsenic after 40 days exposure to only 0.05 mg/L of arsenic. This corresponds to a bioconcentration factor of 50. Interestingly, concentrations as high as 50 mg/L arsenic did not reduce egg survival, but concentrations less than 5 mg/L decreased survival because the higher arsenic concentrations reduced growths of fungus on the fish eggs. In another study, Phillips and Russo (1978) report that arsenic was rapidly accumulated in largemouth bass from both food and water sources, but the arsenic was rapidly eliminated after the exposure was terminated. The arsenic concentrations necessary to control aquatic vegetation would not result in arsenic concentrations in bass considered dangerous to human consumers. Average arsenic bioaccumulations in various freshwater fish species in the Southeastern United States are reported as 0.5 g/g. However, liver oil from the fish averaged almost 40 g/L arsenic.

Lake Michigan plankton and benthic organisms were found to contain about 6 g/g arsenic. Leland and Luoma (1979) have also summarized many past studies of heavy metal bioaccumulation. They reported a range of arsenic tissue concentrations in benthic invertebrates ranging from less than 1 to 1,300 g/g and a range of arsenic concentrations in zooplankton from 700 to 2,400 g/g. Neff, et al. (1978) also reviewed many bioaccumulation studies and found that the bioconcentration factors for macroinvertebrates were usually greater than for other organisms. Concentration factors ranged from 300 to 3,300 for arsenic in macroinvertebrates. Leland and Luoma (1979) also reported another study that examined trace metal uptake by invertebrates in laboratories. They found that arsenic uptake by snails was similar to the arsenic concentrations in the sediment, rather than in the water.

Arsenic can also accumulate in aquatic vegetation. The EPA (1976) reported on a previous study in a Wisconsin lake that had water concentrations ranging from 100 to 450 g/L arsenic. The concentration of arsenic in the bottom mud was about 200 g/g. A sample of cladophora contained more than 1,200 g/g arsenic and fresh shoots of mature Myriophyllum sp. stems contained as much as 550 g/g arsenic.

Cadmium


Callahan, et al. (1979) stated that in most unpolluted waters, the majority of cadmium will exist as the hydrated divalent cation. In polluted waters, complexes with organic materials will be the most important cadmium forms. The affinity of ligands for cadmium follows the order of humic acids greater than carbonates, carbonates greater than hydroxides and hydroxides greater than both chlorides and sulfates. Adsorption of cadmium onto organics, clays, hydrous iron and manganese oxides is also important in polluted water. Cadmium is also strongly bioaccumulated. Durum (1974) stated that concentrations of the carbonate and hydroxide forms of cadmium, with pH values equal to or less than 7, are relatively high and that the USPHS standard of 10 g/L may occur in many stable water systems, including both surface and groundwaters. Pitt and Amy (1973) studied the solubility of cadmium in street dirt, along with other metals, and found that in typical urban runoff concentrations, soluble cadmium values of less than 1 g/L occurred in moderately hard water (hardness equal to 50 mg/L) after an exposure of 25 days. This soluble fraction was 14 percent of the total cadmium in the mixture. Wilber and Hunter (1980), in an urban receiving water study in Lodi, New Jersey, found that with most low flows in the Saddle River, the cadmium was mostly dissolved. However, during wet weather conditions, most of the cadmium was associated with particulates.

Durum (1974) in the nationwide USGS study of water quality conditions that analyzed 727 samples, observed an overall range of cadmium of less than one to as much as 130 g/L. The nationwide minimum and median values for all regions of the country were all less than 1 g/L, except in New England where the median value was 2 g/L. The maximum observed value of 130 g/L was found in the southwest. A maximum value of 90 g/L was observed in the southeast, and maximums of 40, 21 and 32 g/L were observed for the central, northwest and New England areas respectively.

Phillips and Russo (1978) reported that very little cadmium is accumulated in the eatable portions of fish. However, shellfish are capable of accumulating extremely high levels of cadmium in eatable portions. Cadmium is readily available through both food and water to marine and freshwater organisms. Either source can result in toxic symptoms by fish. The fish tissues appear to reach cadmium equilibrium after about 2 to 5 months exposure. Soft water usually results in higher cadmium bioaccumulations in fish than when in hard waters. Cadmium uptake also increases with increasing water temperature and decreasing salinity. Neff, et al. (1978) studied cadmium concentrations in a relatively unpolluted Illinois stream system. Cadmium was found in all components of the aquatic system. Fish and sediment cadmium concentrations were similar, but aquatic insects had cadmium concentrations higher than the sediments. There was, therefore, no indication of food chain cadmium magnification.

Leland and Luoma (1979) reported a study that observed museum fish specimens that had been collected over a 40-year period which did not detect any chronological accumulation of cadmium. Phillips and Russo (1978) also noted another study where cadmium was not detected in any fish samples collected in Wisconsin. Fish samples, however, collected from the Iowa River did contain low concentrations of cadmium. Another study exposed rainbow trout to high cadmium concentrations and found that most of the cadmium in the gills were rapidly lost when the fish were returned to clean water. However, almost no cadmium was lost from the kidney. In another study, cadmium concentrations in the gills of bluegills greater than 150 g/g almost always killed the fish. Three spine sickleback experienced concentration factors from 500 at the lowest cadmium water concentrations to about 0.5 at the highest cadmium water concentrations. The water concentrations ranged from 1 g/L to 100 mg/L and all were lethal. The EPA (1978) reports that cadmium concentrations of more than 0.1 g/g in goldfish were evident from samples from the Hudson River.

Neff, et al. (1978) reported on another study that examined the accumulation of cadmium by freshwater snails. It was found that the initial cadmium uptake rate for the snails was higher in hard water than in soft water, but the total amount of cadmium accumulated in the snails was greater for soft water environments. Leland and Luoma (1979) reported bioconcentration factors of more than 500 for cadmium in crayfish after exposures of 1 mg/L of cadmium for about 1 week. Spehar, et al. (1978) reported that cadmium concentrations in invertebrates increased with increasing water concentrations and were as much as 30,000 times greater than the water cadmium concentrations. Insect and snail body cadmium concentrations of 1 to 10 g/g occurred with water concentrations of 1 g/L, while the insect and snail cadmium concentrations increased to 100 to 200 g/g when the water concentration was increased to 300 g/L. Neff, et al. (1978) reported macroinvertebrate cadmium concentration factors of 82,000 to 182,000.

Phillips and Russo (1978) reported a study that found that the feces of migratory waterfowl contained high levels of cadmium. It was felt that waterfowl can contribute significant quantities of cadmium to the Illinois lake that was studied. In another study, cadmium was found to increase in concentration moving from water to fish to sediment to invertebrates in an Illinois stream. It was found that aquatic insects contained the highest cadmium level possibly due to their close association with the sediment. Rolfe, et al. (1977) found significant retention of cadmium in predator protozoa in another Illinois stream study.

Ray and White (1976) examined cadmium bioaccumulation in aquatic plants. Bioaccumulation factors from 1 to 260 were observed for various plants and plant parts in clean water systems, while the range in polluted water was only 0.2 to 2. Plant tissue cadmium concentrations ranged from about 0.5 to 6 g/g. Leland and Luoma (1979) reported sphagnum moss cadmium concentrations of 1 to 2 g/g. DePinto, et al. (1980) reported a study that found cadmium rapidly taken up by algae. The algae was also a more efficient collector of the cadmium than the sediments. The EPA (1978) reported a study that found cadmium concentrations being greater in the roots of aquatic plants than in the plant shoots. They concluded that roots can take up large quantities of cadmium from solutions, but there are restrictions to cadmium movement through the plant.

Chromium


Phillips and Russo (1978) stated that in water, trivalent (+3) chromium exists as a complex, colloid or precipitate, depending on pH. The more toxic hexavalent (+6) chromium form is usually present only as an ion. Pitt and Amy (1973) found that the solubility of chromium associated with street dirt in moderately hard water was about 4 g/L or about 0.3 percent of the total chromium in the mixture.

The EPA (1976) found, in a nationwide survey of 1,577 analyses, an overall observed chromium concentration range of 1 to 112 g/L, for 386 samples that had detectable chromium concentrations. The mean value for the positive test samples was 9.7 g/L. Durum (1974) reported that a USGS survey of 728 samples resulted in an overall range of chromium of less than 1 to a maximum of 19 g/L. The median observed value was less than 1 g/L.

Phillips and Russo (1978) reported on a study in New Zealand that showed that chromium was concentrated through a simple food chain of sediment to bacteria to tubificid worms. They also reported a study in Wisconsin that showed typical chromium concentrations in fish were less than 1 g/g. Fish exposed to chromium in water can bioconcentration chromium nearly 100 times. Fish, however, were shown to rapidly eliminate chromium when returned to clean water. Therefore, chromium is not likely to accumulate in fish tissue if only exposed to intermittent high chromium concentrations. Chromium is apparently accumulated in fish through the gills and eliminated through the feces. Phillips and Russo (1978) also reported another study of rainbow trout that showed hexavalent chromium bioaccumulations when the chromate water concentration was greater than 10 g/L. The chromium continued to accumulate for at least 30 days. An equilibrium concentration of chromium in rainbow trout appears to be reached rapidly.

Rubin (1976) reported chromium concentration factors of about 10 for fish and more than 250 for mollusks. Leland and Luoma (1979) reported chromium concentrations of 1.8 to 4.6 g/g in aquatic sphagnum moss plants.

Copper


The EPA (1976) stated that copper occurs as a natural or native metal and in various mineral forms, such as cuprite and malachite. Callahan, et al. (1979) stated that copper in unpolluted waters is mostly a carbonate complex and in polluted waters forms complexes with organic materials. Pitt and Amy (1973) found that inorganic copper is mostly found with valence states of plus one and plus two in natural water systems near neutral pH values. The common inorganic copper forms at these pH conditions are copper sulfide, oxide, hydroxide, cyanice, sulfate and iodide. Phillips and Russo (1978) stated that divalent copper ion (+2) and its hydroxides are believed to be the toxic copper forms for fish. Alkalinity and pH are believed to be the major factors controlling copper speciation. Callahan, et al. (1979) stated that copper speciation with organics is most important in polluted waters. The adsorption of copper can reduce is mobility and enrich suspended and settled sediments. Copper is absorbed onto organics, clay minerals, hydrous iron and manganese oxides. They also reported that copper is strongly bioaccumulated.

Wilber and Hunter (1980), in a study of an urban river in Lodi, New Jersey, found that the readily available copper (at a pH of about 7) was about 13 percent of the street dirt and runoff solids total copper content. Pitt and Amy (1973) found that the copper solubility of street dirt was about 160 g/L, or about 36 percent of the total copper in the mixture, with moderately hard water conditions.

The EPA (1976) found that about 74 percent of the more than 1,500 copper analyses in nationwide waters had detectable copper concentrations, with an average value of 15 g/L and a maximum observed value of 280 g/L. Goldschmidt (1958) reported a range of copper concentrations near the estuary of the Mississippi River of 1 to 15 g/L. He also reported a concentration range of 9 to almost 400 g/L observed in three Connecticut lakes, and a range of 65 to 600 g/L for 25 municipal water supplies throughout the country.

Phillips and Russo (1978) reported that copper is bioaccumulated by fresh water and marine fish, shellfish and aquatic insects. They also found that chronic symptoms in fish start to develop very soon after copper bioaccumulation rises above background levels. They also reported on a study conducted in a New Zealand river that confirmed the potential for copper being concentrated as it moves through a simple food chain consisting of metal enriched sediments to bacteria to tubificid worms.

Phillips and Russo (1978) reported that rainbow trout tissue copper concentrations ranged from 1.7 to 12.9 g/g when the trout was collected in a hatchery with a pristine water supply. Bluegill was also found to accumulate copper when the water concentrations were greater than 40 g/L. This concentration also resulted in decreased larval survival for bluegills. In another study, bullheads were found to accumulate copper at all water concentrations exceeding 27 g/L. Rubin (1976) reported copper concentration factors of about 60 for fish, 1,500 for mollusks and about 160 for macrophytes.

Phillips and Russo (1978) reported on a study that found clams and tubificid worms, along with another benthic organisms, containing higher copper concentrations than either omnivorous or carnivorous fish. Neff, et al. (1978) found that concentration factors for macroinvertebrates were higher than for almost all other test organisms. The concentration factors ranged from 2,400 to 3,500. In another study, Phillips and Russo (1978) summarized that insects in a heavy polluted mine stream contained as much as 6,400 g/g copper. They also reported another study's conclusion that mayflies and stoneflies were more resistant to copper pollution than fish, and that their copper accumulation reflected copper water exposure. Neff, reported another study that showed that copper tolerant worms accumulated copper more rapidly than non-tolerant worms.

DePinto, et al. (1980) reported a study that showed rapid copper bioaccumulation in algae when the resultant algae concentrations were greater than sediment copper concentrations. Ray and White (1976) reported copper concentrations in various aquatic plants sampled in polluted and unpolluted reaches of an urban creek. Plant concentrations in an unpolluted stream reach ranged from about 3 to 200 g/g and from about 13 to 240 g/g for other algae and plant species in the polluted reach. The approximate copper bioconcentration factors ranged from about 1 to 20 in the clean stream reach and about 0.1 to 4 in the polluted reach of the stream. Leland and Luoma (1979) reported sphagnum moss copper concentrations of 13 to 540 g/g.

Iron


Phillips and Russo (1978) stated that the soluble ferrous form of iron (+2) is readily oxidized to the insoluble ferric, or trivalent (+3) state in most natural surface waters. A substantial fraction of iron in natural waters is therefore associated with suspended solids. The EPA (1976) stated that the ferrous form can persist in waters void of dissolved oxygen, and originates usually from anaerobic groundwaters or from mine drainage. Iron can exist in natural organometallic, humic, and colloidal forms. Black or brown “swamp waters” may contain iron concentrations of several milligrams per liter in the presence or absence of dissolved oxygen, but this iron form has little effect on aquatic life because it is complexed and relatively inactive chemically or physiologically.

Pitt and Amy (1973) found that the solubility of iron in street dirt was about 50 g/L, or much less than 1 percent of the total iron in a mixture with a moderately hard receiving water. They also stated that the principle inorganic iron forms, with neutral pH water conditions, are iron oxide, hydroxide, sulfate, nitrate and carbonate.

Phillips and Russo (1978) reported that iron is concentrated to a considerable degree by some marine organisms, with most of the iron being accumulated in the gills. Iron was also found to concentrate as it moved through a simple food chain from sediment to benthic worms.

Phillips and Russo (1978) reported that whole bluegill iron concentrations averaged about 150 g/g from a South Carolina reservoir. While carp from Austria was found to have iron tissue concentrations ranging from about 7 to 40 g/g. The gill iron concentrations of this carp was almost 15,000 g/g. The concentrations of iron on the gills were similar to the iron concentrations in the suspended sediments, suggesting that the metals were on particles embedded on the gill surfaces. Rubin (1976) reported iron bioconcentration factors of about 200 for fish and more than 3,600 for macrophytes. Phillips and Russo (1978), in reviewing many studies, were not able to find any age related iron increases. They also reported Lake Michigan benthos iron concentrations of about 1,800 g/g. Leland the Luoma (1979) reported sphagnum moss iron concentrations of about 150 to 2,800 g/g.

Lead


The EPA (1976) stated that most lead salts are of low solubility. Lead exists in nature mainly as lead sulfide (Galena). Other common natural forms of lead are lead carbonate (Cerussite), lead sulfate (Anglesite) and lead chlorophosphate (Pyromorphite). Stable complexes result from the interaction of lead with organic materials. The toxicity of lead in water is affected by pH, hardness, organic materials and the presence of other metals. The aqueous solubility of lead ranges from 500 g/L in soft water to 3 g/L in hard water. Durum (1974) stated that lead carbonate and lead hydroxide are soluble lead forms at pH vales of 6.5, or less, with low alkalinity conditions (less than 30 mg/L alkalinity). The soluble lead concentrations under these conditions can reach 40 to several hundred g/L. If the alkalinity is greater than 60 mg/L and if the pH is near 8, however, the dissolved lead would be less than 10 g/L. Callahan, et al. (1979) stated that lead carbonate and lead sulfate control lead solubility under aerobic conditions and normal pH values. Lead sulfide and lead ions, however, control lead solubility in anaerobic conditions. In polluted water, the organic complexes of lead are most important in controlling lead solubility. Phillips and Russo (1978) stated that most lead is probably precipitated in natural waters due to the presence of carbonates and hydroxides.


Pitt and Amy (1973) found that the solubility of lead in a street dirt mixture was about 40 g/L, or about 3 percent of the total lead, in moderately hard water. Wilber and Hunter (1980) found that readily available lead was about 21 percent of the total lead in street dirt and runoff solids. They also found that under most low flow river conditions, most of the lead was dissolved, but under wet weather conditions, most of the lead was insoluble. Solomon and Natusch (1977) also examined the solubilities of lead associated with street dust. They found solubilities ranging from 500 to 5000 g/L which was 0.03 to 0.3 percent of the initial mixture total lead concentration. However, the test mixture of street dirt with water was very high (1750 mg/L lead).

Rolfe and Reinbold (1977) found that about 46 percent of the total lead input in a test watershed remained airborne. The total input included gaseous and particulate vehicle emissions. About 5 percent of the total lead input to the watershed occurred with rainfall and about 60 percent occurred with atmospheric settleable solids. They found that about 80 percent of the lead in stream water was insoluble and associated with suspended solids, and only 3 percent of the lead input into the watershed exited the watershed in the stream. The streamflow accounted for the majority of all of the lead discharged from the watershed (about 7 to 8 percent of the total lead input).

Pitt and Amy (1973) reported that most inorganic lead in water systems near neutral pH conditions exist in the plus 2 or plus 4 valence states as lead sulfide, carbonate, sulfate, chromate, hydroxide, chloride or iodine.

Drumum (1974) reported lead concentrations in 727 nationwide samples. The reported range was less than 1 to 890 g/L with a median value of 2 g/L. The observed minimum values were all 1 g/L, or less. The median values were all 1 g/L, except for New England where the median value was 6 g/L, and in the southeast where it was 4 g/L. The maximum value of 890 g/L was reported for New England. Maximum lead values of 84, 44, 34 and 23 g/L were reported for the central, southeast, southwest and northwest regions of the country, respectively. The EPA (1976) reported a range of 1 to 10 g/L as the natural mean lead concentration of the world’s lakes and rivers. In 1,500 analyses, less than 20 percent had detectable lead concentrations, with a reported mean value of about 20 g/L and a maximum value of 140 g/L.

Lead is present in all animals and as for many heavy metals, animals higher up in the food chain can bioaccumulate higher quantities of lead in their bodies (EPA 1978). Rolfe, et al. (1977) collected plant and animal tissues from terrestrial and aquatic urban areas. They found that most of the lead was concentrated in the soils, plants, animals and insects in the urban area or near high traffic volume rural highways. They found that the lead concentrations of aquatic organisms varied substantially within and between the urban and non-urban sectors of their test area near Champaign, Illinois. Lead concentrations in organisms from the urban sector were 10 to 20 times higher than those from the rural area. They, however, found no biological magnification of lead through the aquatic food chain, which conflicts with much of the published information. They found that biological lead concentrations were influenced by the amount of contact an organisms had with the polluted stream substrate. They therefore, concluded that external contact is a more important lead uptake mechanism that ingestion. They also found that the uptake rate and final lead body concentrations were proportional to the amount of lead in the water solution when the other lead sources were eliminated.

Phillips and Russo (1978) report that the Canadian Food and Drug limit for lead in fish food is 2 g/g. They also report that most of the lead accumulated by aquatic animals is in the divalent form which increases with decreasing pH values. Neff, et al. (1978) reported that for an unpolluted Illinois stream, that lead concentrations in sediment and aquatic insects were similar and higher than in fish. Fish lead bioaccumulation concentrations, however, were greater than the water lead concentrations. Snails had the next highest bioaccumulations of lead. Again, there was no indication of food chain magnification of lead in this study area.

Rolfe, et al. (1977) reported body tissue lead concentrations in aquatic organisms in a rural stream near Champaign, Illinois, ranging from about 1.4 to 16 g/g. Crayfish samples had concentrations of about 5.4 g/g, mayfly nymphs had concentrations of about 10 g/g and leaches and aquatic worms had concentrations of about 13 g/g. They also found no biological magnification of lead in this food chain. Phillips and Russo (1978) reported that in a study conducted in New Zealand, bioconcentration of lead in a simple food chain did occur from sediments to bacteria to tubificid worms. Almost all studies showed higher lead concentrations in benthic invertebrates than in the sediments, however, predator fish typically had lower lead concentrations than the benthos. Fish usually have greater body lead concentrations than the water concentration. Therefore, the magnification of lead through a complete freshwater aquatic food chain is uncertain.

Phillips and Russo (1978) report that lead concentrations in fish livers greater than 50 g/g and fish kidneys above 100 g/g may indicate a history of unacceptable lead exposures. In another study, fish collected in Wisconsin typically had whole body lead concentrations less than 1 g/g. Leland and Luoma (1979) reported on a study that examined Hudson River fish collected over a 30-year period, ending in 1975, showing no significant chronological lead increases.

Phillips and Russo (1978) report on a study that showed that most of the lead was still retained in rainbow trout after they had been returned to clean water. In another study, pumpkinseed sunfish bioaccumulated lead three times as much at a pH of 6.0 then at a pH of 7.5. The EPA (1978) reported that most freshwater fish contain at least 0.5 g/g lead, with green sunfish containing as high as 16 g/g lead.

Phillips and Russo (1978) summarized a report that discussed isopods that had accumulated lead from both food and water. The most lead-tolerant isopods were found to bioaccumulate the most lead in their tissue. Rolfe, et al. (1977) noted dramatic increases in lead concentrations in tubificid worms during a period of high urban runoff. However, they also found that the amount of lead transported by drifting stream invertebrates was insignificant in this south central Illinois watershed. Spehar, et al. (1978) reported that lead concentrations were as much as 9,000 times greater than corresponding lead concentrations in the water. With 1 g/L lead water concentrations, the bioaccumulation factor was about 10 to 30 times for insects, snails and amphipods. With water concentrations of 600 g/L, however, the bioaccumulation factor was reduced to 2 to 3 times, with resultant insect, snail and amphipod concentrations of 1,000 to 2,000 g/g. Neff, et al. (1978) found macroinvertebrate lead bioconcentration factors higher for almost all other organisms. These lead bioconcentration factors ranged from 7,000 to 100,000. Phillips and Russo (1978) reported a study conducted in a polluted Colorado stream where the insects contained up to 6,000 g/g lead.

Rolfe, et al. (1977) studied the uptake of lead that was deposited on plants from atmospheric sources. They found a complete lack of lead uptake in these plants by this mechanism. Leland and Luoma (1979) report on a study that found 10 to 30 times more lead in algae grown on lead polluted snow, than in a control area. They also reported that duckweed lead bioconcentration factors were highest, when the lead concentrations in the water were the lowest. Ray and White (1976) reported lead concentrations from various aquatic plants collected from polluted and non-polluted streams. The plant tissue concentrations ranged from about 1 to 13 g/g for some plants in the clean water and ranged from about 5 to 570 g/g for other plants in the polluted stream reach. Leland and Luoma (1979) summarized feather moss lead concentrations ranging from 44 to about 310 g/g and lead concentrations of 5 to about 30 g/g for sphagnum moss. In a study of retention of lead in a bacterial food chain, Rolfe, et al. (1977) found that only 10 percent of the lead ingested by protozoa was retained in its cells and the remainder of the lead was found in a water-soluble form in the culture media.

Nickel


Wilber and Hunter (1980) found that the readily available nickel fraction of street dirt and runoff solids was about 4 percent at close to neutral pH conditions. Pitt and Amy (1973) found that the nickel solubility of street dirt solids, in a moderately hard water mixture, was about 30 g/L or about 7 percent of the total nickel in the mixture.

Mercury


Phillips and Russo (1978) reported that inorganic mercury concentration, availability of inorganic mercury, pH, microbial activity and redox potential all affect mercury methylation rates. In general, more methylmercury is produced when more inorganic mercury is present. Chemical agents which precipitate mercury, such as sulfide, reduce the availability of mercury for methylation, but only when present in large quantities. At neutral pH values, the primary product of mercury methylation is monomethylmercury. Methylation can occur under both aerobic and anaerobic conditions, but more mercury is produced when more bacteria are present. Therefore, highly organic sediments which favor bacterial growth have a higher methylation potential than inorganic sediments. Methylmercury is also strongly accumulated by organisms. Fish accumulated more mercury as the temperature and mercury content of the sediment increased. Bacteria not only act as methylators of mercury, but also accumulate large amounts of mercury. However, sediment and water are probably the two most important mercury sinks. Conditions reducing the mercury content of overlaying waters, such as the accumulation of mercury by aquatic organisms, result in the mobilization of mercury from sediment. Virtually any mercury compounds discharged to water may become a bioaccumulation hazard if the environmental conditions are favorable for methylation. Other microbial conversions of mercury have also been reported. Some bacteria are capable of transforming mercuric ion and phenylmercuric acetate to volatile mercury. Under certain conditions, however, the most toxic form of methylmercury can be demethylated.


Callahan, et al. (1979) also stated that almost all of the environmental processes are important when determining the fate of mercury in aquatic environments. The EPA (1976) reported that typical mercury concentrations in 31 states with no known mercury deposits are typically less than 0.1 g/L. Durum (1974) found that in 722 nationwide water analyses, the total mercury concentrations ranged from less than 0.5 to 6.8 g/L, with a median value of less than 0.5 g/L. He also reported dissolved mercury concentrations in 262 samples that ranged from less than 0.1 to a maximum of 4.3 g/L, with a median value of less than 0.1 g/L.

Essentially all animal tissues contain some mercury. Much information exists in the literature on mercury content of various animal tissues. In general, it is found that animals higher up in the food chain bioaccumulate higher amounts of mercury (EPA 1978). Methylmercury is bioconcentrated many times in fish and other aquatic organisms because of the rapid uptake of the methylmercury and the relative inability of the fish to excrete it from their tissues (EPA 1976). In addition, methylmercury appears to persist in the aquatic environment for sufficient time periods to allow uptake by aquatic organisms. Phillips and Russo (1978) reported a study that found more than 80 percent of the mercury that accumulated in mosquito fish was inorganic. Leland and Luoma (1979) also reported on studies that showed biomagnifications of 3 to 5 times for each step in a simple food chain. In a Georgia salt marsh food chain, plants had the lowest percentage of total mercury as methylated mercury. Herbivorous snails had the next highest percentage methylated mercury followed by benthic worms and mollusks, then crabs, fish and finally birds which had the highest percent of total mercury as methylated mercury in their tissues.

Phillips and Russo (1980) reported that fish can tolerate very high tissue concentrations of mercury. Fathead minnows exposed to about 0.1 g/L methylated mercury obtained methylated mercury concentrations in their eatable portions greater than the Food and Drug Administration’s action level (0.5 g/g mercury) without suffering adverse affects. Rainbow trout were reported to accumulate up to 30 g/g mercury without noticeable affects (about 60 times the FDA level). Phillips and Russo also reported that methylmercury is readily accumulated by fish both from food and from water. The biological halftime of methylmercury in fish is between 1 and 3 years. Leland and Luoma (1979) reported that a survey of museum fish collected over a period of 30 years indicated no detectable chronological accumulation of mercury in any species. The EPA (1976) reported concentration factors of 15,000 to 30,000 for methylmercury in fish with resultant tissue mercury concentrations of about 0.5 g/g. Leland and Luoma (1979) also reported that sporadic feeding of mercury to trout resulted in much greater mercury tissue concentrations than continuous feeding. Phillips and Russo (1978) summarized a report that studied fathead minnows in methylmercury concentrations of 0.018 to 0.025 g/L. The resultant minnow tissue concentrations ranged from 1.5 to 11 g/g after 48 weeks of exposure. In another study, mercuric chloride uptake in fathead minnows increased as the water pH decreased, with a sharp increase in uptake at pH values below 7. In another study, fathead minnows accumulated more mercury when their food source was also raised in the test water.

Phillips and Russo (1978) reported on a study of waterfowl mercury accumulation that resulted in waterfowl breast tissue mercury concentrations ranging from about 0.5 to 8 g/g. The waterfowl were sampled from a heavily polluted river. Fish and shellfish from highly polluted Minamata Bay in Japan contained 9 to 24 g/g mercury (EPA 1978).

Aquatic plants accumulate mercury primarily by surface adsorption (EPA 1976). Leland and Luoma (1979) reported mercury plant tissue concentrations ranging from 0.08 to 0.14 g/g in 23 aquatic plants collected in Finland, and a mercury range of 13 to 112 g/g in sphagnum moss from northern Canada.

The concentrations of mercury in invertebrates varies over a wide range. Leland and Luoma (1979) stated that benthic organisms accumulated mercury in their tissues through ingestion of material in the sediments. This mercury is then transferred to their fish predators upon ingestion (EPA 1976). Leland and Luoma (1979) reported mercury concentrations in crayfish from Wisconsin ranging from 0.07 to about 0.6 g/g.

Zinc


Durum (1974) stated that the solubility of zinc is less than 100 g/L at pH values greater than 8, and less than 1,000 g/L for pH values greater than 7, if there is a high concentration of dissolved carbon dioxide. Phillips and Russo (1978) stated that zinc sulfates and halides are soluble in water, but zinc carbonates, oxides and sulfides are insoluble. The EPA (1976) stated that zinc is usually found in nature as the sulfide. It is often associated with the sulfides of other metals, especially lead, copper, cadmium and iron. Callahan, et al. (1979) stated that zinc in unpolluted waters is mostly as the hydrated divalent cation (+2) but in polluted waters complexation of zinc predominates. Pitt and Amy (1973) reported that zinc is mostly found as the divalent form, as a sulfide, oxide, sulfate or hydroxide.


Wilber and Hunter (1980), in a study of an urban stream near Lodi, New Jersey, found that the readily available zinc in street dirt and runoff solids was about 17 percent of the total zinc. Most of the zinc in the river during low flow conditions was dissolved, while during wet weather it was mostly in the solid form. Pitt and Amy (1973) found that the solubility of zinc was about 170 g/L, or about 8 percent of the total street dirt zinc, in a moderately hard water mixture.

Durum (1974) reported zinc concentrations in 727 nationwide water samples ranging from less than 10 g/L to a maximum of 4,200 g/L, with a median value of about 20 g/L. The EPA (1976), in a nationwide survey of over 1,200 positive zinc results, found a mean value of 64 g/L and a maximum value of about 1,200 g/L.

Phillips and Russo (1978) summarized a report that found zinc had concentrated in the upper levels in a simple food chain consisting of sediment to bacteria to tubificid worms. They also reported that zinc is bioaccumulated in fish gills at a modest rate during chronic exposures, but rapidly during acutely lethal zinc exposures. In another study in Wisconsin, zinc concentrations in freshwater fish was found to range from 3 to more than 18 g/g. Studies of zinc bioaccumulation in rainbow trout showed that eyes accumulated the highest concentrations, followed by gills, bone, intestine, liver, kidney and finally skin. Baseline zinc levels ranged from 400 g/g for the eye to 1 g/g for the stomach. In another study, zinc was shown to bioaccumulate in the intestine of goldfish, implying that zinc is excreted through the intestine (Phillips and Russo 1978). In general, zinc begins to accumulate in fish at about the concentration where it becomes quite chronically toxic to the fish. Whole fish zinc uptake was higher in hard water for three spine stickleback, even though the zinc toxicity was much lower under the hard water conditions. Much of the zinc in the gill area could be suspended particulates imbedded on the gills. The half-life for zinc in brown bullhead appears to be about 6 days, after removal to clean water. The zinc half-life in juvenile mosquitofish have, however, ranged from 2 to more than 200 days, depending upon the fraction of the total zinc in the system accumulated within the fish. Rubin (1976) reports bioconcentration factors for zinc in fish of about 230, about 2,300 for mollusks and more than 300 for macrophytes.

Phillips and Russo (1978) reported a study that showed that benthic organisms such as clams and tubificid worms contained higher zinc concentrations than either omnivorous or carnivorous fish. Neff, et al. (1978) reported that concentration factors for macroinvertebrates were higher than for other test organisms and ranged from about 150,000 to 300,000. They also found that zinc tolerant worms were less permeable to zinc and excreted it more rapidly than non-tolerant worms. Phillips and Russo (1978) also reported on a study conducted in a polluted Colorado river that showed zinc concentrations in insects of up to 10,000 g/g. In another study, crayfish were found to bioaccumulate zinc through food, as their major source of zinc uptake.

Leland and Luoma (1979) reported zinc concentrations in feather moss ranging from 54 to more than 130 g/g and from 26 to 40 g/g in aquatic sphagnum moss. Ray and White (1976) reported zinc concentrations in various aquatic plants that were collected from polluted and unpolluted reaches of an urban creek. Zinc concentrations for plants in the clean reach of the creek ranged from about 100 to 3,000 g/g and from about 500 to 6,000 g/g for another plant species in the polluted stream reach. The concentration factors in the clean stream reach ranged from about 2 to 20 while they ranged from about 0.02 to 0.7 in the polluted reach.

1   ...   42   43   44   45   46   47   48   49   ...   64

Похожие:

Assessment Strategy for Evaluating the Environmental and Health Effects of Sanitary Sewer Overflows from Separate Sewer Systems iconEnvironmental effects of ozone depletion and its interactions with climate change: 2002 assessment

Assessment Strategy for Evaluating the Environmental and Health Effects of Sanitary Sewer Overflows from Separate Sewer Systems iconQuantitative Human Health Risk Assessment for 1,3-Butadiene Based Upon Ovarian Effects in Rodents

Assessment Strategy for Evaluating the Environmental and Health Effects of Sanitary Sewer Overflows from Separate Sewer Systems iconAged Care Assessment Service (acas) and Office of the Public Advocate (opa) protocol. Incorporating health professionals undertaking assessment and placement within health networks

Assessment Strategy for Evaluating the Environmental and Health Effects of Sanitary Sewer Overflows from Separate Sewer Systems iconAn evaluation of risk to u. S. Consumers from methylmercury in commercial fish products, including a quantitative assessment of risk and beneficial health effects from fish

Assessment Strategy for Evaluating the Environmental and Health Effects of Sanitary Sewer Overflows from Separate Sewer Systems iconSimulating and Evaluating Local Interventions to Improve Cardiovascular Health

Assessment Strategy for Evaluating the Environmental and Health Effects of Sanitary Sewer Overflows from Separate Sewer Systems iconFollowing guidelines established by the wln collection Assessment Service, which provide a framework for evaluating a library's current holdings and the level

Assessment Strategy for Evaluating the Environmental and Health Effects of Sanitary Sewer Overflows from Separate Sewer Systems iconFollowing guidelines established by the wln collection Assessment Service, which provide a framework for evaluating a library's current holdings and the level

Assessment Strategy for Evaluating the Environmental and Health Effects of Sanitary Sewer Overflows from Separate Sewer Systems iconSystem of design dokuments for construction. Elements оf sanitary engineering systems sumbols
Система проектной документации для строительства. Условные обозначения элементов санитарно-технических систем

Assessment Strategy for Evaluating the Environmental and Health Effects of Sanitary Sewer Overflows from Separate Sewer Systems iconCentre for Environmental Strategy, University of Surrey

Assessment Strategy for Evaluating the Environmental and Health Effects of Sanitary Sewer Overflows from Separate Sewer Systems iconEnvironmental Effects of Off-road Vehicles: Impacts and Management in Arid Regions

Разместите кнопку на своём сайте:
Библиотека


База данных защищена авторским правом ©lib.znate.ru 2012
обратиться к администрации
Библиотека
Главная страница