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Water Quality Ambient Water Quality Guidelines for Chlorophenols 4. THE FATE OF CHLOROPHENOLS IN THE ENVIRONMENT Once released into the environment, chlorophenols are subjected to a number of degradative processes. There are physical and chemical processes such as photodegradation, oxidation, hydrolysis, evaporation/volatilization and sorption, and biological processes such as uptake, breakdown and utilization. These all work with varying degrees of efficiency and speed, under different environmental conditions, to reduce chlorophenols to CO2, H2O and chloride, in times ranging from days to months. There are a great many different and complex organic compounds which may be intermediates in these breakdown processes. The nomenclature and structures of some classes of these intermediates are shown in Figure 4.1. In biologically active, buffered, lake water at pH 7.0 and 25C, 2,4-DCP added at 100, 500 and 1000 µg/L was degraded by 56% in 6 days; all of the 100 µg/L concentration was gone in 9 days. If the water was unbuffered and anaerobic, with high organic levels, then 2,4-DCP persisted for 43 days (345). Soil degradation rates of PCP are primarily a function of organic matter content; cation exchange capacity and pH are less relevant. Soil texture, clay content, base saturation and iron oxide were not correlated with degradation rates. Microbial action is the major cause; no degradation occurs without organics being present (606). The primary products of PCP degradation are 2,3,4,5-TTCP, 2,3,6-TCP and 2,4,6-TCP; other products include 2,3,4,6-TTCP, 2,3,5,6-TTCP, 2,3,5-TCP, 2,3,4-TCP and 2,4,5,-TCP (606, 53). Table 4.1 gives a summary of the fate of chlorophenols in water. This table is adapted from McKee et al. (220).
Photodegradation is only important in shallow water and in high insolation areas. Oxidation under natural conditions is not significant, nor is hydrolysis. Evaporation and volatilization are only important where there is shallow water and vigorous mixing . High temperatures also speed up the process. Adsorption to sediment or suspended particulate matter is an important process for chlorophenols in nature and most chlorophenols introduced into the environment will be found, eventually, in the sediments.
Photodegradation of chlorophenols depends upon the absorption spectrum of the compound; at sea level, sunlight does not contain wavelengths shorter than 290 nm. Since they are weak acids, chlorophenols exist as non-dissociated molecules or as anionic phenolates, depending upon the pH. The absorption spectrum is different for each species; ionized moieties are more susceptible to photodegradation than are the neutral molecules (337, 338, 339). The first step in the process is to split the C-Cl bond to form an alcohol, C-OH. In water of any depth, or with high turbidity or color, the energetic short wavelengths are all absorbed in the upper portion of the water column; photolysis is only a surface phenomenon. The rate may be significant below 280 nm, and/or in alkaline solutions, but is small in waters where the pH is neutral or acidic, and at longer wavelengths. 2-MCP Molecular 2-MCP is converted to pyrocatechol; anionic 2-MCP to cyclopentadienoic acid dimers (335). 3-MCP Molecular and anionic 3-MCP is converted to resorcinol (335). 4-MCP Anionic and molecular 4-MCP is transformed to hydro- and benzo-quinones, trihydroxybenzenes and dihydroxybiphenyls (335). 2,4-DCP Photolysis is a major transformation process for 2,4-DCP in estuarine waters (242), and occurs readily (221, 210), but is much slower in the winter. The breakdown of 2,4-DCP is more complex than that of the monochlorophenols and leads ultimately to chlorinated dimers (335). Tests using a 326 mg/L solution of 2,4-DCP, 5 hours/day of noon sunlight for 10 days, or a mercury arc lamp at 254 nm for 5 to 12 hours, and 20 to 25C temperature, gave 50% losses in 5 minutes at pH 7 in the lab. The decomposition pathway was presumed to be 4-chlorocatechol or 4-chlor-1,2-diol-benzene to 1,2,4-benzenetriol and thence to hydroxybenzoquinone and polymeric humic acids (518). At 42.4 mg/L 2,4-DCP in 90 mL, 4 mm deep vials, using a 600 watt mercury discharge lamp, the time to achieve 50% decomposition was pH dependent. It was 34 minutes at pH 4, 4 minutes at pH 7 and 2 minutes at pH 9. At pH 7 with an initial concentration of 36 mg/L, only 12% was left after 80 minutes (345). 2,4,5-TCP Photolytic degradation of 2,4,5-TCP is fairly rapid and the suggested pathway is to 2,5-DCP or 2,4-dichloro-resorcinol. These then go to 4-chlororesorcinol and ultimately to humic acids (493). In estuarine waters photolysis is a major transformation process for 2,4,5-TCP (242), occurring very readily (221, 210), but much reduced in the winter. 2,4,6-TCP In the presence of an electron acceptor, 2,4,6-TCP can be broken down to 2,6-dichlorophenoxyl semiquinone radical ions; 66% in 17 hours under UV light at 290 nm (597). 2,3,4,5-TTCP Photoreactions of 2,3,4,5-TTCP, at 285 nm in water and acetonitrile, occur through reductive dechlorination. The 4-position is photo-labile, but 3- and 5-position chlorines are not photoreactive. The degradation of 2.3 mg/L of 2,3,4,5-TTCP was 30% complete in 6 hours and 73% after 24 hours. There was 10% 2,3,5-TCP in the products after 6 hours and 9% 2,3,5-TCP after 24 hours (457, 524). 2,3,4,6-TTCP Six hours of irradiation at 285 nm destroyed 51% of the initial 1.2 g/L of 2,3,4,5-TTCP. The breakdown products were pH and time dependent but included 2,3,4- and 2,3,6-TCPs at about the one percent level. The 4 and 6 chlorines were photo-labile, but not the 3 position chlorine (457). PCP Photolysis is a major transformation process for PCP in estuarine waters (242) and occurs very readily (221, 210, 473), but is much reduced in the winter months. In the clear, shallow waters of Transvaal, PCP is a poor molluscicide, but in Egyptian snail-infested, turbid, irrigation ditches, PCP works very well (210). This is because the PCP is rapidly degraded by sunlight in the clear waters. Aqueous solutions of PCP are degraded by sunlight or laboratory ultraviolet lamps; 90% of the starting concentration is broken down in 10 hours at pH 7.3, but only 40% in 90 hours at pH 3.3 (20, 221). A residence time of 2 days is estimated for lakes, where PCP can be degraded by photolysis to 0.1% of the original concentration (321). The products include chlorinated phenols, catechols and benzoquinone; tetrachlororesorcinol and tetrachlorohydroquinone; dichloromaleic acid (289), and octachloro-dibenzodioxin (294, 295, 296, 221). Under laboratory conditions, decomposition of PCP in aqueous solution by sunlight is rapid. A starting concentration of 9.3 mg/L went to 0.4 mg/L in 24 hours and nearly to zero in 48 hours. In effluents and natural waters, other compounds screen out much of the UV. In a rice paddy, 2.4 mg/L PCP dropped to zero in the water in 7 days and from 4 mg/L to 2 mg/L in the soil. In waste from a wood treatment plant, PCP went from 30 mg/L to 0.2 mg/L in 7 days. The products of solar photochemical breakdown include tetrachloro-p-benzoquinone, 3,4,5-trichloro-6-(2-hydroxy-3',4',5'-tetrachlorophenoxy)-0-benzoquinone, and tetrachlororesorcinol. Dechlorination to 2,3,4,6- and 2,3,5,6-tetrachlorophenols is followed by hydroxylation to 3,4,5-6-tetrachlorocatechol, tetrachlororesorcinol and 2,3,5-6-tetrachlorohydroquinone. Further irradiation yields dichlorocyclopentanedione and 2,3, dichloromaleic acid. Ultimate products are H2O, CO2 and HCl (191). Irradiating NaPCP at high concentrations in aqueous solution leads to formation of octachlorodibenzo-P-dioxins (205). The approximate 1/2 life for PCP undergoing photolysis during the summer in California is 1.5 days (16). In a 21 day photolysis study in water between 7.5 and 36.3 degrees centigrade, PCP degraded rapidly under natural sunlight with half-lives of 20, 16.3 and 13.6 minutes at pHs of 5, 7 and 9 respectively. The degradation products included tetrachlorohydroquinone, tetrachlorocatechol, tetrachlororesorcinol and dichloromaleic acid or anhydride (725). A 30 day soil photolysis study on sandy loam at v24 degrees centigrade was conducted under alternating 12 hour light and dark cycles. No degradation occurred in the dark phase. The half life in the light was 37.5 days based on first-order degradation kinetics (726). Vapour phase PCP was seen to degrade under simulated sunlight in alternating 12 hour light and dark periods with a half-life of 36.6 days in the light phase. Degradation products included 2,3,5,6-TTCP (727). Photolysis is likely responsible for increasing TTCP: PCP ratios downstream from PCP effluents and in surface waters compared to deeper waters as the PCP is degraded to TTCPs (291). Photolytic half-lives of PCP are about 4 hours at pH 7.3 and 100 hours at pH 3.3. Light in the 290 to 330 nm range is effective at photolysis which proceeds most rapidly at high levels of light, pH and dissolved oxygen (221, 20, 210, 159). Since light attenuation occurs with depth, particularly UV attenuation, photolysis is primarily a surface phenomenon and 1 meter of depth has about the same effects as a drop of four pH units, or about 24 times the half-life.
There are few good data concerning chlorophenol oxidation. Chlorophenol decay is found in sterilized natural water under conditions where photodegradation and volatilization/evaporation are not likely. The disappearance rate is concentration dependent and increases with temperature. Phenols are susceptible to auto-oxidation and catalysis occurs on the surface of clay and silica in the presence of several metal ions (18, 19, 341, 342, 343, 344). Highly chlorinated organics are usually resistant to oxidation at temperatures well above normal aquatic-habitat ambient levels. Oxidation is not expected to be important in PCP breakdown in nature (473, 598, 340). 2-MCP and 4-MCP Oxygen and hydroxyl radicals attack 2-MCP at the C-2 or C-4 sites to form 1,2- or 1,4-benzosemiquinones, and 4-MCP at the C-4 site to form 1,4-benzosemiquinone (335).
There are few data on chlorophenol hydrolysis, which is, in any case, an unlikely transformation in the environment. Aryl halides, especially those containing hydroxyls (which includes all chlorinated phenols), have low reactivity to halogen displacement or nucleophilic substitution (340). The covalent bond of a substituent on an aromatic ring is resistant to hydrolysis because of the high negative charge density of the aromatic nucleus. Making the pentachlorophenolate anion from hexachlorobenzene requires treatment with concentrated alkali at 130 to 200C. Further hydrolysis should require more extreme conditions and thus is not important under ambient conditions in aquatic habitats (473). 2,3,4-TCP, 2,3,5-TCP, 2,3,4,5-TTCP, 2,3,4,6-TTCP and 2,3,5,6-TTCP The hydrolysis half-lives of these five chlorophenols are about 3 years or more; hydrolysis is not an important natural transformation process (428, 457). PCP Hydrolysis is not a factor in PCP breakdown at pH 7 or 9 and is negligible at the lower pHs of 4 and 5. The calculated half lives, from 80 to 160 days, at these low pHs were likely due primarily to volatilization losses at the 25 and 35 degree centigrade temperatures of the experiments, which included air flow over the solutions (728).
Evaporation is inversely proportional to water depth (336). TCPs, which are soluble and have low vapor pressure, do not generally volatilize from water. They are moderately acidic and will be ionized and solvated in natural waters. Volatilization is likely insignificant in natural environments (598). Volatilization is a function of temperature, water solubility, vapor pressure, solution mixing depth and molar concentration (607). Chlorophenols are quite persistent under ice cover in Canadian rivers during the winter (706). 2-MCP and 4-MCP In stirred solutions the half-life of 2-MCP at 0.38 cm depth is 1.5 hours; at 1 m depth the half-life is 15 days. For 4-MCP the values are 13 hours at 0.38 cm depth and 4.5 days at 3 cm depth (336). 2,4-DCP The volatilization half-life of 2,4-DCP at 64 to 100% relative humidity, was about 1 month at 30C (510). 2,3,4-TCP and 2,3,5-TCP The Neely 100-day partitioning pattern for 2,3,4-TCP and 2,3,5-TCP is air-68.45%, water-12.15%, ground-10.03% and hydrosoil-9.37%. Chemicals with these properties are expected to vaporize rapidly from, and not persist in, open water (428). 2,4,6-TCP The volatilization of 2,4,6-TCP from water and soils at 25C was measured as a percent 2,4,6-TCP per mL of water evaporated. For the first few hours, rates were about 1.5% for water and 1% for sandy soils, dropping to 0.1% for humic soils (462). Volatilization is not a significant process in nature (598). 2,3,4,6-TTCP and 2,3,5,6-TTCP The Neely 100-day partitioning pattern for 2,3,4,6-TTCP is air-63.95%, water-6.31%, ground-15.38% and hydrosoil-14.36%. For 2,3,5,6-TTCP the respective percentages are air-23.86%, water-28.23%, ground-24.78% and hydrosoil-23.13%. Chemicals with these partitioning coefficients, and a log Ko/w value around 4.5, would be expected to vaporize rapidly from open water and accumulate in the soil (428).
PCP The volatilization rates of PCP from water, sand, loam and humus
were 2.57%, 0.13%, 0.31% and 0.10% in the first hour, and 2.11%,
0.12%, 0.15% and 0.12% in the second hour. These rates are expressed
as the percent of applied PCP per mL of water evaporated (581).
Volatilization decreases abruptly with increasing pH. At pH 8
virtually no volatilization occurs. The half-life declined from
130 days at pH 6, to 6 days at pH 4 (607). All the MCPs adsorb to about the same extent; increasing the number of chlorine atoms results in increased sorption, likely due to extra hydrogen bond formation (246). Sorption in river conditions gave concentration ratios of 117 for suspended matter and 440 to 650 for sediments, when water concentrations were 1.8 to 10.0 µg/L (348, 349). Desorption is slower than sorption and is not 100%; a fraction is irrevocably held by the sediment particles. Sediment levels may thus remain high longer than water levels (347), but this gives microorganisms, which are more densely aggregated on sediments, a greater opportunity to break down the chlorophenols. Adsorption is a function of the partition coefficients of the chlorophenols between the water and the solid phase, sediments or suspended particulates (707, 708, 709). Chlorophenols are strongly adsorbed by activated carbon at the µg/L level, which is around the odor threshold for most chlorophenols. Adsorption is a function of pH (707, 708, 709); neutral species, which predominate below the pKa, are adsorbed more strongly than anionic species. While increasing the number of substituted chlorine atoms increases adsorption, it also lowers the pKa. The competition for adsorption sites among a group of chlorophenols is thus pH dependent, and those whose pKa is higher than the given pH, will be preferentially adsorbed, displacing those chlorophenols whose pKa values are lower than the ambient pH (506). Sorption is generally a function of the organic matter in the sediment (707, 708, 709); clays are not as effective. Soil moisture levels do not affect adsorption (507, 508). When natural ponds were dosed with PCP at a total application of 0.3 mg/L, the concentration in the water decreased rapidly and exponentially. Sediment levels rose slowly for 28 weeks then slowly declined. Leaching into surrounding soils did not occur (583). The sorption of some chlorophenols by Bentone 24 and Bentone 18C, Wyoming Bentonite Clays is shown in Table 4.1.5.1
2,4-DCP Sediment enrichment factors for the Rhine River were 440 times the water concentration, and sediment concentrations were generally in the µg/kg dry weight range (349). 2,6-DCP Sediment enrichment over the water column in the Rhine River
is 20 times (349). All six TCP isomers were measured in Rhine River mouth, 13 water samples and 17 sediment samples. TCPs accumulate in the sediments with various enrichment factors from five to 61 as shown in Table 4.1.5.2 adapted from reference (349). 2,3,4-TCP In Nebraska sandy loam soil this compound is tightly bound and some irreversible binding may occur based on the high Ko/c values obtained (730). 2,4,6-TCP In a contaminated lake in Finland, 2,4,6-TCP reached 10.4 to 17.2 µg/kg dry weight in the sediments (599). TTCPs All three TTCP isomers were measured in water and sediments from the mouth of the Rhine River. There were 13 water samples and 17 sediment samples. TTCPs accumulated in the sediments with enrichment factors of 35 to 445 as shown in Table 4.1.5.2, adapted from reference (349). As expected from the log Ko/w values of TTCPs compared to TCPs, the mean enrichment of TTCPs is greater than that of TCPs. 2,3,4,5-TTCP A log Ko/c value of 3.84 indicates a high potential for partitioning to soil. The Neely 100-Day Partitioning Pattern for 2,3,4,5-TTCP is air-39.66%, water-6.05%, ground-28.08% and hydrosoil-26.21%. With a log Ko/w of 4.7, a chemical with these properties would not persist in water; most would end up in soils and sediments (428). In Nebraska sandy loam soil this compound is tightly bound and some irreversible binding may occur based on the high Ko/c values obtained (730).
2,3,4,6-TTCP The log Ko/c, a calculated value of 3.69, indicates that this compound will bind to soil (428). 2,3,5,6-TTCP A log Ko/c value of 3.45 indicates a good potential for this compound to partition to soil (428). The Neely 100-day partitioning pattern for 2,3,5,6-TTCP is air-23.86%, water-28.23%, ground-24.78% and hydrosoil-23.13%. Based on Henry's Constant, 0.00562, and a log Ko/w of 4.5, a chemical with these properties should vaporize rapidly from open water with large amounts partitioning to sediments (428). PCP PCP is sorbed on acidic soil primarily; little if any sorption occurs on neutral soil. The pH is the most important factor controlling sorption and the total amount sorbed is a direct function of organic content of the soil. Sediments and leaf litter retain high levels of PCP and serve as a sink for continual PCP input to the water. Rainfall both leaches PCP from soil and transports leaf litter to lakes (473, 580). PCP binds strongly to soil and extensive leaching experiments rarely desorbed over 10 to 20% of the amount present (13, 191). A study on the sorption and desorption of PCP in Georgia sandy loam, Ohio clay loam, California sandy loam and Nebraska blue sandy loam indicate that PCP is tightly bound or immobile in all but the California sandy loam where it is moderately bound, based on its Koc value (729). Enrichment factors in sediment over the water column ranged from 960 to 3600 in the Bay of Quinte, downstream from a wood preservation facility (291). In the lower Rhine River, PCP was found in all 17 sediment and 13 water samples. The mean sediment enrichment factor was 20 (340). Organic material is a strong sink for PCP and the inhibition of degradation under the anaerobic conditions and low pH often found in such sediments leads to high PCP levels in anaerobic, low pH, high organic sediments such as are found in eutrophic conditions. The rate of PCP dissipation in soil is related to temperature, aeration and organic matter, partly dependent on cation exchange capacity and pH, and independent of texture, clay content, base saturation and free iron oxides. Degradation is usually slower in flooded or anaerobic soils than aerobic moist soils. Where the adapted microflora are anaerobes, and there are no adapted aerobic organisms, the reverse may occur. Anaerobic microorganisms use a different degradative pathway than the aerobes. Degradation is primarily by reductive dehalogenation to TTCPs, TCPs and DCPs. Methylation to the anisole also occurs (584).
Microbiological breakdown is the major mechanism for removal of chlorophenols from the environment. In the absence of more readily metabolized compounds, some microorganisms use chlorophenols as their sole source of carbon for growth. Higher organisms also take up chlorophenols, accumulate them, detoxify them in various ways, and excrete chlorophenols and various breakdown products. All organisms are able to adapt to higher concentrations of chlorophenols over time, becoming better able to break them down as their metabolic processes adapt in response to the continued presence of chlorophenols. In higher organisms the adaptation is at the metabolic or enzymatic level; in bacteria, with their shorter life-cycles, the whole genetic pool of the population shifts so that chlorophenol metabolizing strains predominate. These adaptation processes take time and response to a first chlorophenol exposure is relatively slow, but once adapted, response to subsequent exposures is rapid.
In spite of their extensive use as biocides, the chlorophenols are quite readily broken down by a wide variety of microorganisms, and used as a growth substrate by a number of bacteria. The addition of chlorine to phenol decreases the rate of biodegradation; the more chlorine added, the slower the degradation rate. Biodegradation may be a slow process which only occurs aerobically when other organic carbon sources are not available (184, 185), or may occur rapidly with half-lives up to days (709). Phenols substituted in the 2,4 and 6 positions are more readily degraded than 3-chloro or 5-chloro compounds (18, 19, 49, 159, 292, 293, 350). Monochlorophenols, particularly 3-MCP (18), are more resistant to degradation than many poly-chlorinated congeners (18, 49). The sediment/water interface is the most important site of degradation since the responsible bacteria are generally attached to surfaces (297, 298). Adsorption to soil is pH dependent, more occurs at lower pH and in soils of high organic or humic content. Degradation rates increase with soil moisture and decrease with clay content (175). Degradation rates are a function of the conditions suitable for the growth of the responsible organisms (709). The aeration, temperature, pH, organic content, moisture level, nutrient level, light intensity and duration, and a cation exchange capacity conducive to microbiological growth, encourage the degradation of chlorophenols (48, 289). The optimum levels of these parameters are quite variable and depend upon the organism which is metabolizing the chlorophenol. In addition the concentration of the chlorophenol is also important. Toxicity occurs at high levels and at low levels other carbon sources may be preferentially metabolized. The evidence for fungal and bacterial degradation of the chlorophenols is widespread (14 to 19, 21, 49 to 55, 434, 700). Many genera of bacteria have been shown to break down chlorophenols; these include, Alcoligenes, Aeromonas, Azotobacter, Flavobacterium, Pseudomonas, Cytophaga, Corynebacterium, Arthrobacter, and Brevibacterium (21, 434). Additional organisms which break down PCP include Trichodera viride and Penicillium molds, Coriolus fungi and the peroxidase enzymes produced by some white rot fungi such as Phanerochaete chrysosporium. The chlorophenols are ultimately broken down to chloride and CO2 plus organic fragments, as indicated by work with isotopically labeled carbon. Ring cleavage occurs and the organic fragments are ultimately processed to free CO2. Microbiological breakdown ensures non-persistence in soil, sediment or sewage effluent (155, 191); only continuous input, at rates higher than degradation rates, makes it possible to find substantial quantities of chlorophenols in the aquatic environment where half-lives are days for MCPs and DCPs, weeks for TCPs and months for TTCPs and PCP (220). Much of the following discussion is tabulated in Table 4.2.1. Table 4.2.1 shows the rates of degradation of some chlorophenols in sterile and non-sterile soils and indicates that most degradation is biological. It also gives degradation rates for some chlorophenols in sewage sludge under different pH and temperature regimes. One must remember that soil and sediment bacterial populations, which have not previously been exposed to chlorophenols, will show an appreciable lag phase between the application of chlorophenols and the emergence of significant breakdown products. Initially, very few cells will have the necessary metabolic machinery to survive in, and utilize, the chlorophenol. It takes time for these few cells to multiply and become a significant proportion of the bacterial population and also a significantly large number of cells in absolute terms. Once a large population of adapted cells exists, degradation will proceed rapidly. Subsequent additions of chlorophenols within a short period of time will not require this lag time before breakdown products begin to be seen (351, 352, 353, 354, 438, 710). The pelagic bacteria of oligotrophic lakes which have high humic levels, are an exception to the usual need for lag time. Bacteria living in such habitats already have enzymes adapted to degrading the aromatic compounds of humus and can degrade the structurally similar chlorophenols (700). After adaptation of the microflora, which takes about 2 weeks, complete biodegradation occurs in only a few days. Using activated sludge from waste treatment plants, the lower chlorophenols are readily degraded, see Table 4.2.1, though at high concentrations of 20 to 100 mg/L toxicity occurs and activity is reduced (350, 351). With pre-adapted microflora degradation is rapid through to complete mineralization, CO2, H2O, and chloride, (185, 355, 356). Most degradation studies only monitor the loss of the parent compound. Complete mineralization may not occur and then metabolites may accumulate in the environment. Due to losses from various sources, experimental results reported for nominal or starting concentrations are actually caused by lower levels of chlorophenols in solution. Thus, most published LC50 data are a little optimistic; the actual solute levels were a little lower than the nominal amount added.
2-MCP Water from the Vistula River in Poland degraded 2 to 20 mg/L of 2-MCP in 2 to 13 days, after 2 to 3 weeks of adaptation. Adding adapted bacteria from a previous trial increased the rate of removal (351). In the water column of a small stream, degradation of 2-MCP at 100 mg/L was negligible in 40 days, even at 20C. In aerobic sediments at 0C and 4C, bacterial degradation of 2-MCP occurred more quickly than that of 3-MCP and PCP. Complete removal occurred in 10 to 15 days at 20C, and at the lower temperatures 78% was removed in 30 days (19). Activated sludge will completely break down 100 mg/L of 2-MCP in 3 days; both ring cleavage and dechlorination occur (184). Anaerobic conditions delay biodegradation (18), but anaerobic sewage sludge degrades 50 mg/L in 3 weeks by a different biochemical pathway than that used in aerobic degradation. One mg/L of 2-MCP added to domestic sewage at 20C was not degraded in 20 to 30 days, presumably due to a lack of adapted bacteria in this medium. In polluted river water, degradation occurred in 15 to 23 days (315). In soil, 250 mg/L of 2-MCP took 10 days to be reduced to 34%. Subsequent additions degraded at twice the rate. In a 7-day test, the rate of disappearance was twice as fast in unsterilized soil as opposed to sterilized soil (438). Oxidation of 2-MCP to 3 or 4-chlorocatechol by soil and bacteria was found for three strains of Norcardia, three strains of Pseudomonas and one strain of Bacillus and for Mycobacterium coeliacum (439). In shake culture 2-MCP at 50 µg/L was mixed with 4 g of two types of soil at 30C. Complete degradation occurred in 14 and 47 days for the two soil types (350). The degradation rate of 2-MCP was studied at several concentrations
in sludge pre-acclimated to 3-MCP In aerobic sediments, at both 10C and 4C, bacterial degradation of 3-MCP occurred more slowly than that of 2-MCP, 4-MCP and 2,4-DCP. In the water column degradation was negligible, even at 20C. In sediments at 20C, 42% of 100 mg/L was degraded in 30 days (19). In a shake-culture experiment 4 g of soil in 100 mL of medium at pH 7.2 and 30C was inoculated with 3-MCP. When 50 µg/mL (50 mg/L) of 3-MCP was added to two different soils the time to complete degradation was >72 and >47 days (350). Degradation is very slow, if it occurs at all, in otherwise suitable conditions when acclimated bacteria are not present. Chlorine substitutions in the "meta" position slow down the rate of degradation considerably (18, 350). Well-acclimated bacteria, given no other carbon source for several weeks, were able to degrade 100 mL /L in 2 days at 27C (355), but un-acclimated anaerobic sewage sludge took 7 weeks to completely degrade 50 mg/L (353). In activated sludge 100% ring degradation and 100% chloride ion production was reported in 3 days after an initial inoculation of 100 mg/L of 3-MCP (184). The soil bacteria Norcardia -three strains, Pseudomonas -three strains, Bacillus -one strain and Mycobacterium coeliacum were able to oxidize 3-MCP to either 3- or 4-chlorocatechol (439). In activated sludge from treatment plants previously exposed to phenols, the degradation of 3-MCP in 6 hours was 100% for 1 mg/L, 40% for 10 mg/L and none for 100 mg/L (440).
4-MCP Stream water did not appreciably degrade 100 mg/L of 4-MCP between 0C and 20C. In 30 days, the aerobic sediments removed all the 4-MCP at 20C and 60% of the 4-MCP at 0C, which was faster than the removal rate for 3-MCP or PCP (19). In river and pond water, bacterial transformation rates are about 0.007 µg/L /hour at a concentration of 1 mg/L (358). In Skidway Estuary water at 21C, the half-life of 25 µg/L 4-MCP was 20 days for complete mineralization; at 9C, no CO2 was evolved (352). In North Sea coastal plankton communities, initial concentrations of 0.1 and 1 mg/L of 4-MCP disappeared in 5 and 14 days, respectively. Subsequent additions disappeared more rapidly indicating adaptation and biological degradation were responsible (353, 354). The effect of 4-MCP on urease activity and the loss rate of 4-MCP by bacterial degradation were studied in a clay with 3% organic content, at pH 7.3. Starting with 500 µg of 4-MCP (50 µg/g in 10 grams of soil) urease activity decreased by 30%. After 3, 7 and 14 days the inhibition dropped to 21%, 10% and 3%. On loam soils, the initial depression of urease activity was 37% (447). Soil decomposition of 4-MCP was tested by pre-treating soil for 7 days with 13.3 mg/L 4-MCP, then adding 22.5 mg/L. In 23 days this dropped to 13.0 mg/L (438). In shake culture at 30C using 50 mg/L of 4-MCP and 4 g of soil, total loss times of 9 and 3 days were recorded for two soil types (350). Activated sludge degraded 100 mg/L in 3 days. Ring degradation and chloride ion development, both to 100%, occurred in this time (184). Soil bacteria, three Norcardia strains, one Pseudomonas, one Bacillus and Mycobacterium coeliacum, oxidized 4-MCP to 3 or 4-chlorocatechol (439). In sewage sludge from plants with continuous incoming phenol levels of 0.1 to 0.35 mg/L for a year, there was 100%, 80% and 16% degradation of 1, 10 and 100 mg/L of 4-MCP, respectively, in 6 hours (440). In anaerobic sewage sludge, no significant mineralization of 50 mg/L occurred in 8 weeks (357). 2,4-DCP In aerobic sediments, at both 0C and 4C, bacterial degradation of 2,4-DCP occurred more quickly than that of 3-MCP or PCP. In water at 20C, 2,4-DCP was broken down slowly, but not at 0C and 4C. The half-life of 100 mg/L in the sediment at 20C was 15 to 30 days, and in water at 0C, the test was terminated after 40 days, at which time no appreciable breakdown had occurred (19). Aerated and buffered lake water removed 0.1 mg/L in 9 days, and 0.5 to 1 mg/L in 30 days. The same rates of loss were found in North Sea water (354). In shake culture at pH 7.2 and 30C, 50 µg/L of 2,4-DCP was completely degraded in 9 and 15 days by two types of silt loam soil (350). Arthrobacter sp. enzymatically dechlorinates 2,4-DCP completely in 4 hours (514). The soil fungus, Aspergillis clavatis, degrades 2,4-DCP at concentrations up to 10 mg/L. At higher concentrations growth is inhibited (515). The degradation of 2,4-DCP by microorganisms proceeds through catechol intermediates to succinic acid (441, 516, 517). In acclimated activated sludge, 2,4-DCP at initial concentrations of 50, 100, 200 and 400 mg/L dropped in 2 days at pH 7 and 26C, to 20, 25, 60 and 45% respectively. At 100 mg/L, 100% degradation occurred in 5 days (184).
2,5-DCP In activated
sludge there was 52% ring degradation in 4 days with an initial
concentration
of 100 mg/L. This was acclimated
sludge at pH 7 and 26C (184). In shake culture with an artificial
medium which had been previously sterilized, 50 µg/L 2,5-DCP
was added along with 4 g of silt-loam soil. At 30C and pH 7.2,
the 2,5-DCP persisted through to the end of the experiment at
72 days (350). Adding 5 mg/L of 3,5-DCP to wastewater stimulated oxygen consumption rates; 25 and 50 mg/L depressed oxygen uptake. Sublethal growth effects, as determined by the fermentation tube method, were seen at 3 mg/L (1.3 to 5.0) and the LC50 was 11 mg/L (3 to 20). The onset of the inhibition of cell division in Pseudomonas occurred at 1.9 mg/L (1.25 to 2.50) (444). 2,3,4-TCP Pseudomonas cepacia (AC1100), adapted to 2,3,4-TCP, dehalogenated the 2,3,4-TCP at a rate of 70% in 3 hours when it was present at up to 20 mg/L. When present at concentrations of 40 mg/L the degradation rate decreased (434). 2,3,5-TCP Pseudomonas cepacia (AC1100), dehalogenates 2,3,5-TCP, but it is a slow process and concentrations above 0.2 mM, (39.5 mg/L) are toxic (434). At 30C Pseudomonas takes 100 hours to complete ring cleavage and breakdown of 2,3,5-TCP (424). 2,4,5-TCP Microorganisms in soil suspensions convert 2,4,5-TCP to 3,5-dichlorocatechol, 4-chlorocatechol, succinate, chlorosuccinate, Cis,Cis-2,4-dichloromuconate and 2-chloro-4-(carboxymethelene) but-2-enolide (491). In shake culture at pH 7.2 and 30C, 50 µg/mL of 2,4,5-TCP was added to 4 g of each of two different types of silt loam soils. The time for complete disappearance was greater than 72 and 47 days for the two soils (492). With no adapted microorganisms present, chlorophenol breakdown can be very slow. In an aeration lagoon where the effluents were maintained at 20 to 21C, 18.8 mg/L of 2,4,5-TCP was completely degraded in 7 days (315). In a freshwater nutrient medium, 70% biodegradation of 2,4,5-TCP occurred in 35 days starting with 1 mg/L; in a biological sewage treatment system, no degradation of 2,4,5-TCP was seen in 14 days (600, 601).
2,4,6-TCP A gram-variable bacillus was isolated which could utilize 2,4,6-TCP as its sole carbon source. Growth occurred at 198 mg/L and in 84 hours all the 2,4,6-TCP was used and cell counts rose over 100 times. The minimum inhibitory concentration was over 400 mg/L. At a 2,4,6-TCP/cell ratio of about 400 µg/mg, and 150 minutes incubation, all the 2,4,6-TCP was used and 64% of the chloride had been released (51). Slime organisms, similar to activated sludge bacteria, were grown on normal substrate and then switched to 2,4,6-TCP. In 3 days 75% was degraded as measured by chloride ion production (461). In bacterial cultures, starting with 300 mg/L, 95% degradation occurred in 7 to 10 days; starting with 100 mg/L, 70% degradation occurred in 3 hours (602). Using an initial concentration of 1 mg/L of 2,4,6-TCP, a freshwater nutrient medium degraded 70% in 9 to 18 days (600). In soil culture, 5 to 13 days were needed for complete removal of 2,4,6-TCP, but in acclimated sludge, complete aromatic ring degradation occurred in 5 days (473). In flask cultures with sludge bacteria, it took 7 to 10 days to remove 95% of 300 mg/L of 2,4,6-TCP; at 100 mg/L, 70% was removed in 3 hours. Acclimated sludge can carry out complete aromatic ring degradation of 2,4,6-TCP in 5 days (184). In soils, complete disappearance takes from 1 to 9 days (350). In soil suspension, 2,4,6-TCP disappeared completely in 5 days. Degradation by the parent strain of Pseudomonas took 120 hours at 30C for 100% ring degradation of 200 mg/L. A mutant strain took only 50 hours (424). In shake culture with two loam soils, 50 µg/L of 2,4,6-TCP was completely degraded in 5 and 13 days (350). 3,4,6-TCP The growth of a gram-variable bacillus was supported only by 3,4,6-TCP and 2,3,4,6-TTCP; none of the other chlorophenols would support growth (603).
2,3,4,5-TTCP Pseudomonas cepacia (AC1100), grown on 2,4,5-T as its sole carbon source, was able to dechlorinate 2,3,4,5-TTCP to a limited extent (434). In a continuous-flow-sludge-blanket-reactor, 300 µg/L of 2,3,4,5-TTCP was dechlorinated, at the rate of 90 to 99% . Even at up to 600 µg/L, yeasts were present in the flocs but died after 5 days at 1 mg/L. Increasing the TTCP from 2.4 mg/L to 8 mg/L caused drastic decreases in the rate of chlorophenol removal. Bacteria identified included Klebsiella oxytoca, K. pneumonia, Aeromonas hydrophila and Pseudomonas aeruginosa. One of the yeasts was a Candida sp. (523). In non-sterile soil, after 80 days at pH 7.1 and 23C, only 5% of the added 1 mg/mL of 2,3,4,5-TTCP was decomposed, and 31% decomposition occurred in 160 days. No decomposition occurred in sterile soil (18). A wastewater treatment biomass acclimated to PCP did not readily degrade 2,3,4,5-TTCP at 40 mg/L (458).In soil, 2,3,4,5-TTCP degrades to 2,3,5-TCP, 2,4,5-TCP, 3,4-DCP and 3-MCP (606, 17). 2,3,4,6-TTCP Fungi and bacteria, extracted from litter in broiler houses, were grown in flasks and treated with 2,3,4,6-TTCP. Of the 116 isolates tested, 99 metabolized 2,3,4,6-TTCP; 68 produced 2,3,4,6-tetra-chloroanisol as a metabolic product. Penicillium corylophilum produced the highest amount of methylated product and removed the 2,3,4,6-TTCP in 2 days. P. brevicompactum did not form the chloroanisole and took 8 days to remove the chlorophenol (534). Decomposition, as measured by ring cleavage, not dechlorination, of 2,3,4,6-TTCP in soil suspensions, took longer than 72 days for complete disappearance (424, 350). Pseudomonas cepacia (AC1100) was grown on 2,4,5-T as the sole carbon source. Tests were then run to determine its ability to dehalogenate halophenols, including 2,3,4,6-TTCP. At 23.2 mg/L for 3 hours 86% dechlorination occurred; above 46.4 mg/L the rate decreased due to toxicity (434). Wastewater treatment biomass, acclimated to PCP, was tested for its ability to breakdown 2,3,4,6-TTCP. In flasks at pH 7.8, 8.5 mg of 2,3,4,6-TTCP was broken down, per gram of cells per hour, from an original concentration of 40 mg/L (458). The growth of a gram-variable bacillus was supported by 2,3,4,6-TTCP and 3,4,6-TCP, but not by any of the other chlorophenols (603). In soil, 2,3,4,6-TTCP degrades to 2,4,5-TCP (17,606). 2,3,5,6-TTCP Pseudomonas cepacia, grown on 2,4,5-TCP as the sole carbon source, was able to dechlorinate 2,3,5,6-TTCP by 94% in 3 hours at 23.2 mg/L. The growth medium was between pH 7 and pH 8 and strain AC1100 was used in the tests (434). Mixed culture cells taken from a wastewater fiber wall reactor were suspended in phosphate buffer and 40 mg/L of 2,3,5,6-TTCP was added. The breakdown rate was 3.2 mg/gram of cells/hour (458). In soil, 2,3,5,6-TTCP degrades to 2,3,5-TCP and 2,3,6-TCP (606, 17).
PCP In spite of its extensive use as a biocide (30), pentachlorophenol is quite readily broken down by a wide variety of organisms, and is used as a carbon source for growth by a number of bacteria (49). Pure or reagent grades of PCP are more readily degraded than commercial products. This is likely due to the contamination of commercial products with dioxins, and other compounds more toxic than PCP, inhibiting the organisms responsible for degradation (49). The half-life of PCP in water is a function of light intensity and duration as well as biological activity. Times of 2 to 5 days were found in some experimental ponds. Turbidity and algal blooms prolong the half-life. The biological half-life of PCP in aerobic waters is about 14 to 19 days, and is increased in darkness, in soil-free habitats, at a pH below the pKa of 4.8 and under anaerobic conditions (159). Bacteria can grow with PCP as their sole carbon and energy source and at 25C the conversion rate reached 10 µg PCP per hour per mg of cells (dry weight). In 24 hours, 73% of the added PCP was converted to CO2 (50). A one year aerobic soil metabolism study in the dark on sandy loam soils at 25 degrees centigrade gave a half life of 63 days using first order degradation kinetics. The degradation appeared to follow a process of progressive dechlorination and yielded TTCP and TCP metabolites (731). Under similar conditions a 60 day anaerobic study gave a half life of 13.9 days for the aerobic controls but no degradation occurred in the anaerobic state. Additional degradation products included mucochloric acid and 2-chlorohydroquinone (732). A 30 day aerobic study using flooded blue sandy loam at 25 degrees centigrade gave a half life of 4.93 days using first order kinetics. Degradation products included TTCP, TCP and 3,4-DCP (733). A similar one year anaerobic study gave a half life of 33.8 days using first order degradation kinetics, TTCP and TCP were the degradation products identified (734). The evidence for fungal and bacterial degradation of PCP is widespread. Use of isotopically labeled carbon indicates ring cleavage and CO2 evolution (20). In aerobic sediments, at both 0C and 4C, bacterial degradation of PCP is slower than that of 2-MCP, 4-MCP and 2,4-DCP. In water at 20C, degradation is negligible (19). In rice fields, PCP breakdown via reductive-dechlorination occurs and results in the following metabolites, all containing the persistent meta-chloro substituents which are the hardest to degrade: 2,3,4,5-TTCP, 2,3,5,6-TTCP, 2,3,4,6-TTCP, 2,4,5-TCP, 2,3,5-TCP, 3,4-DCP, 3,5-DCP and 3-MCP (536). Degradation of PCP in sediments begins with dechlorination to form a series of partially dechlorinated products, followed by oxidation to form substituted catechols or hydroquinones and then by ring cleavage and breakdown into CO2 and chloride (175). Many intermediate products are formed in this process and it may be difficult to tell which really are intermediate products and which were contaminants in the original PCP product. Methylation of the hydroxyl forms pentachloroanisole, PCA, which is rapidly accumulated by aquatic life and has a longer half-life than PCP (141). Intermediates include tri- and tetra-chlorophenols, anisols, resorcinols, quinones, hydroquinones, benzoquinones, hydroxybenzoquinones and catechols (Fig. 4.1). Acclimated sludge did not degrade PCP in 4 days, when the PCP was initially present at 100 mg/L. Neither ring degradation nor halide ion production occurred (536). PCP-utilizing bacteria occur, but they are slower growing, and when other nutrients are available these will be preferentially used and PCP degradation rates will decrease (536). Sludges from commercial wood treating operations were readily broken down when composted in permeable soil at PCP levels of 200 mg/L or less. There was 98% breakdown in 200 days, and the soil could then be reused for another batch (191). Activated sludge bacteria break down PCP to CO2, H2O and HCl. In a pilot plant study, wood preserving plant sludge at 23 mg/L was reduced to 0.4 mg/L (191). In sewage treatment plant aeration lagoons, PCP concentrations of 39.5 mg/L and 81 mg/L dropped to 0.5 mg/L and 0.6 mg/L in 3 days and 30 hours, respectively (191). PCP breakdown by sewage sludge bacteria had half-lives of 0.36 days when aerobic, and 192 days when anaerobic (608). The degradation of PCP by Pseudomonas cultures was temperature dependent. No degradation occurred at 0C and the half-lives were 8 to 10 days at 20C and 80 days at 4C (609). In natural sediments where temperatures, oxygen levels, and naive bacteria levels are low, activity rates will also be low and breakdown rates will be very slow (608, 19). The very variable results noted in the literature for half-lives is likely due to the interplay of these factors, some of which were not recorded in all experiments. The adaptability of natural sediment bacterial populations, when repeated PCP doses are applied, is well documented.
Since direct partitioning from water is presumed to be the primary uptake mechanism, ratios in fish and in the water should be the same for different congeners. The rate of uptake is a function of the lipophilicity until the K o/w exceeds about 104 (log K o/w = 2.01); this is probably a function of membrane/water interfaces or boundary layer effects, where no mixing occurs (153). Permeability through a membrane is a function of lipophilicity, and two membranes with an aqueous cytoplasm phase between them form a diffusion barrier. Ionized compounds are polar which restricts their uptake through membranes. As the pH rises the ionization of chlorophenols increases and uptake rates decrease (6). If absorption occurs below the pKa, where the compound is not ionized, uptake is maximal and not pH dependent. Clearance rates for chlorophenols are sufficiently rapid that organisms should be in dynamic equilibrium with water levels. The absorption rate of PCP is proportional to the concentration in water to well beyond the LC50 dose for fish. PCP is readily absorbed from the gastrointestinal tract, skin and gills of fish (225). In people, PCP is readily absorbed from the lungs, gut and skin (405, 199). The uptake rate of PCP in leeches is pH and temperature dependent (86); pH dependency in many organisms is well documented (86, 79, 6, 113, 363). Uptake in eastern oysters, Crassostrea virginica, reaches equilibrium after 4 days at constant ambient levels of PCP (119). Uptake rates of 2,4,5-TCP in Pimephales promelas were 0.2 and 0.34 µg/g/hour from solutions of 4.8 and 49.3 µg/L at 22C, pH 7.5 and alkalinity of 41.5 mg/L . Maximum bioconcentration factors (BCF) of 1900 and 1800 were achieved in 1 or 2 days (605). Fasting fish absorb PCP through the gills and store it in the fat from whence it may be passed along the food chain (121). A 1979 paper by Neely (613) discusses how to model uptake and clearance rates of chemicals by fish and calculate the rate constants. Thus, in the absence of experimental data, a reasonable estimate of the constant, based on fish physiology and chemical properties, can be made.
Bioconcentrating compounds usually have a Log K o/w >4.5 or water solubilities up to 1 mg/L. Usually K o/w is over 1000 for a BCF of 100 (109). Acute toxicity varies six-fold over the range pH 4 to pH 9 and bioconcentration varies similarly (6). At the circumneutral pH of most waters, PCP is over 99% dissociated and the sodium salt is readily soluble. Below the pKa of 4.7 PCP is soluble in most organic solvents but sparingly so in water (199). Assuming that accumulation is correlated with the equilibrium partitioning between lipid and water phases, a regression equation has been developed with a correlation coefficient of 0.97 (378): log BCF = log P - 1.32 BCF = biological concentration factor For leeches at 12C and pH 7, a similar regression equation was developed; the BCF doubles for a 10C rise in temperature (109). log BCF = 1.265 log P - 1.201 One needs to specify pH values for experiments since at higher pH the chlorophenols will be increasingly ionized and thus decreasingly fat soluble. There is reasonable agreement between calculated and measured BCFs in 1-day-whole-body, experiments, but not for longer time periods. Specific organs will have higher BCFs (335). Most experimental BCFs will be derived from waters near neutral, where many chlorophenols are mostly ionized, and thus will be lower than calculations carried out below the pKa would indicate. Comparison of Table 4.2.3.1, which gives the calculated values, and Table 4.2.3.4 which gives measured values for whole fish relative to water, shows this discrepancy clearly. The theoretical BCFs calculated for PCP, based on an undissociated Log Ko/w of 5.2, are too high. At environmental pH levels, PCP is almost completely dissociated and the mean of the measured BCFs is about one quarter of the mean of the calculated BCFs ( the one anomalous high calculated value was not included in the average). Table 4.2.3.7 shows the effects of pH on the toxicity of chlorophenols to guppies, Lebistes reticulata. The toxicity obviously rises as the pH drops; the mg/L needed to show an LC50 response becomes less at lower pH values. There is also, as a rule, a sharp drop in the toxicity between pH 7.3 and pH 7.8; the response is not linear over the entire pH range. Even in the ambient range of pH 6 to pH 7.3 the slope is quite steep and a small pH change has a noticeable effect on the toxicity. One other observation arising from this table is that the ratio of the toxicity at pH 8 over that at pH 6 tends to rise as more chlorines are added to the molecule; that is as the molecule becomes more lipophilic. This means that it is relatively more readily taken up by fish in its less dissociated form, which prevails at lower pH levels. From a practical point of view this means that for a unit change in pH there is a greater change in the toxicity with more highly substituted congeners. PCP accumulates in the fatty tissues of animals; significant levels are found in human adipose tissues (536). Average PCP levels in human female adipose tissue in Florida residents was 0.023 to 0.025 mg/kg; apparently PCP does not bioconcentrate in humans (576). Accumulation in mammalian tissues is low due to very rapid urinary excretion rates, usually as glucuronide conjugates (289). Leeches however have high bioconcentration capacities for chlorophenols and slow elimination rates. This makes them good indicators of long-term contamination and transient events (192, 193, 194). PCP will be found in the liver, kidney and gall bladder prior to elimination. Accumulation or magnification in predators is not appreciably more than that in grazers, due, again, to rapid breakdown and elimination (13). Bioaccumulation depends upon the ratios of absorption, metabolism and elimination (153), and, in fish, the fat content is correlated with uptake and concentration (280). Since they are immersed in the uptake medium, and have very efficient uptake sites in the gills, fish are most susceptible to bioaccumulation of PCP with factors up to 1000 for whole fish in sub-lethal PCP concentrations in the water (204, 279). Concentration factors in the gall bladder, where concentration and conjugation occur preparatory to elimination, may reach 5400 (204). Amphibian tadpole stages and neotenous forms are even more susceptible to chlorophenols since they too are fully immersed, have gills, are smaller, and do not have protective scales on most of their body. Bioconcentration occurs especially in the liver, up to 200-fold for MCPs and DCPs, and up to 10000-fold for 2,4,6-TCP (220). The bioconcentration factor for phenol during a 1-day exposure by Daphnia magna was 634 (47). Bioaccumulation factors for C14 -labeled PCP were: 5 for the alga, Oedegonium cardiacum Of the C14 in the various species, intact, un-degraded, PCP accounted for 11% in the alga, 12% in the snail, 33% in the mosquito, 56% in the daphnia and 51% in the fish (12). Tables 4.2.3.4 and 4.2.3.5 list literature values of bioaccumulation factors relative to water levels for PCP and for other chlorophenols, respectively. Where the reference gives a range of values, only the highest factor is given and only factors over one are listed. Bioaccumulation factors relative to sediments are listed in Table 4.2.3.6. Factors of less than one are listed in this Table, due to the higher chlorophenol concentrations in the sediment as opposed to the concentrations in water. The BCF is a function of species differences and lipid content, but not temperature (682). Figures 4.2.3.1 and 4.2.3.2 show the ranges of bioaccumulation factors reported for various tissues of fish and for various groups of aquatic organisms, respectively. Note that the scales on these figures are logarithmic; there is a great deal of variation in bioaccumulation factors from different species and under different conditions. The ratio of PCP to 2,3,4,6-TTCP in fish is larger than the ratio in the water, suggesting that preferential bioaccumulation of PCP over 2,3,4,6-TTCP is occurring. This is a reflection of the higher Ko/w of PCP. In sculpins the ratio was 1640/440 and in starry flounder 380/100, for a ratio of 3.7 for PCP over 2,3,4,6-TTCP (269, 720). This is in good agreement with the ratio of 4.6 for the Ko/w values of PCP and 2,3,4,6-TTCP as given in Table 2.2, verifying the assumption that uptake rates are a function of lipid content. In sunfish, bass and catfish the bioconcentration factors for PCP were 500x in muscle, 1500x in gills and 8000x in livers, when tests were done using 0.1 mg/L PCP in the water (201). The bioconcentration factor for PCP in Gambusia affinis was 296 at pH 8 (12). Fingerling chinook salmon, Oncorhynchus tshawytscha, were exposed to 2,3,4,6-TTCP and PCP, in "Woodbrite 24" at 0.7C for 62 days. Woodbrite 24 has 250 g/L of chlorophenols, consisting of 156 g/L of TTCPs and 93 g/L of PCP. The solution used for the fish was 0.002 mg/L of total chlorophenols (1.3 µg/L of TTCPs and 0.7 µg/L of PCP), and fish tissues reached 0.224 mg/kg TTCPs and 0.43 mg/kg PCP (203). The ratio of TTCPs to PCP in Woodbrite 24 is 1.68 to 1.00, but the ratio in the tissues of the experimental fish ranged from about 0.65 to 0.52, using NaPCP (203). This is a further illustration of the earlier point that PCP is preferentially bioaccumulated over TTCPs. It is not the relative proportions of chlorophenol compounds in the water which determine their ultimate proportions in tissues, but rather their relative uptake rates from the water. These uptake rates are proportional to the Log Ko/w values rather than to the concentrations in the water. Juvenile rainbow trout, 250 to 450 g, were exposed to a chlorophenol mixture for 6 days at pH 7.3, 12C and water hardness of 80 mg/L. The levels of the various chlorophenols in the bile and in the plasma were measured at the end of the exposure period and are given in Table 4.2.3.8. The ratios of the chlorophenols in the bile, and especially in the plasma, were quite different to the ratios in the water which is a reflection of the relative Log Ko/w s of the compounds. Plasma level ratios approximated the product of the log Ko/w of the chlorophenol and the ratios of the chlorophenol in the water. Small rainbow trout, Oncorhynchus mykiss, 8 to 10 g in size, were exposed to 26 µg/L of PCP for 24 hours. In a subsequent experiment, the previously treated fish were placed in clear running water for a further 24 hours. PCP levels in the liver, blood, fat and muscle were measured at intervals during exposure and subsequent cleansing. The results are given in Tables 4.2.3.9 and 4.2.3.10. These experiments show that muscle levels never got very high, fat levels were slow to clear, blood levels cleared rapidly with a half-life of less than 24 hours, and most PCP was found in the liver. Even the high levels in the liver had a half-life of less than 24 hours. A similar experiment was carried out for days rather than hours using the bluegill sunfish, Lepomis macrochirus. A sub-lethal concentration of 0.1 mg/L of PCP was used with 6-month-old fish of 16 to 42 g, at pH 7.2 to 7.7, 17 to 21C, 135 to 185 µs/cm conductivity and DO of 7.4 to 8.6 mg/L. PCP was measured and adjusted to maintain losses from uptake, evaporation, adsorption, decomposition, photodissociation and bacterial breakdown. Values given are means of 3 to 6 replicates; the fish were not fed. PCP in the tissues is expressed as µg PCP/g wet weight of fish. During the treatment phase, levels rose for 8 days, then, due to elimination, detoxification and metabolic breakdown, fell to an equilibrium value. As in bacterial degradation processes, there is a lag phase while metabolic machinery becomes optimized to deal with the new toxin (121). These results are given in Tables 4.2.3.11 and 4.2.3.12. These experiments demonstrate again the short half-life of PCP in fish tissues; by the fourth day of recovery in clean water, when sampling began, levels were already below one half of the starting concentrations reached after 16 days of exposure. Tables 4.2.3.2 and 4.2.3.3 give some bioaccumulation factors for PCP and TTCPs, relative to the sediments, in various tissues of benthic species of marine and freshwater habitats in the lower Fraser Valley and around Victoria. These are drawn from the same data set as the data in Table 3.5.2, which gives the actual sediment levels. These values demonstrate the wide range of values which may be found when one looks at different organisms and different habitats. The bioconcentration and elimination of C14 -labeled 2-MCP was studied in 100 bluegill sunfish, Lepomis macrochirus. Work was done in continuous-flow-aquarium-culture with a turnover rate of six to seven aquarium volumes per day. Water parameters were DO 5.9 to 8.6 mg/L, pH 6.3 to 7.9 and hardness of 35 mg/L. The fish were exposed for 28 days at 16C to 9.18 µg/L of 2-MCP and then transferred to clean water to measure elimination over 7 days. The bioconcentration factor at equilibrium was 214 and the half-life in the tissues was less than 1 day (379, 423). Rainbow trout, Oncorhynchus mykiss, were exposed to pulp mill bleaching effluents in Baltic Sea water, salinity 7 ppt, diluted to 2.5 ppt, at 9C and pH 7, in a flow-through system. The concentrations of chlorophenols and other contaminants in the effluent were not measured. A 2-week exposure to effluent from the chlorination step led to 16 µg of chlorophenol/g fat in the livers; 5 weeks led to 17 µg/g fat. Using effluents from the extraction step, a 2 week exposure gave 25 µg/g fat, and 5 weeks, 45 µg/g fat. These were all liver samples with 2.3 to 2.9% fat content. When effluent from a chlorination step using 5% chlorine dioxide and 95% chlorine was used, instead of the preceding 100% chlorine, the liver fat levels were 2.9 µg/g after 6 weeks in chlorination step effluent and 1.5 µg/g in extraction step effluent. After an 11 week exposure these levels were 2.8 µg/g and 2.1 µg/g respectively, and 1 and 0.8 µg/g fat in muscle tissue. In all cases levels in the liver dropped to zero 3 weeks after exposure was discontinued, demonstrating the short half-life of chlorophenols in fish livers. (187). In goldfish subjected to 200 µg/L PCP, the gall bladder levels reached 44 µg/g in 5 hours and 539 µg/g after 24 hours, at which time the whole body load was 2475 µg/g. After a further 24 hours in clean running water, the whole body load dropped to 1720 µg/g as excretion took place, but the gall bladder levels rose to 1077 µg/g since excretion took place via the gall bladder (136). In Bluegill sunfish continuously exposed for 28 days to 2.5 µg/L of PCP the mean steady state concentration of PCP in the edible tissues was 500 µg/kg for a BCF or bioconcentration factor of 190X, in the non-edible tissue was 2100 µg/kg for a BCF of 790X and in the whole body was 1300 µg/kg for a BCF of 490X (735). In the Weser Estuary and German Bight, the actinian, Sagartia troglodytes lives tightly attached to the polychaete worm Lanice conchilega thus both were exposed to the same PCP levels which averaged 0.04 mg/L. The BCF for S. troglodytes was about 70 to 180 while for L. conchilega it was 2600 to 8500 (641, 642, 643). This is a good demonstration of the inter-species variability in bioaccumulation of chlorophenols. Oysters, Crassostrea virginica, concentrated NaPCP 41 and 78 fold when exposed to 25 and 2.5 µg/L in the water. Purging took only 4 days in clean water (119). The eel, Anguilla anguilla, exposed to 0.1 mg/L of PCP in sea water for 8 days, accumulated 33.4 mg/kg in liver, 9.4 mg/kg in muscle and 4.4 mg/L in blood. After 8 days, levels dropped to 11.9 mg/kg, 3.6 mg/kg and 2.1 mg/L respectively. In fresh water, exposure to 0.1 mg/L of PCP resulted in accumulations of 8.8 mg/kg in liver, 0.81 mg/kg in muscle and 1.7 mg/L in blood after 4 days, and in 55 days levels dropped to 1.3 mg/kg in liver and 0.08 mg/kg in muscle. After 38 days blood levels dropped to 0.31 mg/L (667). Table 4.2.3.13 gives the relative PCP accumulations in a fish, a shrimp and an oyster, relative to water. These are all marine organisms and the results are from 96-h exposures in flowing sea-water.
In the gall bladder, chlorophenols are conjugated, usually with glucuronides, for excretion in the urine. Thus, high transient levels may be found in the gall bladder (204, 289). The rapid excretion rate of chlorophenols keeps the bioaccumulation factor relatively low, in spite of the high calculated values in Table 4.2.3.1 based on Ko/w figures, and permits periodic low doses to be tolerated with little permanent toxic effect (263). It takes naive fish about 8 days to develop a fully functioning elimination mechanism for chlorophenols (121). The liver is the main metabolic organ for breaking down or detoxifying chlorophenols, before transport to the gall bladder for conjugation and ultimate elimination in the urine via the kidney (208, 289). Uptake and elimination of chlorophenols by fish is rapid with a half-life of 2 days for MCPs and DCPs and 10 days for TCPs, TTCPs and PCP (220). Oysters, Crassostrea virginica, purge themselves of PCP in 4 days (119). Some depuration half-lives of chlorophenols, mostly PCP, are given in Table 4.2.4.1. The values are recorded in hours and the longest half-life reported is a whole body value for trout of 168 hours or 7 days. 2-MCP The half-life of 2-MCP in the whole body of the bluegill sunfish, Lepomis macrochirus, is less than 1 day (379). Dogs excreted 87% of the 2-MCP as glucuronic and sulphate conjugates (315, 372) and rabbits also conjugate 2-MCP (315). 2,4,5-TCP Depuration in Pimephales promelas, which had reached BCFs of 1800 to 1900 in 2 days, was rapid with a biological half-life of 12 hours following transfer to clean water. These fathead minnows lost 84% to 92% of their 2,4,5-TCP load during the first day after exposure stopped (605). 2,4,6-TCP Rapid clearance from the body, in the urine, occurs for 2,4,6-TCP (225). In male rats, radioactive 2,4,6-TCP was given by stomach tube for 3 days at 1 mg/kg diet. Of the amount applied, 80% was eliminated in the urine and 20 % in the feces within 5 days. A subsequent autopsy showed no detectable levels in liver, lungs or fat (460). Rainbow trout livers were cleared of 2,4,6-TCP 21 days after dosing was discontinued (187). 2,3,4,6-TTCP This chlorophenol was excreted unchanged in the urine, for the most part, when given by intraperitoneal injection to rats at 10 mg/kg. Within 72 hours of the dose being given, 66% could be recovered unchanged from the urine. Over 95% is eliminated within 48 hours. Trichloro-p-hydro-quinone is a minor metabolite (450). 2,3,5,6-TTCP About 35% of an intraperitoneal dose of 2,3,5,6-TTCP, given to a rat, was excreted as tetrachloro-p-hydroquinone, within 24 hours (45).
PCP PCP is reportedly excreted in the urine unchanged for the most part (278, 406). In rats and mice, some PCP is dechlorinated to TTCP- and TCP-hydro-quinones (225). In goldfish, PCP is excreted as a conjugate (135, 136, 204) and the half-life for clearance is 10 hours (204). The half-life of PCP in mammals, including people, is measured in hours (276, 277); in trout it is about 7 days for the whole body (131, 577), 6.7 hours for blood, 4.8 hours for liver, 23.0 hours for fat, 6.9 hours for muscle, 10.3 hours for gills, and 6.9 hours for the heart (134, 141). Work with midges, Chironomus riparius, showed that 89% of the PCP was still unchanged after the experiment, and that pH had no effect on the amounts or nature of the breakdown products (6). The clam, Tapes philippinarum, forms sulphate conjugates of PCP as a means of detoxification prior to excretion, instead of the more common formation of glucuronides (143). In mammals, there is little fecal excretion, long-term tissue accumulation, or storage, of PCP. Detoxification is by oxidative conversion to quinone or glucuronic acid conjugation and excretion in the urine (199). In people, mice and rats, the primary mode of PCP elimination is the urine. The amount and rate of excretion increases with increasing levels of PCP in the body. Initial elimination is rapid, but complete removal of all the residue may take a month. In the mouse, 72 to 83% is excreted in 4 days and in the rat 68.3% in 10 days. Trace amounts are respired (536). Urinary excretion in mice and rats is mostly free PCP or tetrachlorohydroquinone but glucuronide conjugates also occur. The plasma half-lives of 10 mg/kg were 15 hours in rats and 78 hours in Macaca mulatta, monkeys (464). In Rhesus monkeys essentially all excretion is urinary. The half-life in plasma was 84 hours for females and 72 hours for males; excretion half lives were 92 hours for female and 41 hours for males (570). In male and female rats given 10 or 100 mg/kg PCP, the rapid phase of elimination had a half-life of 17 hours in females and 13 hours in males; this phase eliminated over 90% of the dose of PCP. Excretion was rapid in the urine as PCP, its glucuronide conjugate, or as tetrachlorohydroquinone (277). Female mice were given 15 to 37 mg/kg PCP by either subcutaneous or intraperitoneal injection. The half-life was about 24 hours, and 72 to 83% was excreted in the urine in 4 days, mostly as PCP (407). PCP is rapidly excreted in fish, after formation of the glucuronide and sulphate conjugates, with tissue half-lives of less than 24 hours (205). Rainbow trout, Oncorhynchus mykiss, fed PCP in their diet reached equilibrium between constant uptake and elimination after a period of time which was a function of the dose. Fish receiving 10 µg/kg in their diets reach equilibrium of 2 µg/kg in 28 days. Higher dose levels reached a higher equilibrium level and took longer to achieve it. The half-life of PCP in the whole body was about 7 days (131). In Bluegill sunfish continuously exposed for 28 days to 2.5 µg/L of PCP the half life for depuration or elimination from the whole body, of 1300µg/kg, was 1 day and 98% was eliminated in 14 days. Metabolism of PCP by fish tissues was minimal (735).
Once released to the environment, physical, chemical and biological processes break down chlorophenols, ultimately to CO2, H2O and chloride, with half-lives ranging from hours to months. Table 4.1 gives a summary of the fate of chlorophenols in water. Photodegradation is only important in shallow water under high irradiation levels; hydrolysis and oxidation are not important in nature; evaporation and volatilization are only important in nature in shallow water subject to vigorous mixing; adsorption is a major process and most chlorophenols introduced into the environment will eventually be found in the sediment, usually on organic sediments. Microorganisms can adapt their metabolic processes to use virtually any source of carbon for growth. Once the adaptive phase is over and a large population has built up, breakdown is rapid and subsequent additions to the environment are broken down quickly, if the concentrations are not excessive. If the organisms have never been exposed to chlorinated organics there will be an initial lag period while adaptation occurs. The evidence for fungal and bacterial degradation in nature is widespread and Table 4.2.1 gives degradation rates under different pH and temperature conditions. Chlorophenols with chlorines in the 2,4 or 6 positions are more readily degraded than those with chlorines in the 3 or 5 positions. Bioaccumulation of chlorophenols is species specific; under identical conditions different species may accumulate very different amounts of chlorophenols. Bioaccumulation in the same organism will vary depending upon the environmental conditions. Bioaccumulation factors are different for accumulation from water than for accumulation from sediments and differ from marine to fresh waters. The pH affects the accumulation factor by affecting the dissociation of the chlorophenol and thus the fat solubility and uptake rate. The bioaccumulation factors for different tissues in the same organism vary widely, and whole body values are low compared to values for detoxification and excretion organs. Temperature affects bioaccumulation, much more in poikilotherms than in homeotherms, since all enzymatic reactions are temperature dependent with rates generally doubling with a 10C rise in temperature. Compounds which bioconcentrate usually have Log Ko/w values over 4.5 or water solubilities under 1 mg/L. Usually for a BCF of 100 the Ko/w will be over 1000. In mammalian tissues, bioaccumulation is generally low due to high urinary excretion rates, but in poikilotherms accumulation may be much higher. Biomagnification does not generally occur, again due to rapid excretion and detoxification rates. Preferential bioaccumulation of compounds with higher Ko/w values occurs even when other compounds are present at much higher concentrations. The half-lives of bioaccumulated chlorophenols in vertebrates are quite short, hours or days, due to very efficient excretory mechanisms. In mammals, chlorophenols are not accumulated in fat to a very high level due to very rapid excretion as glucuronide conjugates, thus keeping bioconcentration factors low. Since fish are immersed in their uptake medium and have gills which are very efficient uptake organs, they bioaccumulate chlorophenols up to 1000 times for whole body loads, in spite of very efficient conjugation and elimination. Tables 4.2.3.4 and 4.2.3.5 give bioaccumulation factors relative to the water and Table 4.2.3.6 factors relative to the sediments for fish. The half-life of chlorophenols in fish is less than 1 day as shown in Tables 4.2.3.10 and 4.2.3.12; thus the existence of high levels in fish tissues is indicative of chronic exposure. There is an initial lag period while the metabolic processes of the fish adapt to conjugating and excreting chlorophenols, but once adapted, the depuration half-lives are short as indicated in Table 4.2.4.1., generally less than 24 hours. In nature many aquatic organisms do not leave their contaminated sites and thus the effectiveness of the rapid detoxification is limited (711, 712).
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