Literatura académica sobre el tema "Pentachlorophenol"

Background: This study aimed to investigate the biosorption of pentachlorophenol on Aspergillus niger biomass as a method for removal of pentachlorophenol from aqueous solutions. Methods: Modified A. niger biomass with NaOH was used to absorb the pentachlorophenol. The impacts of various experimental parameters like primary pentachlorophenol concentration, pH of the solution, contact time, and biomass dosage on the biosorption of pentachlorophenol were investigated. Results: The correlation of contact time, pH and initial concentration with the biosorption of pentachlorophenol by A. niger biomass was statistically significant (P<0.001). Pentachlorophenol removal increased with decreasing pH of the solution and the maximum efficiency was obtained at pH=3. The equilibrium adsorption capacity was increased from 4.23 to 11.65 mg/g by increasing initial pentachlorophenol concentration from 10 to 40 mg/L, while pentachlorophenol removal efficiency decreased from 87 to 55%. Both Langmuir and Freundlich isotherms efficiently described adsorption equilibrium of pentachlorophenol on A. niger biomass. Correlation coefficients for the second order kinetic model were almost equal to one. Conclusion: A. niger biomass can be used to reduce the toxicity of aqueous solutions containing pentachlorophenol in acidic pH conditions.

The 2nd year of a 2-year study of the fate of pentachlorophenol in outdoor artificial streams focused on details of microbial degradation by a combination of in situ and laboratory measurements. Replicate streams were dosed continuously at pentachlorophenol concentrations of 0, 48, and 144 μg/L, respectively, for an 88-d period during the summer of 1983. Pentachlorophenol was degraded both aerobically and anaerobically. Aerobic degradation was more rapid than anaerobic degradation. Mineralization of pentachlorophenol was concommitant with pentachlorophenol disappearance under aerobic conditions, but lagged behind loss of the parent molecule under anaerobic conditions. Biodegradation in the streams, or in specific stream compartments such as the sediment or water column, was characterized by an adaptation period (3–5 weeks for the stream as a whole, and reproducible from the previous year), which was inversely dependent on the concentration of pentachlorophenol and microbial biomass. The adaptation in the streams could be attributed to the time necessary for selective enrichment of an initially low population of pentachlorophenol degraders on surface compartments. The extent of biodegradation in the streams (percent loss of initial concentration of pentachlorophenol) increased with increasing pentachlorophenol input, which was explicable by an increase in the pentachlorophenol degrader population with increasing pentachlorophenol concentration. The sediment zone most significant to overall pentachlorophenol biodegradation was the top 0.5- to 1-cm layer as shown by pentachlorophenol migration rates and depth profiles of degrader density within the sediment. Pentachlorophenol profiles in sediment cores taken during and after the adaptation period for degradation showed that diffusion of pentachlorophenol into the sediment was rate limiting to degradation in this compartment. Degradation rates were independent of temperature within the temperature range of the streams during the dosing season (19–30 °C), but became increasingly slower below 19 °C. The impact of sudden increases in toxicant level (to 10 or 100 mg/L) on degradation was significant (negative), and was assessed by laboratory experiments with sediments. Total heterotrophic activity of sedimentary communities over a major part of the season was unaffected by pentachlorophenol at all stream concentrations tested.

This study is focused on the removal of pentachlorophenol from its aqueous phase by electrochemically induced degradation with Pt electrodes. The objective of this study was to contrast the electrochemical removal of pentachlorophenol at the oxidative and the reductive potentials, and further to understand how to apply the electrochemical treatment on PCP degradation. Lab experiments were conducted in a Pt electrolyzer, and the voltage source was supplied and precisely controlled by an electrochemical analyzer. In these experiments, the variables including electrolyte species, pH, voltage supply, and reaction time were examined to compare the efficiency of pentachlorophenol removal. Experimental results showed that pentachlorophenol was completely degraded after being electrolyzed for 1 h at−1.5 V in a 0.5 M KCl solution, while the removal of pentachlorophenol is negligible under the similar condition when 0.5 M NaNO3 or Na2CO3 was used as the electrolyte. The electrolyte concentration below 0.5 M is unfavourable for the electrochemical removal of pentachlorophenol. The removal efficiency of pentachlorophenol is slightly affected by pH, and the strong basic environment might impede the degradation of pentachlorophenol. Comparing with those under positive potentials, the experiments conducted under negative potentials have shown a better removal of pentachlorophenol with a higher current efficiency. It implies that pentachlorophenol degradation followed the reductive pathway. Based on the analysis of GC/MS, the intermediates of pentachlorophenol degradation were identified as 1,2-dichlorocyclohexane and 2-chlorocyclohexanol.

The effect of a chromated-copper-arsenate wood preservative on the degradation of pentachlorophenol by Flavobacterium sp. strain ATCC 53874 was examined in liquid culture. Both a commercially available and a laboratory-prepared formulation were tested. Each increased the lag time required for measurable pentachlorophenol degradation and the time required for complete degradation to nondetectable levels. This response was noted at all pentachlorophenol concentrations examined (10, 25, 50, 75, and 100 μg∙mL−1). The commercial formulation of chromated-copper-arsenate had the more significant impact on pentachlorophenol degradation. Inhibitory effects were evident at chromated-copper-arsenate component metal concentrations 0.1–0.5 mg∙L−1. These levels are thousands of times below those used commercially.Key words: pentachlorophenol, biodegradation, chromated-copper-arsenate, toxicity.

Pentachlorophenol is used as an industrial wood preservative for utility poles, crossarms, fence posts, and other purposes (79%);for NaPCP (12%); and miscellaneous, including mill uses, consumer wood preserving formulations and herbicide intermediate (9%) (CMR, 1980). As a wood preservative, pentachlorophenol acts as both a fungicide and insecticide (Freiter, 1978). The miscellaneous mill uses primarily involve the application of pentachlorophenol as a slime reducer in paper and pulp milling and may constitute ∼6% of the total annual consumption of pentachlorophenol (Crosby et al., 1981). Sodium pentachlorophenate (NaPCP) is also used as an antifungal and antibacterial agent (Freiter, 1978). Pentachlorophenol also is used as a general herbicide (Martin and Worthing, 1977). Photolysis and microbial degradation are the important chemical removal mechanisms for pentachlorophenol in water. In surface waters, pentachlorophenol photolyzes rapidly (ECETOC, 1984; Wong and Crosby. 1981; Zepp et al., 1984); however, the photolytic rate decreases as the depth in water increases (Pignatello et al., 1983). Pentachlorophenol is readily biodegradable in the presence of accli-mated microorganisms; however, biodegradation in natural waters requires the presence of microbes that can become acclimated. A natural river water that had been receiving domestic and industrial effluents significantly biodegraded pentachlorophenol after a 15-day lag period, while an unpolluted natural river water was unable to biodegrade the compound (Banerjee et al., 1984). Even though pentachlorophenol is in ionized form in natural waters, sorption to organic particulate matter and sediments can occur (Schellenberg et al., 1984), with desorption contributing as a continuing source of pollution in a contaminated environment (Pierce and Victor, 1978). Experimentally determined BCFs have shown that pentachlorophenol can significantly accumulate in aquatic organisms (Gluth et al., 1985; Butte et al., 1985; Statham et al., 1976; Veith et al., 1979a,b; Ernst and Weber, 1978), which is consistent with its widespread detection in fish and other organisms. Direct photolysis may be an important environmental sink for pen tachlorophenol present in the atmosphere. The detection of pen tachlorophenol in snow and rain water (Paasivirta et al., 1985; Bevenue et al., 1972) suggests that removal from air by dissolution is possible. Soil degradation studies indicate that pentachlorophenol is biodegrad able; microbial decomposition is an important and potentially domin ant removal mechanism in soil (Baker et al., 1980; Baker and Mayfield, 1980; Edgehill and Finn, 1983; Kirsch and Etzel, 1973; Ahlborg and Thunberg, 1980). The degree to which pentachlorophenol leaches in soil is dependent on the type of soil. In soils of neutral pH, leaching may be significant, but in acidic soils, adsorption to soil generally increases (Callahan et al. , 1979; Sanborn et al. , 1977). The ionized form of pentachlorophenol may be susceptible to adsorption in some soils (Schellenberg et al., 1984). In laboratory soils, pen tachlorophenol decomposes faster in soils of high organic content as compared with low organic content, and faster when moisture content is high and the temperature is conducive to microbial activity. Half- lives are usually ∼2-4 weeks (Crosby et al., 1981). Monitoring studies have confirmed the widespread occurrence of pentachlorophenol in surface waters, groundwater, drinking water and industrial effluents (see Table 2). The U.S. EPA's National Urban Runoff Program and National Organic Monitoring Survey reported frequent detections in storm water runoff and public water supplies (Cole et al., 1984; Mello, 1978). Primary sources by which pen tachlorophenol may be emitted to environmental waters may be through its use in wood preservation and the associated effluents and its pesticidal applications. Pentachlorophenol can be emitted to the atmosphere by evaporation from treated wood or water surfaces, by releases from cooling towers using pentachlorophenol biocides or by incineration of treated wood (Skow et al., 1980; Crosby et al., 1981). Pentachlorophenol has been detected in ambient atmospheres (Caut reels et al., 1977), in snow and rain water (Paasivirta et al,. 1985; Bevenue et al., 1972) and in emissions from hazardous waste incinera tion (Oberg et al., 1985). The U.S. Food and Drug Administration's Total Diet Study (conducted between 1964 and 1977) found pen tachlorophenol residues in 91/4428 ready-to-eat food composites (See Tables 4 and 5). The average American dietary intake of pen tachlorophenol during 1965-1969 was estimated to range from <0.001-0.006 mg/day (Duggan and Corneliussen, 1972). The most likely source of pentachlorophenol contamination in many food prod ucts may be the exposure of the food to pentachlorophenol-treated wood materials such as storage containers (Dougherty, 1978). Acute toxicity data indicated that salmonids are more sensitive to the toxic effects of pentachlorophenol than other fish species, with LC50 values of 34-128 μ g/l for salmonids and 60-600 μ g/l for other species. More recent data showed that carp larvae, bluegills, channel catfish and knifefish also had LC50 values < 100 μ gl (see Table 10). The most sensitive marine fishes were pinfish larvae, the goby, Gobius minutus, and eggs and larvae of the flounder, Pleuronectes platessa, all with LC50 values <100 μ g/l (Adema and Vink, 1981). The most sensitive freshwater invertebrate species were the chironomid, Chironomus gr. thummi (Slooff, 1983) and the snail, Lymnaea luteola (Gupta et al., 1984). The most sensitive marine invertebrates were the Eastern oyster (Borthwick and Schimmel, 1978), larvae of the crusta ceans, Crangon crangon and Palaemon elegans (VanDijk et al. , 1977), and the copepod, Pseudodiaptomus coronatus (Hauch et al., 1980), all with LC50 values <200 μ g/l. In chronic toxicity tests, the lowest concentration reported to cause adverse effects was 1.8 μ g/l (NaPCP), which inhibited growth of sockeye salmon (Webb and Brett, 1973). The marine species tested displayed similar thresholds for chronic toxicity. Both acute and chronic toxicity increased at lower pH, probably because a lower pH favors the un-ionized form of pentachlorophenol, which is taken up more readily and is therefore more toxic than ionized pentachlorophenol (Kobayashi and Kishino, 1980; Spehar et al., 1985). Data concerning the effects of pentachlorophenol on aquatic plants were highly variable. Therefore, it was difficult to draw conclusions from these data. Pentachlorophenol did not appear to bioaccumulate in aquatic or ganisms to very high concentrations. BCFs for pentachlorophenol were <1000 for most species tested. The highest BCF was 3830 for the polychaete, Lanice conchilega (Ernst, 1979). Some species appear to have an inducible pentachlorophenol-detoxification mechanism, as evidenced in several experiments in which pentachlorophenol tissue levels peaked in 4-8 days and declined thereafter despite continued exposure (Pruitt et al., 1977; Trujillo et al., 1982). A study by Niimi and Cho (1983) indicated that uptake of waterborne pentachlorophenol from gills was much greater than uptake from food, indicating that bioconcentration of pentachlorophenol through the food chain is unlikely. Biomonitoring data of Lake Ontario fishes showed that similar pentachlorophenol levels were found in predators andforage species. Studies with experimental ecosystems have indicated that ecological effects may occur at pentachlorophenol levels as low as those causing chronic toxicity in sensitive species in single-species tests. The lowest concentration that caused adverse effects in these studies was 15.8 μ g/l, which caused a reduction in numbers of individuals and species in a marine benthic community (Tagatz et al., 1978). Pentachlorophenol is readily absorbed from the gastrointestinal tract of rats, mice, monkeys and humans (Braun et al. , 1977, 1978; Ahlborg et al., 1974; Braun and Sauerhoff, 1976). Peak plasma concentrations are reached within 12-24 hours after oral administration to monkeys (Braun and Sauerhoff, 1976), but 4-6 hours after oral administration to rats (Braun et al., 1977). After oral administration, the highest concentration of radioactivity was found in the liver and gastrointesti nal tract of monkeys (Braun et al., 1977). In rats and mice, tet rachlorohydroquinone was identified in the urine (Jakobson and Yllner, 1971; Braun et al., 1977; Ahlborg et al., 1974) as well as unmetabolized pentachlorophenol and glucuronide-conjugated pen tachlorophenol. Although Ahlborg et al. (1974) reported that oxidative dechlorination of pentachlorophenol occurs in humans, as evidenced by the presence of tetrachlorohydroquinone in the urine of workers occupationally exposed (probably by inhalation), analysis of human urine after ingestion of pentachlorophenol revealed the presence of conjugated pentachlorophenol and unmetabolized pentachlorophenol (Braun et al., 1978). The primary route of excretion after oral administrtation of all species studied is in the urine (Braun et al. , 1977, 1978; Ahlborg et al., 1974; Larsen et al., 1972; Braun and Sauerhoff, 1976). Although urinary excretion followed second-order kinetics in rats (Larsen et al., 1972; Braun et al., 1977) except in females receiving a single high dose (100 mg/kg) of pentachlorophenol, urinary excretion of pentachlorophenol in humans and monkeys followed first-order kinetics (Braun and Sauerhoff, 1976; Braun et al., 1978). Enterohepatic circulation played an importation role in the pharmacokinetics of pen tachlorophenol. The half-life of pentachlorophenol in the plasma is longer in female rats and monkeys than it is in male rats and monkeys (Braun et al. , 1978; Braun and Sauerhoff, 1976). Because many preparations of pentachlorophenol are contaminated with small but measurable amounts of highly toxic substances, such as dibenzodioxins, special attention must be paid to the composition of the pentachlorophenol solution tested. In studies where technical and purified pentachlorophenol have been evaluated (Schwetz et al., 1974; Goldstein et al., 1977; Kimbrough and Linder, 1978; Knudsen et al., 1974; Johnson et al., 1973; Kerkvliet et al., 1982), only the results of the experiments using purified pentachlorophenol were reported in detail. Oral exposure to pentachlorophenol was not carcinogenic in mice (BRL, 1968; Innes et al., 1969) or rats (Schwetz et al., 1977), regardless of the composition of the pentachlorophenol solution tested. Although there are a few studies that suggest pentachlorophenol may be mutagenic in B. subtilis (Waters et al., 1982; Shirasu, 1976), in yeast, Saccharomyces cerevisiae (Fahrig et al., 1977) and in mice, as evidenced by the coat-color spot test (Fahrig et al., 1977), no evidence of mutagenicity was reported in S. typhimurium (Anderson et al. , 1972; Simmon et al., 1977; Lemma and Ames, 1975; Moriya et al. , 1983; Waters et al., 1982; Buselmaier et al., 1973) or in E. coli (Simmon et al., 1977; Fahrig, 1974; Moriya et al., 1983; Waters et al., 1982) with or without metabolic activation. Three teratogenicitylreproductive toxicity studies (Schwetz et al., 1974, 1977; Courtney et al., 1976) indicate that pentachlorophenol is fetotoxic in rats at oral dose levels ≥5 mg/kg/day. At the highest dose tested (500 ppm) in a fourth teratogenicity/reproductive toxicity study (Exon and Koller, 1982), there was a statistically nonsignificant decrease in litter size. The lowest dose tested (5 mg/kg/day) by Schwetz et al. (1977) was the lowest dose at which any evidence offetotoxicity, as indicated by delayed ossification, was observed. No adverse fetal or reproductive effects were reported at ≤3 mg/kg/day (Schwetz et al., 1977; Exon and Koller, 1982). In subchronic and chronic toxicity studies, adverse effects occurred primarily in the liver (Kerkvliet et al., 1982; Johnson et al., 1973; Knudsen et al. , 1974; Goldstein et al. , 1977; Kimbrough and Linder, 1978; Schwetz et al., 1977), the kidney (Johnson et al., 1973; Kimbrough and Linder, 1978; Schwetz et al., 1977) and the immune system (Kerkvliet et al., 1982). Knudsen et al. (1974) reported increased liver weights in female rats and centrilobu lar vacuolization in male rats exposed to diets containing ≧50 ppm commercial pentachlorophenol, which contained 282 ppm dioxins. In the remaining studies, increased liver weight (Johnson et al., 1973) and increased pigmentation of hepatocytes (Schwetz et al., 1977) were observed at oral doses of≥10 mg/kg/day (∼90%), and SGPT levels significantly increased in rats ingesting 30 mg/kg/day pentachloro phenol (∼90%) for 2 years (Schwetz et al., 1977). Increased kidney weight unaccompanied by renal histopathology was reported in rats exposed to dietary concentration ≧20 ppm of pentachlorophenol (>99%) for 8 months (Kimbrough and Linder, 1978) and in rats ingesting 30 mg/kg/day (∼90%) for 90 days (Johnson et al., 1973). Increased pigmentation of the renal tubular epithelial cells was re ported in rats ingesting 10 or 30 mg/kg/day pentachlorophenol for 2 years (Schwetz et al., 1977). Although decreased immunocompetence was reported in mice exposed to dietary levels of 50 or 500 ppm of pentachlorophenol (>99%) for 34 weeks (Kerkvliet et al., 1982), the decrease was statistically significant only at the higher dose. An ADI of 0.03 mg/kg/day or 2.1 mg/day for a 70 kg human was derivedfrom the NOAEL of 3 mg/kg/day in rats in the chronic dietary study by Schwetz et al. (1977). An uncertainty factor of 100 was used. An RQ of 100 was derived based on the fetotoxic effects of pen tachlorophenol in rats in the study by Schwetz et al. (1974). Based on guidelines for carcinogen risk assessment (U.S. EPA, 1984b) and inadequate evidence for animal carcinogenicity or absence of human cancer data, pentachlorophenol is classified as Group D, meaning that it is not classified as a human carcinogen.

Abstract A new, simple, and sensitive spectrophotometric method is described for determination of pentachlorophenol, a widely used insecticide and herbicide, in various environmental samples. The method is based on the reaction of pentachlorophenol with concentrated nitric acid to form chloranil, which liberates iodine from potassium iodide. The liberated iodine then selectively oxidizes leucocrystal violet to form crystal violet, which has an absorption maximum at 592 nm. Beer's law is obeyed over the concentration range of 0.1-1.6 μg pentachlorophenol/ 25 ml_ (0.004-0.064 ppm). The method was applied satisfactorily to determination of pentachlorophenol in air, water, plant material, textile effluent, and biological samples.

Pentachlorophenol (PCP) is a chlorophenol with a pronounced biocidal activity that has led to its use in a number of applications. It was introduced in the 1930s as a preservative for timber and lumber and since then has found wide use as a biocide in agricultural and industrial applications. Many different physical, chemical and biological methods have been tried for the removal of PCP from wastewater. However, using microalgae for the removal of PCP and other organochlorine compounds from water may prove to be a cheaper alternative and give complete degradation of the compounds. The aim of this project was to study the efficiency of microalgae to degrade PCP. An algal strain named VT -1 and a bacterial strain named AT -14 were isolated from PCP containing conditions in the laboratory. The growth of VT -1 in the presence of PCP was compared with Chlorella emersonii and Chlorella vulgaris in two different autotrophic media. It was observed that VT-1 had the highest IC50 value of 25-26mg }-l PCP and EC50 value of 11.3mg }-1 PCP in S&K medium. With glucose as an additional carbon source the IC50 value for VT-1 in S&K medium was 29-30mg t 1 PCP. Bacterium AT-14 could grow in the presence of PCP, only with glucose as a carbon source. Mineralization of PCP by VT -1 and the two Chlorella strains was compared by using 14C_PCp. With all the three algae exposed to 14C_PCp, only VT-1 showed release of 14C02, which was evidence of mineralization of PCP by VT-1 which occurred only in the presence of light. Bacterium AT-14 did not produce 14C02. However, the consortium of VT-1 and AT-14 showed enhanced 14C02 evolution in the presence of glucose. The release of chloride ions from PCP can also indicate PCP dehalogenation and degradation. The evolution of 14C02 lagged behind chloride release (90 %) indicating that dechlorination of PCP could be the first step in its biodegradation. Breakdown of PCP was also followed by its extraction from the cells and medium. Normally dichloromethane (DCM) was used to extract PCP. The changes in the label extracted in DCM and iso-butanol were studied under different light condjtions, which showed that the 14C counts in DCM reduced and those in iso-butanol extract increased with time. The 14C counts in the iso-butanol extract could be a metabolite of PCP which is more hydrophilic. VT-1 appeared not to degrade PCP completely, since only 15% of 14C was recovered as 14C02. It appears that intermediates are formed which are distributed in the growth medium and in the biomass. It can thus be concluded that VT -1 is tolerant of PCP, appears to dechlorinate PCP and then releases some part of it as CO2.

Microbial populations from subsurface soil collected from a hydrocarbon contaminated site and a pristine site with no history of contamination had the ability to degrade pentachlorophenol (PCP) in anaerobic enrichment cultures. Increasing concentrations of PCP in nitrate, sulfate and yeast extract-mineral salts media were used to acclimate the cultures. Nitrate enrichments, previously incubated in an anaerobic phenol-mineral salts medium, showed 23% degradation in medium containing 40 μg ml⁻¹ PCP during a 32 d incubation period. Cultures not adapted to phenol degradation did not degrade PCP at concentrations over 20 μg ml⁻¹. Enrichment cultures grown in the anaerobic yeast extract-mineral salts medium did not degrade PCP at concentrations over 20 μg ml⁻¹ and phenol adaptation did not enhance PCP degradation. The sulfate reducing enrichment containing 1 μg ml⁻¹ PCP showed 71.3% degradation after 32 d incubation. No degradation occurred at or above 5 μg ml⁻¹ PCP. PCP intermediates, 2,4,6-trichlorophenol (TCP) and 3,4,5 TCP were found in the spent culture of the nitrate reducing enrichment. In the spent culture of the sulfate reducing enrichment, 3,4,5 TCP and 2,3,4,5-tetrachlorophenol were found. Attempts to obtain a pure culture of an anaerobic PCP degrading bacterium were unsuccessful.
Master of Science

Dissertation presented to obtain the Ph.D degree in Biochemistry
Pentachlorophenol (PCP) is a synthetic compound introduced as a wood preservative and used for decades in agricultural and industrial applications. Today, the production and use of PCP is restricted due to its toxicity and environmental impact. However, extensive use in the past and its environmental persistence has resulted in substantial contamination of PCP worldwide. The detection of PCP in the human population and in remote environments, such as the Arctic, is still being reported.(...)

Pentachlorophenol (PCP) is one of the prominent environmental pollutant that has penetrated into food chain and is present in humans. Health concerns have been raised since daily intake of PCP by the US population is estimated to be 16-19 µg. PCP facilitates dissipation of electrochemical potential gradients of hydrogen ions across energy transducing membranes, which are the energy sources for the conversion of adenosine diphosphate into adenosine triphosphate. Closely linked to these dissipative effects is the development of electrical conductivity in lipid membranes, induced by the presence of PCP. Three modes of PCP - induced membrane electrical conductivity were theoretically analyzed and experimentally verifiable formulations of each models were developed. Experimental studies using the charge - pulse method involved characterization of the time dependent transmembrane voltage over a wide pH range, from 1.8 to 9.5, for 30 µM concentrations of PCP. Lipid membranes were prepared from dioleoyl phosphatidylcholine. It was shown that three PCP molecular species were determining the transmembrane transfer of hydrogen ions: electrically neutral PCP molecules (HA), negatively charged pentachlorophenolate ions (A⁻) and negatively charged heterodimers (AHA⁻). It was found that at pH>9 the membrane electrical conductivity was determined by the transmembrane movement of A⁻ ions, whenever pHAHA⁻ species. Two new membrane surface reactions were proposed as supplementary mechanisms for the generation of AHA⁻ in addition to the formation of AHA⁻ by the recombination of HA and A⁻, HA + A⁻→ AHA⁻. These new reactions are, (i) 2HA → H⁺ + AHA⁻, and (ii) H20 + 2A⁻ → OH' + AHA⁻. Reaction (i) provides formation of membrane permeable heterodimers AHA⁻ at pH < < 5.5 and reaction (ii) at pH> > 5.5. The maximum surface density of AHA" heterodimers was 0.09 pmol/cm² • The rate constant of formation of AHA' by recombination, HA + A⁻ → AHA' was estimated to be k[subscript f] = 2.6xl0⁹ cm² mol⁻¹ s⁻¹ and the dissociation rate constant for AHA⁻ Further, it was possible to determine the rate constants of transmembrane translocation for A' and AHA⁻ ions to be k[subscript a] = 6.6x10⁻⁵ s⁻¹ and k[subscript aha] = 1200 S⁻¹, respectively.