Stein, I.H. “The Wyclif Manuscript in Florence.” Speculum 5 (1930): 95-97. [This is about MS Florence, Biblioteca Medicea Laurenziana Plutei.19.33 which .]
—. “Two Parallel Trains of Anti-Hierocratic Thought in the Fourteenth Century: Marsilius of Padua and John Wyclif.” Rivista Di Storia Della Philosophia 1997 (52.1): 91-110.
Manganese can be released to water by discharge from industrial facilities or as leachate from landfills and soil (US EPA, 1979, 1984; Francis & White, 1987; TRI91, 1993). Sea disposal of mine tailings and liquor is another source of manganese to the marine environment, particularly in tropical areas (Florence et al., 1994). Nriagu & Pacyna (1988) estimated that total worldwide anthropogenic inputs of manganese to aquatic ecosystems during 1983 ranged from 109 000 to 414 000 tonnes, with the predominant sources being domestic wastewater and sewage sludge disposal. In the USA, reported industrial discharges of manganese in 1991 ranged from 0 to 17.2 tonnes for surface water, from 0 to 57.3 tonnes for transfers to public sewage, and from 0 to 0.114 tonnes for underground injection (TRI91, 1993). An estimated total of 58.6 tonnes, or 1% of the total environmental release of manganese in the USA, was discharged to water in 1991 (TRI91, 1993). In 1996, the estimated release of manganese to water was 870 tonnes (TRI96, 1998).
Aita, Shuichi. “Negation in the Wycliffite Sermons.” Arthurian and Other Studies Presented to Shunichi Noguchi. Ed. T. Suzuki and T. Mukai. Woodbridge: Boydell and Brewer, 1993. 241-45.
Manganese is often transported in rivers adsorbed to suspended sediments. Most of the manganese from industrial sources (metallurgical and chemical plants) found in the Paraiba do Sul-Guandu River, Rio de Janeiro, Brazil, was bound to suspended particles (Malm et al., 1988). A positive correlation between manganese concentrations and suspended sediment levels has been reported for a wide variety of rivers in the United Kingdom (Laxen et al., 1984; Neal et al., 1998, 2000). The tendency of soluble manganese compounds to adsorb to soils and sediments can be highly variable, depending mainly on the cation exchange capacity and the organic composition of the soil (Hemstock & Low, 1953; Schnitzer, 1969; McBride, 1979; Curtin et al., 1980; Baes & Sharp, 1983; Kabata-Pendias & Pendias, 1984). Laxen et al. (1984) proposed that the "particulate" and "dissolved" phases for rivers and streams can be decoupled with weathering processes, leading to suspended sediment and influxes of Mn(II) species leaching from anoxic soil and groundwaters. The speciation in any particular river or stream will depend principally on the hydrogeological conditions of the catchment at time of sampling. Suspended sediment, with a manganese content dependent upon the catchment geology, will be mixed with Mn(II) species in varying proportions.
Primary chemical factors controlling sedimentary manganese cycling are the oxygen content of the overlying water, the penetration of oxygen into the sediments, and benthic organic carbon supply (Lynn & Bonatti, 1965; Grill, 1978; Balzer, 1982; Sundby et al., 1986; Hunt & Kelly, 1988). Manganese exchange between water and sediment is an interdependent process. A cycle between sediment and water is maintained, since dissolved Mn(II) is particle-reactive (Hunt, 1983). Once incorporated into sediments, solid-phase manganese oxides (manganese dioxide) undergo reduction to soluble Mn(II) during anaerobic decomposition of organic matter (Pohl et al., 1998). Release from sediment to water occurs by diffusion processes as a result of a steep Mn(II) concentration gradient across the sediment pore water and bottom water interface (Balzer, 1982; Kremling, 1983; Jung et al., 1996). Recycling at a redox boundary is involved in the formation of enriched manganese horizons. Manganese precipitating on the oxic side of a redox boundary consists of a Mn(IV) oxide. If the boundary is displaced towards the sediment surface or into the water column, the oxide undergoes rapid reduction and dissolution. Removal of Mn(II) by diffusion in the pore water is a slow process, and so supersaturation and precipitation of carbonate are likely to occur, transforming labilized oxide to stable carbonate. Under intermittently anoxic conditions, fixation of an enriched horizon may occur by precipitation of manganese dioxide from the water column during oxic periods, burial in sediment, and transformation to carbonate (Schaanning et al., 1988). A clear enrichment of dissolved manganese was observed at low salinities (
Combustion of MMT leads to the emission of manganese phosphates and manganese sulfate, with manganese oxides such as manganese tetroxide a minor component (NICNAS, 2003). The size of particles emitted to the atmosphere varies from 0.1 to 0.45 µm (Waldron, 1980). Combustion products of MMT also include manganese phosphate and manganese sulfide (Zayed et al., 1999; Zayed, 2001). One of the principal sources of inorganic manganese as a pollutant in the urban atmosphere is the combustion of MMT, particularly in areas of high traffic density (Sierra et al., 1998). MMT was used as a gasoline additive in the USA for a number of years, resulting in manganese emissions. Davis et al. (1988) found that motor vehicles made a significant contribution to levels of airborne manganese in areas such as southern California (around 40% of total airborne manganese) compared with, for example, central and northern California, where the addition of manganese to gasoline was much lower. According to a statistical model of source apportionment, the calculated average vehicular contribution of manganese in southern California was about 13 ng/m3, around 4 times the value calculated for both central and northern California.
In soils, manganese solubility is determined by two major variables: pH and redox potential. Water-soluble manganese in soils is directly proportional to pH, with oxidation state being another major determinant of manganese solubility. The lower oxidation state, Mn(II), predominates in reducing conditions, resulting in higher concentrations of dissolved manganese in flooded soils or other reducing situations (Stokes et al., 1988). This is normally reflected in higher manganese bioavailability in flooded soils; in some situations, however, there is competition by iron, and plant absorption of manganese is decreased or unaffected by flooding (Adriano, 1986). The oxidation state of manganese in soils and sediments can be altered by microbial activity (Geering et al., 1969; Francis, 1985). Geering et al. (1969) observed that Mn(II) in suspensions of silt or clay loams was oxidized by microorganisms, leading to precipitation of manganese minerals. Fungi are known to enhance the bioavailability of micronutrients. Accordingly, the solubilization of the sparingly soluble manganese dioxide by the fungus was reported by Altomare et al. (1999). Herzl & Roevros (1998) found that microbial uptake represented around 60% of the transfer of dissolved manganese to the particulate phase in the Scheldt estuary, Belgium. While microorganisms are believed to play an important role in the cycling of manganese in aquatic environments, specific microbial groups indigenous to these systems have not been well characterized (Thamdrup et al., 2000; Stein et al., 2001). There are two main mechanisms involved in the retention of manganese by soil. Firstly, through cation exchange reactions, manganese ions and the charged surface of soil particles form manganese oxides, hydroxides, and oxyhydroxides, which in turn form adsorption sites for other metals. Secondly, manganese can be adsorbed to other oxides, hydroxides, and oxyhydroxides through ligand exchange reactions (Evans, 1989).
MMT degradation in natural aquifers and sediment systems was determined to be very slow under anaerobic conditions. MMT has been found to be persistent in natural aquatic and soil environments in the absence of sunlight, with a tendency to sorb to soil and sediment particles. Calculated half-lives of MMT in aquatic and soil environments range from approximately 0.2 to 1.5 years at 25 °C (Garrison et al., 1995). In the presence of light, photodegradation of MMT is rapid, with identified products including a manganese carbonyl that readily oxidizes to manganese tetroxide (Garrison et al., 1995). MMT is photolysed rapidly by sunlight in the atmosphere, with a very short half-life of less than 2 min (Ter Haar et al., 1975; Garrison et al., 1995). MMT is photolysed rapidly in purified, distilled water exposed to sunlight, with degradation following first-order kinetics and a calculated half-life of less than 1 min (Garrison et al., 1995). Maneb released to water may be subject to abiotic degradation, with the rate of degradation dependent on the aeration of the water and the pH. In addition, maneb may undergo some photodegradation in sunlit water. Maneb is not expected to undergo significant volatilization from water. Mancozeb hydrolyses rapidly in water, with a half-life of less than 1–2 days at pH 5–9 (ATSDR, 2000).
The hydrophobicity of MMT (octanol–water partition coefficient [log ow] = 3.7) suggests that it can sorb to soil or sediment particles (Garrison et al., 1995). MMT was found to be stable in stream bottom sediments under anaerobic conditions. Photodegradation of MMT is not likely to occur in sediments, and MMT may equilibrate between the sediment, sediment pore water, and water column manganese (Garrison et al., 1995). Calumpang et al. (1993) reported a half-life of 2.9 days for mancozeb determined in a silty clay loam soil. In other studies, the half-life of maneb in soil was estimated to be between 20 and 60 days (Rhodes, 1977; Nash & Beall, 1980). Using chemical and physical properties, Beach et al. (1995) estimated the half-life of maneb and mancozeb in soils to be 70 days.