This report contains the collective views of an international group of experts and does not necessarily represent the decisions or the stated policy of the United Nations Environment Programme, the International Labour Organization, or the World Health Organization.

Concise International Chemical Assessment Document 63

First draft prepared by Mr P.D. Howe, Mr H.M. Malcolm, and Dr S. Dobson,
Centre for Ecology & Hydrology, Monks Wood, United Kingdom

Published under the joint sponsorship of the United Nations Environment Programme, the International Labour Organization, and the World Health Organization, and produced within the framework of the Inter-Organization Programme for the Sound Management of Chemicals.

World Health Organization

Geneva, 2004

The International Programme on Chemical Safety (IPCS), established in 1980, is a joint venture of the United Nations Environment Programme (UNEP), the International Labour Organization (ILO), and the World Health Organization (WHO). The overall objectives of the IPCS are to establish the scientific basis for assessment of the risk to human health and the environment from exposure to chemicals, through international peer review processes, as a prerequisite for the promotion of chemical safety, and to provide technical assistance in strengthening national capacities for the sound management of chemicals.

The Inter-Organization Programme for the Sound Management of Chemicals (IOMC) was established in 1995 by UNEP, ILO, the Food and Agriculture Organization of the United Nations, WHO, the United Nations Industrial Development Organization, the United Nations Institute for Training and Research, and the Organisation for Economic Co-operation and Development (Participating Organizations), following recommendations made by the 1992 UN Conference on Environment and Development to strengthen cooperation and increase coordination in the field of chemical safety. The purpose of the IOMC is to promote coordination of the policies and activities pursued by the Participating Organizations, jointly or separately, to achieve the sound management of chemicals in relation to human health and the environment.

WHO Library Cataloguing-in-Publication Data

Manganese and its compounds : environmental aspects.

(Concise international chemical assessment document ; 63)

1.Manganese - toxicity 2.Risk assessment 3.Environmental
exposure I.International Programme on Chemical Safety
II.Series

ISBN 92 4 153063 4           (LC/NLM Classification: QV 290)

ISSN 1020-6167

©World Health Organization 2004

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Technically and linguistically edited by Marla Sheffer, Ottawa, Canada, and printed by Wissenchaftliche Verlagsgesellschaft mbH, Stuttgart, Germany

FOREWORD

Concise International Chemical Assessment Documents (CICADs) are the latest in a family of publications from the International Programme on Chemical Safety (IPCS) — a cooperative programme of the World Health Organization (WHO), the International Labour Organization (ILO), and the United Nations Environment Programme (UNEP). CICADs join the Environmental Health Criteria documents (EHCs) as authoritative documents on the risk assessment of chemicals.

International Chemical Safety Cards on the relevant chemical(s) are attached at the end of the CICAD, to provide the reader with concise information on the protection of human health and on emergency action. They are produced in a separate peer-reviewed procedure at IPCS. They may be complemented by information from IPCS Poison Information Monographs (PIM), similarly produced separately from the CICAD process.

CICADs are concise documents that provide summaries of the relevant scientific information concerning the potential effects of chemicals upon human health and/or the environment. They are usually based on selected national or regional evaluation documents or on existing EHCs. Before acceptance for publication as CICADs by IPCS, these documents undergo extensive peer review by internationally selected experts to ensure their completeness, accuracy in the way in which the original data are represented, and the validity of the conclusions drawn.

The primary objective of CICADs is characterization of hazard and dose–response from exposure to a chemical. CICADs are not a summary of all available data on a particular chemical; rather, they include only that information considered critical for characterization of the risk posed by the chemical. The critical studies are, however, presented in sufficient detail to support the conclusions drawn. For additional information, the reader should consult the identified source documents upon which the CICAD has been based.

Risks to human health and the environment will vary considerably depending upon the type and extent of exposure. Responsible authorities are strongly encouraged to characterize risk on the basis of locally measured or predicted exposure scenarios. To assist the reader, examples of exposure estimation and risk characterization are provided in CICADs, whenever possible. These examples cannot be considered as representing all possible exposure situations, but are provided as guidance only. The reader is referred to EHC 170.¹

While every effort is made to ensure that CICADs represent the current status of knowledge, new information is being developed constantly. Unless otherwise stated, CICADs are based on a search of the scientific literature to the date shown in the executive summary. In the event that a reader becomes aware of new information that would change the conclusions drawn in a CICAD, the reader is requested to contact IPCS to inform it of the new information.

Procedures

The flow chart on page 2 shows the procedures followed to produce a CICAD. These procedures are designed to take advantage of the expertise that exists around the world — expertise that is required to produce the high-quality evaluations of toxicological, exposure, and other data that are necessary for assessing risks to human health and/or the environment. The IPCS Risk Assessment Steering Group advises the Coordinator, IPCS, on the selection of chemicals for an IPCS risk assessment based on the following criteria:

• there is the probability of exposure; and/or
• there is significant toxicity/ecotoxicity.

Thus, it is typical of a priority chemical that

• it is of transboundary concern;
• it is of concern to a range of countries (developed, developing, and those with economies in transition) for possible risk management;
• there is significant international trade;
• it has high production volume;
• it has dispersive use.

The Steering Group will also advise IPCS on the appropriate form of the document (i.e., a standard CICAD or a de novo CICAD) and which institution bears the responsibility of the document production, as well as on the type and extent of the international peer review.

The first draft is usually based on an existing national, regional, or international review. When no appropriate source document is available, a CICAD may be produced de novo. Authors of the first draft are usually, but not necessarily, from the institution that developed the original review. A standard outline has been developed to encourage consistency in form. The first draft undergoes primary review by IPCS to ensure that it meets the specified criteria for CICADs.

The second stage involves international peer review by scientists known for their particular expertise and by scientists selected from an international roster compiled by IPCS through recommendations from IPCS national Contact Points and from IPCS Participating Institutions. Adequate time is allowed for the selected experts to undertake a thorough review. Authors are required to take reviewers’ comments into account and revise their draft, if necessary. The resulting second draft is submitted to a Final Review Board together with the reviewers’ comments. At any stage in the international review process, a consultative group may be necessary to address specific areas of the science. When a CICAD is prepared de novo, a consultative group is normally convened.

The CICAD Final Review Board has several important functions:

• to ensure that each CICAD has been subjected to an appropriate and thorough peer review;
• to verify that the peer reviewers’ comments have been addressed appropriately;
• to provide guidance to those responsible for the preparation of CICADs on how to resolve any remaining issues if, in the opinion of the Board, the author has not adequately addressed all comments of the reviewers; and
• to approve CICADs as international assessments.

Board members serve in their personal capacity, not as representatives of any organization, government, or industry. They are selected because of their expertise in human and environmental toxicology or because of their experience in the regulation of chemicals. Boards are chosen according to the range of expertise required for a meeting and the need for balanced geographic representation.

Board members, authors, reviewers, consultants, and advisers who participate in the preparation of a CICAD are required to declare any real or potential conflict of interest in relation to the subjects under discussion at any stage of the process. Representatives of nongovernmental organizations may be invited to observe the proceedings of the Final Review Board. Observers may participate in Board discussions only at the invitation of the Chairperson, and they may not participate in the final decision-making process.

This CICAD on manganese and its compounds (environmental aspects) was based primarily on the report Toxicological profile for manganese (update), prepared by the Agency for Toxic Substances and Disease Registry of the US Department of Health and Human Services (ATSDR, 2000). Secondary sources of information included CICAD No. 12 on manganese and its compounds (IPCS, 1999a) and data identified following a comprehensive literature search of relevant databases conducted up to December 2002 to identify any relevant references published subsequent to those incorporated in these two reports. For information regarding the assessment of human health effects of manganese, the reader should refer to CICAD No. 12 (IPCS, 1999a). Manganese fungicides have been referred to in the document for source and fate information only, and no attempt has been made to evaluate this group of chemicals for environmental effect. Information on the preparation and peer review of the source document is presented in Appendix 1. Information on the peer review of this CICAD is presented in Appendix 2. This CICAD was considered and approved as an international assessment at a meeting of the Final Review Board, held in Varna, Bulgaria, on 8–11 September 2003. Participants at the Final Review Board meeting are presented in Appendix 3. International Chemical Safety Cards on selected manganese compounds (ICSCs 174, 175, 290, 754, 977, 1169, and 1398), produced by the International Programme on Chemical Safety in a separate, peer-reviewed process (IPCS, 1999b,c, 2001, 2003a,b,c,d), have also been reproduced in this document.

Manganese (Mn) is a naturally occurring element that is found in rock, soil, and water. It is ubiquitous in the environment and comprises about 0.1% of the Earth’s crust. Crustal rock is a major source of manganese found in the atmosphere. Ocean spray, forest fires, vegetation, and volcanic activity are other major natural atmospheric sources of manganese. Important sources of dissolved manganese are anaerobic environments where particulate manganese oxides are reduced, the direct reduction of particulate manganese oxides in aerobic environments, the natural weathering of Mn(II)-containing minerals, and acidic environments. The major pool of manganese in soils originates from crustal sources, with other sources including direct atmospheric deposition, wash-off from plant and other surfaces, leaching from plant tissues, and the shedding or excretion of material such as leaves, dead plant and animal material, and animal excrement. The major anthropogenic sources of environmental manganese include municipal wastewater discharges, sewage sludge, mining and mineral processing, emissions from alloy, steel, and iron production, combustion of fossil fuels, and, to a much lesser extent, emissions from the combustion of fuel additives.

Manganese is released to air mainly as particulate matter, and the fate and transport of the particles depend on their size and density and on wind speed and direction. Some manganese compounds are readily soluble in water. Manganese exists in the aquatic environment in two main forms: Mn(II) and Mn(IV). Movement between these two forms occurs via oxidation and reduction reactions that may be abiotic or microbially mediated. The environmental chemistry of manganese is largely governed by pH and redox conditions; Mn(II) dominates at lower pH and redox potential, with an increasing proportion of colloidal manganese oxyhydroxides above pH 5.5 in non-dystrophic waters. 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. Manganese in soil can migrate as particulate matter to air or water, or soluble manganese compounds can be leached from the soil. In soils, manganese solubility is determined by two major variables: pH and redox potential.

Manganese in water can be significantly bioconcentrated by aquatic biota at lower trophic levels. Bioconcentration factors (BCFs) of 2000–20 000 for marine and freshwater plants, 2500–6300 for phytoplankton, 300–5500 for marine macroalgae, 800–830 for intertidal mussels, and 35–930 for fish have been estimated. Uptake of manganese by aquatic invertebrates and fish significantly increases with temperature and decreases with pH, whereas dissolved oxygen has no significant effect. Uptake of manganese has been found to increase with decreasing salinity.

Manganese concentrations in air tend to be lowest in remote locations (about 0.5–14 ng/m³ on average), higher in rural areas (40 ng/m³ on average), and still higher in urban areas (about 65–166 ng/m³ on average). Manganese concentrations in air tend to be highest in source-dominated areas, where values can reach 8000 ng/m³. Annual averages of manganese concentrations may rise to 200–300 ng/m³ in air near foundries and to over 500 ng/m³ in air near ferro- and silicomanganese industries.

Concentrations of dissolved manganese in natural waters that are essentially free of anthropogenic inputs can range from 10 to >10 000 µg/litre. However, dissolved manganese concentrations in natural surface waters rarely exceed 1000 µg/litre and are usually less than 200 µg/litre.

Manganese concentrations in river sediments ranged from 410 to 6700 mg/kg dry weight; sediment from an urban lake receiving inputs from industrial and residential areas, as well as windborne dust from old mine dumps, contained manganese at concentrations ranging up to 13 400 mg/kg dry weight. Sediment manganese concentrations of 100–1000 mg/kg dry weight have been reported for intertidal mudflats; similar total manganese values were found in the northern Adriatic Sea. Surface sediments in the Baltic Sea contained manganese at mean concentrations of 3550–8960 mg/kg dry weight; the high manganese concentrations were thought to be due to ferromanganese concretions and riverine loads.

Natural ("background") levels of total manganese in soil range from <1 to 4000 mg/kg dry weight, with mean values around 300–600 mg/kg dry weight.

Mean manganese concentrations in seaweed range from 130 to 735 mg/kg dry weight, whereas concentrations in shellfish range from 3 to 660 mg/kg dry weight; higher concentrations in shellfish are associated with manganese-rich sediment. Concentrations of manganese found in tissues of marine and freshwater fish tend to range from <0.2 to 19 mg/kg dry weight. Higher manganese concentrations — above 100 mg/kg dry weight — have been reported for fish in polluted surface waters.

Concentrations of manganese in terrestrial plants tend to range from 20 to 500 mg/kg. Members of the Ericaceae family, which includes blueberries, are regarded as manganese accumulators. There are numerous reports of foliar manganese levels in excess of 2000–4000 mg/kg. Mean manganese concentrations in birds’ eggs from a variety of geographical areas range from 1 to 5 mg/kg dry weight, mean liver concentrations range from 3 to 11 mg/kg dry weight, and mean feather concentrations range from 0.3 to 40 mg/kg dry weight. Mean manganese concentrations of up to 17 mg/kg dry weight have been found in tissues (liver, kidney, and whole body) from a variety of reptiles and wild mammals.

Manganese is an essential nutrient for microorganisms, plants, and animals. Nutritional manganese requirements for terrestrial plants are around 10–50 mg/kg tissue. Critical nutritional levels vary widely between species and among cultivars of a species. Calcareous soils, especially those with poor drainage and high organic matter, are the types of soil that produce manganese-deficient plants.

Most toxicity tests have been carried out using ionic manganese. Little is known about the aquatic toxicity of colloidal, particulate, and complexed manganese; in general, however, toxicities of metals bound into these forms are assumed to be less than those of the aquo-ionic forms. For algae and protozoa, there is a wide range of toxicity values; the most sensitive species appear to be the marine diatom Ditylum brightwellii, with a 5-day EC₅₀, based on growth inhibition, of 1.5 mg/litre, and a freshwater alga Scenedesmus quadricauda, with a 12-day EC₅₀, based on total chlorophyll reduction, of 1.9 mg/litre. Tests on aquatic invertebrates reveal 48-h LC₅₀/EC₅₀ values ranging from 0.8 mg/litre (Daphnia magna) to 1389 mg/litre (Crangonyx pseudogracilis), the lowest LC₅₀ being observed under soft water conditions (25 mg calcium carbonate/litre). A significant reduction in survival and hatching of yellow crab (Cancer anthonyi) embryos at >0.01 mg manganese/litre was found in 7-day tests in seawater. For fish, 96-h LC₅₀s range from 2.4 mg manganese/litre for coho salmon (Oncorhynchus kisutch) to 3350 mg/litre for Indian catfish (Heteropneustes fossilis), with the lowest LC₅₀ values obtained under soft water conditions (25 mg calcium carbonate/litre). Significant embryonic mortality was observed in rainbow trout (Oncorhynchus mykiss) eggs exposed to 1 mg manganese sulfate/litre for 29 days. A single embryo-larval test with a 7-day LC₅₀ of 1.4 mg manganese/litre was identified for amphibians. Acute toxicity in aquatic invertebrates and fish decreased with increasing water hardness; the addition of chelating agents can reduce the toxicity of manganese. There is evidence that manganese can protect organisms against the effects of more toxic metals.

In the field, the high frequency of blue crabs (Callinectes sapidus) with shell disease (lesions) in a metal-contaminated estuary was ascribed to manganese toxicity, and the deposition of manganese dioxide on the gills of Norway lobster (Nephrops norvegicus) gave rise to a brown or black discoloration of the gills and black corroded areas on the carapace following hypoxic conditions in the south-east Kattegat, Sweden. Increased mortality of rainbow trout (Oncorhynchus mykiss) at a hatchery was found to be positively correlated with manganese concentration (<0.5–1 mg/litre). Acid precipitation has caused acid episodes and elevated concentrations of metals. Cage experiments with yearling brown trout (Salmo trutta) showed that pH (4.5–5.4) and the concentration of labile inorganic manganese (0.1–0.4 mg/litre) explained all of the observed mortality.

Symptoms of manganese toxicity to terrestrial plants vary widely with species and include marginal chloroses, necrotic lesions, and distorted development of the leaves. Toxic manganese concentrations in crop plant tissues vary widely, with critical values ranging from 100 to 5000 mg/kg. Manganese toxicity is a major factor limiting crop growth on acidic, poorly drained, or steam-sterilized mineral soils. There is a wide range of variation in tolerance to manganese between and within plant species. Factors affecting manganese tolerance include genotype (inter- and intraspecific variation), silicon concentration, temperature, light intensity, physiological leaf age, microbial activity, and the characteristics of the rhizosphere.

Surface freshwater data suggest that higher manganese concentrations occur during periods of higher stream flow, such as spring runoff, and lower concentrations tend to occur downstream of lakes that act as settling areas for sediment. Soft water streams, rivers, and lakes appear to be the most sensitive freshwater environments, with laboratory tests and field observations showing that dissolved manganese concentrations of around 1 mg/litre can cause toxic effects in aquatic organisms. An overall guidance value for the protection of 95% of species with 50% confidence was derived at 0.2 mg manganese/litre for soft waters for the freshwater environment. Other factors, such as acid precipitation, acid mine drainage, land use, and municipal wastewater discharges, can increase dissolved manganese levels and thus increase the risk to sensitive species, especially in soft water areas. Evaluation of the likely toxicity of manganese to organisms in the field has to take account of speciation conditions in both the test and the specific field area. In the marine environment, manganese can be taken up and accumulated by organisms during hypoxic releases of dissolved manganese from manganese-rich sediments. Even taking into account the possible mitigating effects of suspended sediment, salinity, and oxygen levels in natural environments, adverse effects in the field have been observed. An overall guidance value for the protection of 95% of species with 50% confidence was derived at 0.3 mg manganese/litre for the marine environment.

When evaluating the risk to the terrestrial environment from anthropogenic releases of manganese, account must be taken of local natural ("background") levels, which are in turn controlled by a variety of physical and chemical parameters. Different communities and ecosystems would also respond differently, depending on their "normal" exposure to manganese. For these reasons, deriving a single guidance value for the terrestrial environment is inappropriate.

Table 1 lists common synonyms and other relevant information on the chemical identity and properties of manganese and several of its most important compounds. Manganese is a naturally occurring element that is found in rock, soil, water, and food. It can exist in 11 oxidation states ranging from –3 to +7, but the most common ones are +2 (e.g., manganese chloride [MnCl₂]), +4 (e.g., manganese dioxide [MnO₂]), and +7 (e.g., potassium permanganate [KMnO₄]). Manganese and its compounds can exist as solids in the soil and as solutes or small particles in water. Most manganese salts are readily soluble in water, with only the phosphate and the carbonate having low solubilities. The manganese oxides (manganese dioxide and manganese tetroxide) are poorly soluble in water. Manganese can also be present in small dust-like particles in the air.

Table 1: Chemical identity of manganese and its compounds.^a

^a Adapted from ATSDR (2000). All information obtained from Sax & Lewis (1987), except where noted.

^j Garrison et al. (1995).

Additional physical/chemical properties for manganese and select manganese compounds are presented in the International Chemical Safety Cards reproduced in this document.

Atomic absorption spectrophotometric analysis is the most widely used method for determining manganese in biological materials and environmental samples. Fluorimetric, colorimetric, neutron activation analysis, and plasma atomic emission techniques are also recommended for measuring manganese in such samples. Inductively coupled plasma (ICP) atomic emission analysis is frequently employed for multianalyte analyses that include manganese. In most cases, distinguishing between different oxidation states of manganese is impossible, so total manganese is measured. The detection limits of these methods range from <0.01 to 0.2 mg/kg for biological tissues and fluids, from 5 to 10 µg/m³ for air, and from 0.01 to 50 µg/litre for water (Kucera et al., 1986; Abbasi, 1988; Lavi et al., 1989; Mori et al., 1989; Chin et al., 1992; ATSDR, 2000).

Determination of manganese requires an acid extraction/digestion step before analysis. The details vary with the specific characteristics of the sample, but treatment usually involves heating in nitric acid, oxidation with hydrogen peroxide, and filtration and/or centrifugation to remove insoluble matter (ATSDR, 2000).

Meneses et al. (1999) and Llobet et al. (2002) used ICP mass spectrometry with a detection limit of 0.02 mg/kg for soil and herbage. Pandey et al. (1998) used a sequential ICP optical emission spectrometer with an ultrasonic nebulizer for atmospheric particulates at a detection limit of 0.001 µg/litre, whereas ICP with atomic emission spectrophotometry was used for atmospheric particulates (Brewer & Belzer, 2001; Espinosa et al., 2001), sediment (Leivuori, 1998; Leivuori & Vallius, 1998), shellfish (Blackmore et al., 1998; Blackmore, 1999; Rainbow & Blackmore, 2001), feathers (Connell et al., 2002), and liver tissue (Mason & Stephenson, 2001).

Beklemishev et al. (1997) used the catalytic kinetic method for analysis of manganese in water. The method relies on an indicator reaction that is catalysed by Mn(II) (the oxidation of 3,3',5,5'-tetramethylbenzidine by potassium periodate [KIO₄]) and is carried out on the surface of a paper-based sorbent. The method has a detection limit (0.005 µg/litre) that is much lower than those of other, more established methods.

A nuclear magnetic resonance method (Kellar & Foster, 1991) and a method using on-line concentration analysis (Resing & Mottl, 1992) were used to determine both free and complexed manganese ions in aqueous media. The latter method was highly sensitive, with a detection limit of 36 pmol/litre (1.98 ng/litre when concentrating 15 ml of seawater). A similar detection limit was achieved by Sarzanini et al. (2001) for seawater using flow-injection preconcentration coupled with electrothermal atomic absorption spectrometry.

The technique of diffusive gradients in thin films (DGT) has been used for in situ trace manganese speciation measurements. A concentration–depth profile of labile manganese was obtained in a stratified estuary by deployment of a string of DGT devices across the redoxcline (Denney et al., 1999). Gauthreaux et al. (2001) used a modified sequential extraction procedure to speciate the chemical forms of manganese in sediment using flame atomic absorption spectrophotometry. Concentrations were determined in five different fractions for each sample: manganese in the exchangeable form, manganese bound to carbonates, manganese bound to manganese/iron oxides, manganese bound to organic matter, and manganese in the residual form. Techniques such as X-ray absorption near-edge structure spectroscopy enable in situ quantification of the oxidation state of manganese (Bargar et al., 2000).

Manganese is ubiquitous in the environment. It comprises about 0.1% of the Earth's crust (NAS, 1973; Graedel, 1978). Manganese does not occur naturally as a base metal but is a component of more than 100 minerals, including various sulfides, oxides, carbonates, silicates, phosphates, and borates (NAS, 1973). The most commonly occurring manganese-bearing minerals include pyrolusite (manganese dioxide), rhodocrosite (manganese carbonate [MnCO₃]), rhodonite (manganese silicate), and hausmannite (manganese tetroxide [Mn₃O₄]) (NAS, 1973; Windholz, 1983; US EPA, 1984; HSDB, 1998).

Ferromanganese minerals such as biotite mica and amphiboles contain large amounts of manganese, and manganese-rich nodules (especially found in the North-east Pacific; Schiele, 1991) have been identified on the seafloor in conjunction with cobalt, nickel, and copper (Reimer, 1999; Ahnert & Borowski, 2000). Similarly, manganese crusts occur as pavement-like encrustations of ferromanganese oxides on exposed abyssal hard substrates related to all types of submarine elevations (Ahnert & Borowski, 2000). Black smokers (hydrothermal vents on the seafloor releasing predominantly iron and sulfide) are an additional source releasing manganese into the (oceanic) hydrosphere (Gamo et al., 2001). Important sources of manganese include soils, sediments, and metamorphic and sedimentary rocks (Reimer, 1999).

Crustal rock is a major source of manganese found in the atmosphere. Ocean spray, forest fires, vegetation, and volcanic activity are other major natural sources of manganese in the atmosphere (Schroeder et al., 1987; Stokes et al., 1988). Stokes et al. (1988) estimated that two-thirds of manganese air emissions were from natural sources. Atmospheric particulate matter collected in the Antarctic indicated that manganese was derived from either crustal weathering or the ocean (Zoller et al., 1974). Air erosion of dusts and soils is also an important atmospheric source of manganese, but no quantitative estimates of manganese release to air from this source were identified (US EPA, 1984). An important source of dissolved manganese is anaerobic environments where particulate manganese oxides are reduced, such as some soils and sediments, wetlands, and the anaerobic hypolimnia of lakes and fjords. Other possible sources include the direct reduction of particulate manganese oxides in aerobic environments by organics, with or without ultraviolet light, the natural weathering of Mn(II)-containing minerals, and acid drainage and other acidic environments. The major pool of manganese in soils originates from crustal sources. Addition of manganese to soils can also result from direct atmospheric deposition, wash-off from plant and other surfaces, leaching from plant tissues, and the shedding or excretion of material such as leaves, dead plant and animal material, and animal excrement (Stokes et al., 1988).

The major anthropogenic sources of environmental manganese include municipal wastewater discharges, sewage sludge, mining and mineral processing (particularly nickel), emissions from alloy, steel, and iron production, combustion of fossil fuels, and, to a much lesser extent, emissions from the combustion of fuel additives.

The manganese content in ore produced worldwide was estimated to be 8.8 million tonnes in 1986. Production levels of manganese ore and its total manganese metal content remained nearly the same through 1990 (US Department of the Interior, 1993). Levels of ore produced worldwide in 1995, 1996, and 1997 declined slightly, with total manganese metal content declining proportionately to 8.0, 8.1, and 7.7 million tonnes, respectively (US Department of the Interior, 1996, 1998). Sites of substantial workable manganese–iron deposits include the former USSR, South and North Africa, South America, India, and China (Schiele, 1991). Most manganese is mined in open-pit or shallow mines (NAS, 1973). Although modern steelmaking technologies call for lower unit consumption of manganese, worldwide demand for steel is projected to increase moderately in the future, particularly in developing countries (US Department of the Interior, 1995, 1998). The demand for manganese in other industries (e.g., dry-cell battery manufacturing) might also increase, but the overall effect of these other uses on global trends in manganese production and use is minor (EM, 1993; US Department of the Interior, 1995, 1998).

Manganese compounds are produced from manganese ores or from manganese metal. Metallic manganese (ferromanganese) is used principally in steel production along with cast iron and superalloys to improve hardness, stiffness, and strength (NAS, 1973; US EPA, 1984; HSDB, 1998). The predominant portion (approximately 90%) of manganese is processed into ferromanganese in blast furnaces (Schiele, 1991). Manganese compounds have a variety of uses. Manganese dioxide is commonly used in the production of dry-cell batteries, matches, fireworks, porcelain and glass-bonding materials, and amethyst glass; it is also used as the starting material for the production of other manganese compounds (NAS, 1973; Venugopal & Luckey, 1978; US EPA, 1984). Manganese chloride is used as a precursor for other manganese compounds, as a catalyst in the chlorination of organic compounds, in animal feed to supply essential trace minerals, and in dry-cell batteries (US EPA, 1984; HSDB, 1998). Manganese sulfate (MnSO₄) is used primarily as a fertilizer and as a livestock supplement; it is also used in some glazes, varnishes, ceramics, and fungicides (Windholz, 1983; US EPA, 1984; HSDB, 1998). Maneb (manganese ethylene-bis-dithiocarbamate) is used as a broad-spectrum contact fungicide and is also used for seed treatment of small grains such as wheat. Maneb is therefore a potential source of manganese in soil and plants (Ferraz et al., 1988; Ruijten et al., 1994). Potassium permanganate is used as an oxidizing agent, disinfectant, and antialgal agent; for metal cleaning, tanning, and bleaching; as a purifier in water and waste treatment plants; and as a preservative for fresh flowers and fruits (HSDB, 1998). The organomanganese compound MMT (methylcyclopentadienyl manganese tricarbonyl), an antiknock additive in unleaded gasoline, is produced by the addition of molten sodium metal to methylcyclopentadiene to give methylcyclopentadienylsodium. Anhydrous manganese dichloride is then added to afford methylcyclopentadienylmanganese, which is subsequently reacted with carbon monoxide to give MMT (NAS, 1973; US EPA, 1984; Sax & Lewis, 1987; HSDB, 1998; Kirk & Othmer, 2001). MMT has been approved for use in Argentina, Australia, Bulgaria, the USA, France, and the Russian Federation and has been conditionally approved for use in New Zealand (Zayed et al., 1999; Zayed, 2001); more recently, Ethyl Corp. (a major producer of MMT) noted that MMT is now sold in 25 countries (Kaiser, 2003).

The main anthropogenic sources of manganese release to air are industrial emissions (such as ferroalloy production and iron and steel foundries, power plants, and coke ovens), combustion of fossil fuels, and re-entrainment of manganese-containing soils (Lioy, 1983; US EPA, 1983, 1984, 1985a,b; Ruijten et al., 1994; ATSDR, 2000). Problems with air pollution — especially dust and smoke containing manganese dioxide and manganese tetroxide — arise during the mining, crushing, and smelting of ores as well as during steel production (Schiele, 1991). Approximately 2 tonnes of manganese ore are required to make 1 tonne of ferromanganese alloy (NAS, 1973). Steel emissions were found to be the predominant source of manganese in urban particulate matter (Sweet et al., 1993). Manganese can also be released to the air during other anthropogenic processes, such as welding and fungicide application (Ferraz et al., 1988; MAK, 1994; Ruijten et al., 1994). Nriagu & Pacyna (1988) estimated that total worldwide emissions of manganese in 1983 ranged from 10 560 to 65 970 tonnes, with the predominant sources being coal combustion, secondary non-ferrous metal production, and sewage sludge incineration. Total emissions to air from anthropogenic sources in the USA were estimated to be 16 400 tonnes in 1978, with about 80% (13 200 tonnes) from industrial facilities and 20% (3200 tonnes) from fossil fuel combustion (US EPA, 1983). Air emissions by US industrial sources reportable to the Toxics Release Inventory (TRI) for 1987 totalled 1200 tonnes (TRI87, 1989). In 1991, air emissions from TRI facilities in the USA ranged from 0 to 74 tonnes, with several US states reporting no emissions (TRI91, 1993). Estimated releases of manganese to air in 1996 were 4000 tonnes, representing 15% of total environmental releases (TRI96, 1998). Figures in Table 2 (see section 6) show decreasing emissions of manganese to air in the USA as a result of air pollution control.

Table 2: Average levels of manganese in air.

^a Adapted from Stokes et al. (1988).

^b Adapted from US EPA (1984).

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 Canada, MMT use as a fuel additive has gradually increased since 1977. Manganese emissions from gasoline combustion rose sharply from 1977 through the early 1980s, reaching an estimated 220 tonnes by 1985 (Jaques, 1984). In 1990, lead was completely replaced by MMT in gasoline in Canada (Loranger & Zayed, 1994). MMT use peaked in 1989 at over 400 tonnes, which was more than twice the usage in 1983 and 1.5 times the usage in 1986. MMT use declined to about 300 tonnes by 1992, owing to reductions in its concentration in gasoline. However, ambient monitoring data for manganese in Canadian cities without industrial sources for the 1989–1992 period did not reflect this peak in MMT use. Air manganese levels (PM_2.5, or particulate matter with an aerodynamic diameter less than or equal to 2.5 µm) remained constant at 11–13 ng/m³ for small cities and 20–25 ng/m³ for large cities (Health Canada, 1994; Egyed & Wood, 1996). Manganese emission levels can vary depending on the concentration of MMT in gasoline and gasoline usage patterns. One study reported a correlation between atmospheric manganese concentrations in 1990 air samples and traffic density in Montreal, Canada (Loranger & Zayed, 1994). However, a later study by these investigators reported that atmospheric manganese concentrations in Montreal decreased in 1991 and 1992, despite an estimated 100% increase in manganese emission rates from MMT in gasoline (Loranger & Zayed, 1994). Another study suggested that the high manganese levels in Montreal were, in part, due to the presence of a silico- and ferromanganese facility that ceased operation in 1991 (Egyed & Wood, 1996).

It is clear that the contribution of MMT to overall manganese levels in the environment is complex. The contribution of MMT to atmospheric manganese concentrations is difficult to establish, since it may be masked by more substantial variation associated with other industrial activities as well as road dust and windblown dust (Bankovitch et al., 2003). However, even though manganese may be a small percentage of total suspended particulate matter measured in cities, such as Montreal, the contribution of MMT to air manganese levels could be significant, in that it may account for stable manganese levels in the face of declining total suspended particulate concentrations. Factors such as unfavourable meteorological conditions and high traffic density could lead to an increase in manganese levels (PM_2.5) attributable to MMT (Wallace & Slonecker, 1997; Davis et al., 1998).

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).

Land disposal of manganese-containing wastes is the principal source of manganese releases to soil. Nriagu & Pacyna (1988) estimated that total worldwide anthropogenic releases of manganese to soils during 1983 ranged from 706 000 to 2 633 000 tonnes, with the predominant source being coal fly ash. In 1991, reported industrial releases to land in the USA ranged from 0 to 1000 tonnes. More than 50% of the total environmental release of manganese (3753 tonnes) was to land (TRI91, 1993). Estimated releases of manganese to soil in 1996 were 21 600 tonnes, representing 80% of total environmental releases (TRI96, 1998).

Elemental manganese and inorganic manganese compounds have negligible vapour pressures but can exist in air as suspended particulate matter derived from industrial emissions or the erosion of soils (US EPA, 1984). In the troposphere, manganese is likely to be found in oxide, sulfate, or nitrate forms or as mineral complexes related to its natural origin in soil or rock (Stokes et al., 1988). Manganese-containing particles are removed from the atmosphere mainly by gravitational settling or by rain (US EPA, 1984).

Soil particulate matter containing manganese can be transported in air. The fate and transport of manganese in air are largely determined by the size and density of the particles and by wind speed and direction. An estimated 80% of the manganese in suspended particulate matter is associated with particles with a mass median equivalent diameter (MMED) of <5 µm, and 50% of this manganese is estimated to be associated with particles that are <2 µm in MMED. Whether these data are for particles in urban or rural areas is unclear. However, it is known that the size of manganese particles in the air tends to vary by source; small particles dominate around ferromanganese and dry-cell battery plants, whereas large particles tend to predominate near mining operations (WHO, 1999). Airborne particles (>2 µm) collected over the oceans contained a mean manganese concentration of 1338 mg/kg (Lee & Duffield, 1979). Based on these data, widespread airborne distribution would be expected (IPCS, 1981). Fergusson & Stewart (1992) stated that unlike deposition of other metals, such as copper, lead, cadmium, and zinc, manganese deposition showed little spatial variation between urban and rural areas; however, Mielke et al. (2002) reported a 4-fold increase in manganese concentrations between rural and urban areas of New Orleans, Louisiana, USA. Very little information is available on atmospheric reactions of manganese (US EPA, 1984). Although manganese can react with sulfur dioxide and nitrogen dioxide, the occurrence of such reactions in the atmosphere has not been demonstrated.

The arithmetically averaged annual wet flux of manganese was 1190 µg/m² for Chesapeake Bay, USA (Scudlark et al., 1994), and 1900 µg/m² for a Scottish sea loch (Hall et al., 1996). Rates of atmospheric deposition of manganese into the western Mediterranean Sea (northwestern Corsica) between 1985 and 1987 ranged between 0.0023 and 0.0072 µg/cm² per day. Sporadic but intense Saharan sandstorms were responsible for the highest atmospheric deposition of manganese (Remoudaki et al., 1991). A manganese deposition rate of 350 µg/m² per day was calculated for Burnaby Lake, British Columbia, Canada. The estimated annual deposition rate was estimated to be 7.7 tonnes for the entire watershed (Brewer & Belzer, 2001).

Manganese exists in the aquatic environment in two main forms: Mn(II) and Mn(IV). Transition between these two forms occurs via oxidation and reduction reactions that may be abiotic or microbially mediated (Nealson, 1983; Thamdrup et al., 2000; Heal, 2001). The environmental chemistry of manganese is largely governed by pH and redox conditions; Mn(II) dominates at lower pH and redox potential, with an increasing proportion of colloidal manganese oxyhydroxides above pH 5.5 in non-dystrophic waters (LaZerte & Burling, 1990). In waters receiving acid mine drainage, dissolved manganese concentrations were <40 µg/litre above pH 5.5; however, below pH 3, dissolved manganese concentrations ranged from 250 to 4400 µg/litre (Filipek et al., 1987). Similarly, in another study, sediment concentrations of manganese decreased from 400 mg/kg at pH 5.6–5.9 to 8 mg/kg below pH 3 due to manganese dissolution influenced by acid mine drainage (Cherry et al., 2001).

A complex series of oxidation/precipitation and adsorption reactions occurs when Mn(II) is present in aerobic environments, which eventually renders the manganese biologically unavailable as insoluble manganese dioxide. However, the kinetics of Mn(II) oxidation are slow in waters with pH below 8.5 (Zaw & Chiswell, 1999). The time required for the oxidation and precipitation of manganese ranges from days in natural waters to years in synthetic waters (Stokes et al., 1988). However, oxidation rates of manganese increase with increasing pH or the presence of catalytic surfaces such as manganese dioxide (Huntsman & Sunda, 1980). In a stream receiving manganese-rich inflows caused by acid mine drainage, there was rapid oxidation and precipitation of manganese oxides (Scott et al., 2002). The sequence of reactions involving the oxidation of Mn(II) and subsequent precipitation as manganese dioxide includes simultaneous occurrence of several manganese forms (i.e., dissolved Mn(II), hydrous oxides of Mn(III), Mn(II) adsorbed to particulates, and Mn(II)–ligand complexes), with individual concentrations dependent on factors that include pH, inorganic carbon, organic carbon, sulfate, chloride, temperature, and time (Stokes et al., 1988). In groundwater with low oxygen levels, Mn(IV) can be reduced both chemically and bacterially to the Mn(II) oxidation state (Jaudon et al., 1989).

There is little evidence for manganese–organic associations in natural waters, with manganese only weakly bound to dissolved organic carbon (L'Her Roux et al., 1998). Hence, organic complexation does not play a major role in controlling manganese speciation in natural waters. Field studies have confirmed that organically bound manganese is minor, even with high natural dissolved organic carbon levels (Laxen et al., 1984). The Mn(II) ion is more soluble than Mn(IV); therefore, manganese will tend to become more bioavailable with decreasing pH and redox potential (Heal, 2001). The presence of chlorides, nitrates, and sulfates can increase manganese solubility and thus increase aqueous mobility and uptake by plants (Reimer, 1999). Hart et al. (1992) studied the speciation of manganese in Magela Creek in tropical north Australia. They hypothesized that higher temperatures (30 °C) and increased rates of bacterially mediated oxidation could result in equilibrium between Mn(II) and oxidized species within the normal residence time of water in the creek. This was one mechanism by which colloidal manganese could dominate speciation.

There is evidence that afforestation of upland areas has increased manganese concentrations in surface waters. Analysis of sites in the United Kingdom between 1988 and 1996 shows a significant positive correlation between mean manganese concentrations and the percentage of conifer cover in the catchment (Heal, 2001). Enhanced manganese concentrations arise from foliar leaching and wash-off of manganese in fine mist and dry particles that are captured from the atmosphere by the trees (Shanley, 1986; Heal, 2001). Litter from conifer plantations may also enhance manganese leaching from soil into runoff. Soil and water acidification in catchments planted with conifers has been widely documented and is associated with enhanced manganese concentrations in surface waters (Heal, 2001). The extent to which land use influences manganese concentrations in upland catchments is modified by catchment hydrology and soil type (Heal, 2001; Heal et al., 2002). Heal et al. (2002) identified summer baseflow and the summer–autumn hydrological transition as critical periods for increased manganese concentrations in runoff. It is only when manganese enters lakes, estuaries, and the ocean, where residence times are considerably longer, that chemical processes will become dominant and the system will approach an equilibrium speciation (Laxen et al., 1984).

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 (<7.5‰) during estuarine mixing (L'Her Roux et al., 1998).

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 Trichoderma harzianum 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 K_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.

In the laboratory, microorganisms have been shown to transform both soluble and solid manganese; thus, they potentially have substantial effects on local manganese cycles. Physiological, biochemical, and structural studies of manganese oxidizers and reducers in the laboratory form the basis on which models of the participation of microorganisms in the cycling of manganese have been proposed. Field analyses of the distribution of manganese oxidizers and reducers, structural properties of manganese precipitates, and in situ activity measurements support the hypothesis that microorganisms play an integral role in the cycling of manganese in some environments (Nealson, 1983). Microbial oxidation of Mn(II) occurs at rates up to 5 orders of magnitude greater than those of abiotic Mn(II) oxidation (Tebo, 1991). Johnson et al. (1995) found that microbial catalysis was overwhelmingly responsible for manganese oxidation in the lower epilimnion of a freshwater dam during the summer months. Microbial oxidation of Mn(II) to Mn(IV) by spores of the marine Bacillus sp. was observed by Bargar et al. (2000), whereas Stein et al. (2001) found three freshwater bacterial isolates capable of manganese oxidation. Elevated manganese levels on the carapace of crayfish (Cherax destructor) are thought to be the result of manganese-oxidizing bacteria forming biofilms (King et al., 1999). Nealson et al. (1991) isolated and identified manganese-reducing bacteria in the Black Sea. The major group of organisms isolated from the 80- to 90-m (manganese reduction) zone were in the genus Shewanella. Microbially mediated reduction of complexed Mn(III) has also been observed in the laboratory (Kostka et al., 1995). Further studies have isolated manganese-reducing bacteria in marine sediments, oxic regions of lake water columns, and the rhizosphere of non-mycorrhizal plants (Posta et al., 1994; Bratina et al., 1998; Thamdrup et al., 2000).

Manganese is an essential element (see section 7.1) and is, therefore, actively assimilated and utilized by both plants and animals; however, it can be significantly bioconcentrated by aquatic biota at lower trophic levels. Bioconcentration factors (BCFs) of 2000–20 000 for marine and freshwater plants, 2500–6300 for phytoplankton, 300–5500 for marine macroalgae, 800–830 for intertidal mussels, and 35–930 for fish have been estimated (Folsom et al., 1963; Thompson et al., 1972; Bryan & Hummerstone, 1973; Pentreath, 1973; Rai & Chandra, 1992). Ichikawa (1961) reported that marine fish did not accumulate manganese to the same extent as organisms at lower trophic levels, with typical BCFs of about 100. Uptake of manganese by aquatic invertebrates and fish significantly increases with temperature (Miller et al., 1980) and decreases with pH (Rouleau et al., 1996), whereas dissolved oxygen has no significant effect (Miller et al., 1980; Baden et al., 1995). BCFs of 140 000–300 000 were reported for annelids (Lamellibrachia satsuma) living near hydrothermal vents (Kagoshima Bay, Japan) (Ando et al., 2002). Uptake of manganese has been found to increase with decreasing salinity (Struck et al., 1997). There are conflicting reports on whether biomagnification of manganese (i.e., increasing concentrations up the food-chain) occurs. Kwasnik et al. (1978) found that there was no biomagnification in a simple freshwater food-chain, with maximum BCFs of 911, 65, and 23 for algae, Daphnia magna, and fathead minnows (Pimephales promelas), respectively. In contrast, other authors have found weak biomagnification (Stokes et al., 1988).

Manganese in its reduced form, Mn(II), is bioavailable and can be readily taken up by benthic fauna. Lobsters living on fine cohesive mud sediments rich in manganese can accumulate manganese following autumnal hypoxia in the south-east Kattegat, Sweden (Eriksson, 2000a). Entry of dissolved manganese occurs mainly via gill transport (Baden et al., 1995), and when individuals are exposed to Mn(II) concentrations above 1.8 mg/litre, they accumulate manganese (Baden et al., 1995; Baden & Neil, 1998). Laboratory experiments with dissolved manganese have shown that the majority of the manganese taken up by lobsters is lost from the inner tissues after 5 days of excretion in "clean" seawater (Baden et al., 1995, 1999). Manganese concentrations in lobster eggs remained stable at around 5 mg/kg dry weight during oocyte maturation and throughout most of embryogenesis; however, concentrations started to increase at the end of embryonic development and had reached 120 mg/kg dry weight at the time of hatching (Eriksson, 2000b). Sea stars (Asterias rubens) accumulated dissolved ⁵⁴Mn linearly with time to a BCF of 19 after 23 days. Manganese accumulated from seawater was eliminated according to first-order kinetics, with a half-life of 36 days. Sea stars fed a diet containing ⁵⁴Mn assimilated 69–83%. Elimination of manganese accumulated from food was described as a two-compartment system, with half-lives of 1.8 and 25 days (Hansen & Bjerregaard, 1995). Bioaccumulation studies of wastewater from a thermo-mechanical paper mill using the freshwater crayfish Cherax destructor consistently demonstrated elevated levels of manganese in the crayfish. However, the authors suggest that the elevated levels observed were due to manganese-oxidizing bacteria forming biofilms on the carapace followed by manganese dioxide precipitation rather than active uptake by the crayfish (King et al., 1999).

Manganese was readily accumulated in all tissues of brown trout (Salmo trutta) at concentrations reflecting the lowest natural concentrations found in circumneutral lakes. Trout exposed to 0.1 µg ⁵⁴Mn/litre accumulated 1.78 mg manganese/kg (whole body) within 6 weeks, representing 95% of the steady-state concentration. The depuration rate was initially rapid, with 22% loss of radiolabelled manganese after 1 week (Rouleau et al., 1995). The addition of humic and fulvic acids had little effect on the uptake of manganese; however, other chelating agents, such as potassium ethylxanthate, sodium diethyldithiophosphate, and sodium dimethyl- and diethyldithiocarbamate, decreased bioaccumulation by 40% (Rouleau et al., 1992).

Terrestrial plant species vary a great deal in their ability to accumulate manganese. The absolute concentration of manganese in soils is generally less important to plants than the availability of manganese, which is determined by pH, cation exchange capacity, concentration of other cations, organic content, temperature, and microbial activity. Plants take up manganese from soil primarily in the divalent state. Differences in plant uptake can be explained in part by differences in the ability of plants to bring about the dissolution of oxidized manganese (Stokes et al., 1988). The application of chelating agents significantly reduced the uptake of manganese in roots, stems, and leaves of okra (Abelmoschus esculentus) at manganese concentrations of 500 and 1000 mg/kg (Denduluri, 1994).

A BCF of 2 was calculated for earthworms (Lumbricus terrestris) exposed to ⁵⁴Mn in litter for 20 days, with a depuration half-time of 40 days (Sheppard et al., 1997).

According to a National Research Council of Canada report (Stokes et al., 1988), manganese concentrations in air tend to be lowest in remote locations (about 0.5–14 ng/m³ on average), higher in rural areas (40 ng/m³ on average), and still higher in urban areas (about 65–166 ng/m³ on average) (see Table 2). Similar concentrations have been reported elsewhere, leading to the conclusion that annual manganese concentrations average 10–30 ng/m³ in areas far from known sources and 10–70 ng/m³ in urban and rural areas without major point sources of manganese (WHO, 1999). Manganese concentrations in air tend to be highest in source-dominated areas (e.g., those with foundries), where values can reach 8000 ng/m³ (US EPA, 1984; Stokes et al., 1988). Annual averages of manganese concentrations may rise to 200–300 ng/m³ in air near foundries and to over 500 ng/m³ in air near ferro- and silicomanganese industries (WHO, 1999). Manganese concentrations in air have been measured in many specific locations. In the Vancouver, Canada, area, for example, annual geometric mean concentrations of manganese ranged from <10 to 30 ng/m³ in 1984 (Stokes et al., 1988). Over the period 1981–1992, Loranger & Zayed (1994) found average manganese concentrations in Montreal, Canada, of 20 and 60 ng/m³ in areas of low and high traffic density, respectively. More recently, Loranger & Zayed (1997) found the average concentration of total manganese in an urban site in Montreal to be 27 ng/m³. In selected periods in the 1970s, annual mean concentrations of manganese were reported to range from 3 to 16 ng/m³ in two German cities (Frankfurt and Munich), from 42 to 455 ng/m³ in Belgium, and from 20 to 800 ng/m³ in Japanese cities (WHO, 1999). More recent analysis of urban aerosols revealed mean manganese concentrations ranging from 80 to 350 ng/m³ for the city of Kayseri, Turkey (Kartal et al., 1993), 154 ng/m³ for Bhilai, India (1995–1996) (Pandey et al., 1998), and 16.5 ng/m³ for Seville, Spain (Espinosa et al., 2001).

As Table 2 shows, manganese concentrations in air in the USA have decreased over the past three decades (Kleinman et al., 1980; US EPA, 1984), a trend believed to be due primarily to the installation of industrial emission controls (US EPA, 1984, 1985a). In Ontario, Canada, as well, annual average manganese concentrations in air have decreased, along with total suspended particulate levels (Stokes et al., 1988).

In a review of worldwide data on trace metals in precipitation, median concentrations of manganese in wet deposition were 23, 5.7, and 0.19 µg/litre for urban, rural, and remote locations, respectively (Galloway et al., 1982).

Concentrations of manganese in open seawater range from 0.4 to 10 µg/litre (US EPA, 1984; Zeri et al., 2000). In the North Sea, the north-east Atlantic Ocean, the English Channel, and the Indian Ocean, manganese content was reported to range from 0.03 to 4.0 µg/litre. Levels found in coastal waters of the Irish Sea and in the North Sea off the coast of the United Kingdom ranged from 0.2 to 25.5 µg/litre (Alessio & Lucchini, 1996). Higher concentrations (up to 500 µg/litre) have been reported for anaerobic layers of open seawater (Lewis & Landing, 1991, 1992). Hypoxic concentrations below 16% saturation can increase the concentration of dissolved manganese above that normally found in seawater to concentrations approaching 1500 µg/litre (Balzer, 1982).

Concentrations of dissolved manganese in natural waters that are essentially free of anthropogenic sources can range from 10 to >10 000 µg/litre (Reimer, 1999). However, manganese concentrations in natural surface waters rarely exceed 1000 µg/litre and are usually less than 200 µg/litre (Reimer, 1999). In a 1974–1981 survey of 286 US river water samples, concentrations of dissolved manganese ranged from less than 11 µg/litre (25th percentile) to more than 51 µg/litre (75th percentile) (Smith et al., 1987), with a median of 24 µg/litre. Mean groundwater concentrations were 20 and 90 µg/litre from two geological zones in California, USA (Deverel & Millard, 1988). In a number of cases, higher levels in water (in excess of 1000 µg/litre) have been detected at US hazardous waste sites, suggesting that, in some instances, wastes from industrial sources can lead to significant contamination of water (ATSDR, 2000). Concentrations of dissolved manganese of up to 4400 µg/litre have been recorded in waters receiving acid mine drainage (pH <2.5) (Filipek et al., 1987). The surface waters of Welsh rivers were reported to contain from 0.8 to 28 µg manganese/litre (Alessio & Lucchini, 1996). Neal et al. (1986) reported mean manganese concentrations of 41 and 30 µg/litre for two Welsh streams with mean rainfall concentrations measured at 2 µg manganese/litre. They found that stormflow waters were acidic (pH ~4.5) and enriched in soluble manganese. Concentrations of manganese ranged from 1 to 530 µg/litre in 37 rivers in the United Kingdom and in the Rhine and the Maas and their tributaries (Alessio & Lucchini, 1996). Mean dissolved manganese concentrations ranging from 6 to 117 µg/litre were found for six United Kingdom rivers with particulate manganese concentrations of 7–93 µg/litre (Neal et al., 2000). Higher manganese concentrations were associated with increasing suspended sediment levels caused by higher river flows. In Chesapeake Bay, USA, manganese concentrations of up to 237 µg/litre have been recorded during anoxic conditions (Eaton, 1979), and Kremling (1983) found manganese concentrations of around 700–800 µg/litre in anoxic bottom waters in the Baltic Sea. During the late summer, Horsetooth Reservoir, Colorado, USA, is fully stratified and exhibits seasonally high fluxes of iron, manganese, and metal-rich particles into the water column. Stein et al. (2002) monitored manganese concentrations during August 1999. The total manganese concentration in water prior to filtration and measured by atomic absorption spectrometry was 93 µg/litre; after sedimentation of particles, the total manganese concentration measured by ICP with atomic emission spectrophotometry was 213 µg/litre.

Manganese has been measured in snow core samples dated from 1967 to 1989 collected in central Greenland at concentrations ranging from 0.016 to 0.236 µg/kg. A large fraction of the manganese was found to originate from rock and soil dust; "excess" manganese sources suggested include volcanoes, natural vegetation fires, continental biogenic emissions, and anthropogenic sources such as industrial outputs and MMT (Veysseyre et al., 1998).

Manganese concentrations in river sediments from the South Platte River Basin, USA (1992–1993), ranged from 410 to 6700 mg/kg dry weight, with a geometric mean of 1260 mg/kg dry weight (Heiny & Tate, 1997). Total manganese levels in sediment of a Chesapeake Bay, USA, tributary ranged from 940 to 2400 mg/kg dry weight (Hartwell et al., 2000). Sediment from an urban lake (Germiston Lake, South Africa) receiving inputs from industrial and residential areas, as well as windborne dust from old mine dumps, contained manganese concentrations ranging from 970 to 13 400 mg/kg dry weight (Sanders et al., 1998). Mangrove sediments in the Arabian Gulf contained manganese concentrations ranging from 29 to 170 mg/kg dry weight, with higher concentrations associated with the local geology (Shriadah, 1999). Sediment manganese concentrations of 100–1000 mg/kg dry weight have been reported for an intertidal flat off Korea (Jung et al., 1996). Similar total manganese values were found in the northern Adriatic Sea (Italy), ranging from 200 to 800 mg/kg dry weight, with a mean of 370 mg/kg dry weight (Fabbri et al., 2001). Surface sediments in the Baltic Sea contained mean manganese concentrations of 3550 (Bothnian Sea), 5070 (Gulf of Finland), and 8960 (Bothnian Bay) mg/kg dry weight; the high manganese concentrations were thought to be due to ferromanganese concretions and riverine loads (Leivuori, 1998). Sediment pore water may contain dissolved manganese at concentrations of 0.2–24 mg/litre (Bryan & Hummerstone, 1973; Aller, 1994; Eriksson & Baden, 1998), whereas bottom water concentrations are normally around 0.2–17 µg/litre (Hall et al., 1996).

Natural ("background") levels of total manganese in soil range from <1 to 4000 mg/kg dry weight, with mean values around 300–600 mg/kg dry weight (Shacklette et al., 1971; Cooper, 1984; US EPA, 1985b; Adriano, 1986; Schroeder et al., 1987; Eckel & Langley, 1988; Rope et al., 1988). Reimer (1999) reported mean total manganese concentrations from uncontaminated soils for British Columbia, Canada, ranging from 284 to 1359 mg/kg dry weight. Similar total manganese concentrations were found in volcanic soils in Romania at varying distances from a lead smelter (Donisa et al., 2000). Median manganese concentrations of up to 1980 mg/kg dry weight were reported for urban areas of Honolulu, Hawaii, USA (Sutherland & Tolosa, 2001). Accumulation of manganese in soil usually occurs in the subsoil and not on the soil surface (IPCS, 1981). Manganese was found at higher concentrations in the upper layers of soil adjacent to highways in Canada following 25 years of the use of MMT as a fuel additive. However, this use has not led to a significant increase in either total or exchangeable manganese in these soils. There was a trend of decreasing manganese concentrations with distance from the highway, but this was not statistically significant (Bhuie et al., 2000; Bhuie & Roy, 2001). In general, soil manganese concentrations in the Kola Peninsula, Russian Federation, ranged from 100 to 500 mg/kg dry weight, except for an extremely contaminated (copper/ nickel smelter complex) and eroded podzol containing 3500 mg/kg (Barcan & Kovnatsky, 1998). Higher concentrations have been reported for some acid soils; for example, a major portion of agricultural land on Oahu, Hawaii, USA, consists of oxisols of basaltic origin, with total manganese concentrations ranging from 10 000 to 40 000 mg/kg (Fujimoto & Sherman, 1948).

Geometric mean manganese concentrations in seaweed from south-west England ranged from 128 to 392 mg/kg dry weight (Bryan & Hummerstone, 1973). Manganese concentrations in seaweed (Fucus vesiculosus) and mussels (Mytilus edulis) were 350 and 29 mg/kg dry weight for the North Sea for the two species, respectively, and 735 and 46 mg/kg dry weight for the Baltic Sea (Struck et al., 1997). High manganese concentrations have been found in all tissues of blue crabs (Callinectes sapidus) from a metal-contaminated estuary in North Carolina, USA (Weinstein et al., 1992), and in Norway lobsters (Nephrops norvegicus) following autumnal hypoxia in the south-east Kattegat, Sweden. Lobsters living on sediments rich in manganese contained whole-body mean manganese concentrations of 92 mg/kg dry weight (Eriksson, 2000a). Similarly, river crabs (Potamonautes warreni) in an urban lake with manganese-rich sediment contained a mean manganese concentration of 662 mg/kg dry weight (Sanders et al., 1998).

Mean manganese levels in mussels (Mytilus trossulus and Crenomytilus grayanus) from the north-west Pacific Ocean ranged from 2.8 to 9.3 mg/kg dry weight (Kavun et al., 2002). Manganese concentrations in barnacles (Balanus amphitrite and Tetraclita squamosa) from Xiamen Harbour and Hong Kong coastal waters ranged from 5.9 to 277 mg/kg dry weight (Blackmore et al., 1998; Blackmore, 1999; Rainbow & Blackmore, 2001). During the 1990s, manganese concentrations in barnacles had increased at some sampling sites, probably due to the resuspension of metal-rich sediments from dredging and reclamation associated with major construction projects (Blackmore, 1999). A manganese concentration of 6 mg/kg for the muscle tissue of annelids (Lamellibrachia satsuma) living near hydrothermal vents was reported by Ando et al. (2002).

Concentrations of manganese found in tissues of marine and freshwater fish tend to range from <0.2 to 19 mg/kg dry weight (Greichus et al., 1977, 1978; Capelli et al., 1987; Sindayigaya et al., 1994; Heiny & Tate, 1997). Higher manganese concentrations of up to 40 mg/kg wet weight (equivalent to dry weight concentrations of >100 mg/kg) were reported for fish from the Lower Savannah River, USA, probably related to acid mine drainage (Winger et al., 1990), and mean liver concentrations of 54 mg/kg dry weight were reported for a polluted lake (Saad et al., 1981). Manganese levels in the opercula and scales of brook trout (Salvelinus fontinalis) were 1.6 times higher in fish from acidified lakes (pH 5.2–5.5) than in fish from non-acidified lakes (pH 6.8–7.0) (Moreau et al., 1983). Bendell-Young & Harvey (1986) found significantly more manganese in all tissues of white suckers (Catostomus commersoni) from an acidified lake (pH 4.8) compared with other lakes (pH 5.0–6.0) in south-central Ontario, Canada.

Concentrations of manganese in terrestrial plants tend to range from 20 to 500 mg/kg (Stokes et al., 1988). Rautio et al. (1998) analysed scots pine (Pinus sylvestris) needles at various distances from a smelter complex (Finnish Lapland and the Kola Peninsula, Russian Federation). Manganese concentrations ranged from <50 to >1200 mg/kg dry weight, with the lowest concentrations in the vicinity of the smelter, possibly because of increased foliar leakage and/or leaching of cations from the upper soil layers due to sulfur and heavy metal deposition. Mean levels of manganese in soil and vegetation near a municipal incinerator in Montcada, Spain, were 390–420 mg/kg dry weight and 49–54 mg/kg dry weight, respectively (Meneses et al., 1999), whereas seven species of mushroom in the vicinity of a nickel/copper smelter contained 11–67 mg/kg dry weight (Barcan et al., 1998). Members of the Ericaceae family, which includes blueberries, are regarded as manganese accumulators. There are numerous reports of foliar manganese levels in excess of 2000–4000 mg/kg, particularly for Vaccinium angustifolium and V. vitis-idaea (Korcak, 1988). Creeping snowberry (Gaultheria hispidula) and velvet-leafed blueberry (V. myrtilloidies) from a bog soil (pH 4.0) contained 3000 and 2200 mg manganese/kg, respectively (NAS, 1973).

Menta & Parisi (2001) found mean manganese concentrations of 200–300 mg/kg dry weight in the digestive gland of terrestrial snails (Helix pomatia and H. aspersa) collected from a semirural area in northern Italy and 10–20 mg/kg in the foot; however, manganese concentrations for the slug Arion rufus were 92 mg/kg for the digestive gland and 394 mg/kg for the foot. Similar concentrations were found in the slug Arion ater at a site with vegetation levels of 150 mg manganese/kg; however, at a contaminated site near a disused lead/zinc mine (with vegetation levels of 207 mg/kg), manganese concentrations in the digestive gland and foot were 210 and 26 000 mg/kg dry weight, respectively (Ireland, 1979). The authors hypothesized that the lower levels of manganese found in the foot of the snail compared with the slug were related to the quantity of calcium required for shell formation. Earthworms (Lumbricus rubellus) from the Rhine Delta floodplain contained mean manganese concentrations of 100–120 mg/kg dry weight (Hendricks et al., 1995).

Mean manganese concentrations in birds' eggs from a variety of geographical areas range from 1 to 5 mg/kg dry weight (Hothem et al., 1995; Hui et al., 1998; Burger et al., 1999); mean liver concentrations range from 3 to 11 mg/kg dry weight (Hui et al., 1998; Burger & Gochfeld, 1999, 2000b). Burger & Gochfeld (2000a) analysed the feathers of 12 species of seabird from the northern Pacific Ocean and found mean manganese concentrations ranging from 0.3 to 2 mg/kg dry weight. Higher manganese concentrations (4.5 mg/kg) were found in Laysan albatross (Diomedea immutabilis) feathers from the same location (Burger & Gochfeld, 2000b). Little egret (Egretta garzetta) and black-crowned night-heron (Nycticorax nycticorax) feathers contained 1.7–22.6 mg manganese/kg dry weight (Connell et al., 2002). Mean levels of manganese were lower than previously reported for egrets and herons in the early 1990s (4.2–63.1 mg/kg) (Burger & Gochfeld, 1993). Great tit (Parus major) feathers had mean concentrations ranging from 17.4 to 43.8 mg/kg dry weight (Janssens et al., 2001).

American alligators (Alligator mississippiensis) from lakes in central Florida, USA, contained a mean manganese concentration of 1.2 mg/kg wet weight in liver tissue; highest concentrations (3.1 mg/kg) were found in the regenerated tail (Burger et al., 2000). Slider turtle (Trachemys scripta) eggs contained 4.5 mg manganese/kg dry weight (Burger & Gibbons, 1998), whereas pine snakes (Pituophis melanoleucus) contained whole-body mean concentrations ranging from 11.9 to 17 mg manganese/kg dry weight (Burger, 1992).

Mean manganese concentrations in Baikal seal (Phoca sibirica) muscle, liver, and kidney tissue were 0.12, 2.1, and 0.84 mg/kg wet weight, respectively; manganese concentrations decreased with age (Watanabe et al., 1998). Mean manganese concentrations in European otter (Lutra lutra) livers ranged from 3.5 to 7.4 mg/kg dry weight (Mason & Stephenson, 2001). Similar mean concentrations were found in shrew (Crocidura russula and Sorex araneus) kidneys, ranging from 7.3 to 8.8 mg manganese/kg dry weight (Hendricks et al., 1995).

Manganese is an essential nutrient for microorganisms, plants, and animals (Underwood, 1977; Woolhouse, 1983). Manganese must be provided as a micronutrient in the culture media for growing algae (McLachlan, 1973) and in the diet of captive fish (Satoh et al., 1987), birds, and mammals (Underwood, 1977). Many calcareous soils require the application of manganese fertilizers for optimum crop growth (Woolhouse, 1983).

Manganese deficiency has been shown to limit the growth rate of marine phytoplankton, particularly where manganese-depleted deep seawater is upwelled to the surface (Huntsman & Sunda, 1980). Manganese levels can affect the microflora species composition of streams. In experiments, it has been shown that under low manganese concentrations (<0.04 mg/litre), blue-green and green algal flora can predominate, whereas at higher manganese levels, diatoms dominate (Patrick et al., 1969). Neutral streams with elevated levels of iron and manganese can develop blooms of ferromanganese-depositing bacteria with oxide deposition zones. Algal abundance declines within these blooms (Wellnitz & Sheldon, 1995).

Overall assemblage composition of macroinvertebrates in upland streams (Wales and Cornwall, United Kingdom) correlated with manganese concentration, pH, and nitrate concentration, with assemblage scores (relative abundance, species richness, and diversity) increasing with increasing manganese concentrations (0.003–0.6 mg/litre) (Hirst et al., 2002).

Rainbow trout (Oncorhynchus mykiss) fed on a manganese-deficient diet (4.4 mg manganese/kg diet) for 60 weeks developed lens cataracts and short body dwarfism, although growth was not affected (Yamamoto et al., 1983). Knox et al. (1981) found no effect of low dietary manganese (1.3 mg/kg diet) on trout growth during a 24-week feeding period; there were effects on plasma ion levels, hepatic mineral levels, and hepatic enzyme activity. Experiments have shown that the addition of 10 mg manganese/kg (as manganese sulfate or manganese chloride) to white fishmeal diets (containing 2–3 mg manganese/kg) is necessary to obtain normal growth of captive carp (Cyprinus carpio) (Satoh et al., 1987). The dietary uptake of manganese dioxide and manganese carbonate by carp was found to be low.

Manganese is an essential nutrient for plant growth as a constituent of a number of metalloenzymes that occupy key roles in metabolism (Clarkson & Hanson, 1980; Woolhouse, 1983; Burnell, 1988). Nutritional requirements for terrestrial plants are around 10–50 mg manganese/kg tissue (Hannam & Ohki, 1988; Reisenauer, 1988). Critical nutritional levels vary widely between species and among cultivars of a species (Reisenauer, 1988). Calcareous soils, especially those with poor drainage and high organic matter, are the types of soil that produce manganese-deficient plants. There are numerous examples of manganese deficiency, especially among crop plants. Manganese deficiency of peanut (Arachis hypogaea) is a common problem on some soils of the coastal plain region of the southern USA. Manganese deficiency occurred at pH levels of 6.8; maintaining soil pH at 6 provided a desirable medium for plant growth without the need for manganese fertilizer. Critical manganese levels ranged between 12 and 15 mg/kg in leaves (Parker & Walker, 1986). Application of manganese sulfate (15 mg manganese/kg) to highly calcareous soils enhanced the growth of soybean (Glycine max) plants (Ahangar et al., 1995). Critical limits for a variety of plant species have been calculated: for example, for corn (Zea mays), 10.6 mg manganese/kg in the ear leaf and 4.9 mg manganese/kg in the grain (Uribe et al., 1988); for oats (Avena sativa), 4.5 mg diethylenetriaminepentaacetic acid (DTPA)-extractable manganese/kg soil and 19 mg manganese/kg dry matter for mature leaf blades; and for cowpea (Vigna unguiculata), 2.4 mg DTPA-extractable manganese/kg soil and 41 mg manganese/kg dry matter for leaf blades (Bansal & Nayyar, 1996, 1998).

The importance of manganese for birds has been recognized for almost 50 years, ever since Wilgus et al. (1936) demonstrated that manganese could prevent perosis (a disease causing bone deformities) in the chicken. Subsequent research has shown that manganese is vital for growth, egg production, and proper development of the chick embryo and is essential in the activation of numerous enzymes (Underwood, 1977). Experiments have shown that the minimal manganese requirement for chicks on a casein-dextrose diet was 14 mg/kg diet (Halpin & Baker, 1986).

Plants and animals require manganese as an essential nutrient up to a certain level often referred to as the deficiency limit; however, at concentrations higher than this level, toxic effects are often observed.

Most toxicity tests have been carried out using ionic manganese. Little is known about the aquatic toxicity of colloidal, particulate, and complexed manganese; in general, however, toxicities of metals bound into these forms are assumed to be less than those of the aquo-ionic forms. Manganese fungicides have been referred to in this CICAD for source and fate information only, and no attempt has been made to evaluate this group of chemicals for environmental effect. Toxicity tests for the effects of manganese on aquatic biota are summarized in Table 3. For algae, there is a wide range of toxicity values; the most sensitive species appear to be the marine diatom Ditylum brightwellii, with a 5-day EC₅₀, based on growth, of 1.5 mg manganese/litre, and a freshwater alga Scenedesmus quadricauda, with a 12-day EC₅₀, based on chlorophyll inhibition, of 1.9 mg manganese/litre.

Table 3: Toxicity of manganese to aquatic species.^a