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Like many other elements, natural background levels of trace elements exist in crustal rocks, such as shales, sandstones, and metamorphic and igneous rocks (Benjamin and Honeyman, 2000). In particular, the majority of trace metals are derived from igneous rocks, simply based on the relative fraction of igneous rocks in comparison with sedimentary and metamorphic rocks in the Earth’s crust. The release of trace metals from crustal sources is largely controlled by the natural forces of physical and chemical weathering of rocks, notwithstanding large-scale anthropogenic disturbances such as mining, construction, and coal burning (release of fly ash). As discussed later in the chapter, adjustments can be made for anthropogenic loading to different ecosystems based on an enrichment factor which compares metal concentrations in the ecosphere to average crustal composition. Biological effects of weathering, such as plant root growth and organic acid release associated with respiration also contribute to these weathering processes. As some trace metals are more volatile than others, release due to volcanic activity represents another source of metals with such properties (e.g., Pb, Cd, As, and Hg). Just as Goldschmidt (1954) grouped elements (e.g., siderophiles, chalcophiles, lithophiles, andatomophiles) based on similarities in geochemical properties, trace metals also represent a group of elements with similar chemical properties. One particularly important distinguishing feature of these elements is their ability to bond reversibly to a broad spectrum of compounds (Benjamin and Honeyman, 2000). Thus, the major inputs of trace metals to estuaries are derived from riverine, atmospheric, and anthropogenic sources. Although trace elements typically occur at concentrations of less than 1 ppb (part per billion) (or μg L−1, also reported in molar units), these elements are important in estuaries because of their toxic effects, as well as their importance as micronutrients for many organisms. The fate and transport of trace elements in estuaries are controlled by a variety of factors ranging from redox, ionic strength, abundance of adsorbing surfaces, and pH, just to name a few (Wen et al., 1999).

Human demands on aquatic and terrestrial ecosystems are on the increase globally and have likely exceeded the regenerative capacity of the Earth since the 1980s. Demands on our aquatic resources will increase in coming decades as it is projected that 75% of the world’s population (6.3 billion) will reside in coastal areas by 2025 (Tilman et al., 2001). The Earth’s population is expected to reach 9 billion during this century, and the projected effects of contaminant loading and human encroachment on biodiversity still remain unclear. The disturbance on global coastal ecosystems and the threat it will have on the economically critical resources they provide, has been estimated to be valued at 12.6 trillion U.S. dollars (Costanza et al., 2001). It has become increasingly apparent that in many regions of the world, Earth systems, which have been viewed as being primarily controlled by natural drivers such as climate, vegetation, and lithology, are now controlled by social, societal, and economic drivers (e.g., population growth, urbanization, industrialization water engineering) (Meybeck, 2002, 2003). This replacement of natural drivers over the past 50 to 200 years has recently been referred to as the Anthropocene era (first postulated by Vernadski, 1926), as a next phase that follows the Holocene era (Crutzen and Stoermer, 2000). Other studies that have effectively made large-scale linkages between human effects on the Earth systems (Turner et al., 1990) and aquatic systems (Costanza et al., 1990, 1997; Meybeck, 2002, 2003; Meybeck and Vörösmarty, 2004) have all concluded that a more comprehensive and fine-scale interpretation of the Anthropocene is needed if we are to make future predictions and management decisions effectively. The growth and movement of human populations have resulted in a significant stressor in the form of invasive species that has altered global biodiversity patterns. For example, the introduction of invasive species worldwide has changed the community composition and physical structure of many ecosystems (Elton, 1958; Vitousek et al., 1997). Estuarine systems, like the northern San Francisco Bay, have experienced serious declines in productivity at the base of the food web over recent decades after the introduction of the Asian clam, Potamocorbula amurensis, in 1987 (Carlton et al., 1990).