Anthropogenic emissions of mercury (Hg) to the atmosphere have increased the actively cycled pool of mercury by threefold to fivefold, and Hg contamination of aquatic ecosystems in the United States and across the globe is widespread [Lindberg et al., 2007; Munthe et al., 2007]. Methylmercury (MeHg) is of particular concern because this form bioaccumulates and biomagnifies in aquatic and terrestrial food webs, and as a result, Hg fish consumption advisories have been listed for surface waters in all 50 U.S. states [U.S. Environmental Protection Agency (EPA), 2009]. The source of Hg to most catchments is believed to be atmospheric deposition that originates from human and natural emissions of which coal burning is the leading source [Pacyna et al., 2006]. In northeastern North America, atmospheric Hg deposition appears to be decreasing in recent years [Butler et al., 2008], but Hg in lake sediment cores has not yet shown widespread and universal declines [Muir et al., 2009]. Annual loads of Hg exported from catchments, however, are generally only a small fraction of annual deposition, and even when gaseous losses are estimated, many sites continue to be net accumulators of Hg [Shanley et al., 2008; Brigham et al., 2009].
 Soils are the largest store of Hg in most catchments and typically the mineral soil has a greater store of Hg than the forest floor despite higher Hg concentrations per unit of soil mass in the organic horizons [Grigal, 2003]. Even higher Hg concentrations are observed in wetlands and in peaty soils, and this Hg is believed to be largely associated with soil organic carbon, suggesting that organic carbon distribution in a catchment landscape controls Hg distribution [Grigal, 2003]. Concentrations of THg and MeHg vary greatly among surface waters, and this spatial variation typically exceeds the spatial variations in atmospheric Hg deposition [Wiener et al., 2006]. Wetlands are believed to be one of the key landscape features that affect Hg storage and transport, and percent wetland area is often strongly related to surface water Hg concentrations [Brigham et al., 2009]. Despite the importance of wetlands to Hg transport, wetland area is sometimes unrelated to Hg concentrations in adjacent waters, which highlights the pivotal role that the hydrologic connection of wetlands with local waters can play in the catchment Hg cycle [Selvendiran et al., 2008]. Other landscape factors are known to affect the Hg cycle in catchments such as the presence of forests, the relative amount of coniferous forest, and the hydrologic connection of upland and wetland or riparian areas [Kolka et al., 2001; Choi et al., 2008; Mitchell et al., 2008b].
 Wetlands generally act as net sinks of inorganic Hg [Kolka et al., 2001], and these losses are assumed to be largely through sorption processes. Although Hg in the elemental form (Hg0) can readily exchange as a gas into and out of soils [Choi and Holsen, 2009], most wetlands are likely accumulating ionic Hg, particularly those with peaty soils [Klaminder et al., 2008]. Some of the Hg pool can become methylated by bacteria under anaerobic conditions; this process is generally believed to be carried out largely by sulfate reducing bacteria (SRB) [Benoit et al., 2003; Drott et al., 2007] for which the availability of sulfate and labile DOC are often controlling factors [Mitchell et al., 2008a]. Additional losses of Hg within catchments can occur by photoreduction of Hg and volatilization of the resulting Hg0, directly from surface waters, and MeHg can be removed through biological and physical demethylation processes in streams and ponds [Krabbenhoft et al., 2005].
 The transport of Hg between upland areas and riparian areas is often favored by the flux of water between these two landscape features [Selvendiran et al., 2009]. Rapid runoff from uplands can speed the transport of Hg to wetlands, where methylation may occur [Kolka et al., 2001]. The loads of Hg and DOC that reach the outlet of a catchment may be less dependent on the amount of Hg stored in the soils, than on the efficiency of physical and chemical mobilization processes [Demers et al., 2010].
 Numerous studies show strong positive relationships between total mercury and DOC or particulate organic carbon (POC) concentrations in stream water [Driscoll et al., 1995; Schuster et al., 2008; Selvendiran et al., 2008; Dittman et al., 2010], which reflects the dependence of Hg(II) on binding to soluble organic matter to facilitate transport [Haitzer et al., 2002; Dittman et al., 2010]. Furthermore, the hydrophobic organic acid fraction (HPOA) of DOC has been found to explain a higher fraction of the variability of Hg concentrations in surface waters than can be explained by DOC concentrations [Shanley et al., 2008; Dittman et al., 2009].
 The mobilization of Hg from soils is driven in large part by changing hydrological conditions, especially high-flow events such as the observed direct responses of THg concentrations in stream water to rain storms and snowmelt events [Bishop et al., 1995; Shanley et al., 2002; Bushey et al., 2008; Schuster et al., 2008; Demers et al., 2010; Dittman et al., 2010].
 New conceptualizations of the flushing of mercury and DOC caused by changing flowpaths in soils as well as changing source areas [Demers et al., 2010; Dittman et al., 2010], often represented by the strong coupling of DOC and THg concentrations [Dittman et al., 2009], suggest the possibility of developing joint conceptualizations for both of these constituents. Numerous model applications have been applied to investigate the process of hydrological DOC mobilization to surface waters through chemical mixing models [Brown et al., 1999; Christophersen and Hooper, 1992; van Verseveld et al., 2008] and rainfall-runoff models [Hornberger et al., 1994; Boyer et al., 1997; Weiler and McDonnell, 2006], often referred to as a “flushing mechanism” [Burns, 2005]. Whereas initial attempts to conceptualize hydrologic processes of Hg mobilization through the application of simple mixing models in small watersheds have shown some potential [e.g., Demers et al., 2010], little has been done to conceptualize landscape-scale hydrologic processes of Hg mobilization by the application of spatially explicit hydrologic models in larger and more complex watersheds.
 Recent research in the Adirondack Mountains of New York indicates that the connection of small riparian wetlands to major hydrologic flow paths provide important controls on THg, MeHg and DOC in streams [Inamdar and Mitchell, 2006; Selvendiran et al., 2008]. We assume that similar processes would also operate at a catchment scale that is greater than that of the above mentioned studies, and hypothesize that the riparian area in small upland tributary streams, and the wetland/upland interface in higher-order stream channels represent major landscape controls on DOC and Hg species. Further we hypothesize that high-flow conditions will increase the connectivity of upland areas to riparian wetlands that dominate in lower valley bottoms and mobilize Hg and DOC from surface soil layers. Therefore, the aim of this study was to investigate the role of hydrological controls on the mobilization of THg, MeHg and DOC during high-flow conditions in the Fishing Brook (FB) Catchment, Adirondack Mountains, New York State. A combined field and modeling approach was used, based on saturation state simulations with the hydrological model TOPMODEL [Beven and Kirkby, 1979; Ambroise et al., 1996; Ibbitt et al., 2009] to identify first-order controls on the mobilization of Hg and DOC. We define the following terms for use in this study: (1) “hydrological flushing” as the process of transport of Hg or DOC from the terrestrial to the aquatic system and (2) “net mobilization” as the combination of a flushing process and additional processes within the stream network which may alter Hg and DOC concentrations prior to sampling at the catchment outlet.