Fate and Transport of Metals and Particulates within the Roadside Environment – A Review - page 4

 

Infiltration of stormwater

Typically, overland flow only occurs when run–on rates are greater than infiltration rates. Infiltration is dependent on soil type and aqueous electrical conductivity (Borselli, et al., 2001), vegetated cover (Greene et al., 1994), litter cover (Agassi et al., 1998; Costantini and Lcoh, 2002; Hart and Frasier, 2003), surficial crusting (Lasanta et al., 2000), water quality (Kim and Miller, 1996; Borselli et al., 2001), and most importantly, the presence and abundance of macro-pores (Haria et al., 1998; Heppell et al., 2000; Weiler and Naef, 2003; Di Pietro et al., 2003). Macro-pores can be responsible for up to 86% of infiltration (Heppell et al., 2000).

Infiltration also plays a part in metals transport. Particulate bound metals delivered via overland flow are essentially strained by the soil matrix during infiltration. Aqueous phase metals are adsorbed onto and retained by the soil matrix during infiltration (Dierkes and Geiger, 1999; Turer et al., 2001) or adsorb onto vegetation, most preferentially senescent vegetation (Ratcliffe and Beeby, 1980).

Erosion of metals laden particulates

Once metals laden particulates are deposited, they can be subjected to both overland flow and, when vegetation is minimal, raindrop erosive forces, both of which will transport particulates further down-slope. Erosive capacity of overland flow is a function of velocity and existing sediment load (Zheng et al., 2000). Overland flow carrying a maximum sediment load has little erosive ability, while sediment free overland flow of the same velocity can cause considerably more erosion. Varying only sediment load in introduced overland flow will result in constant down-slope sediment delivery. Bare soils are also subject to particulate dislodgement and scouring by the kinetic energy of rain drops (Meyer and Harmon, 1992; Thompson et al., 2001). This process increases particulate availability to the transport mechanisms of overland flow.

Metals transport within the soil matrix

Once retained by the soil matrix, either through deposition via overland flow of particulate bound metals or adsorption of aqueous phase metals following infiltration, metals are susceptible to various transport or leaching mechanisms. Transport of metals through soils is enormously complex. Much research has been done regarding metals transport through soil columns (Igloria et al., 1996b; Papini and Majone, 1997; Delmas et al., 2002; Marcos et al., 2002). Transport of some metals is a function of pH (Sauve et al., 2000), soil organic matter (Igloria et al., 1996a), dissolved soil organic matter (Jordan et al., 1997; Liu and Gonzalez, 2000), solute colloid concentration (Amrhein et al., 1993), salinity from deicing salts (Amrhein et al., 1992), temperature (Vandenabeele and Wood, 1972), and reducing conditions caused by inundation (Charlatchka and Cambier, 2000; Ma and Dong, 2004). Adsorption/desorption kinetic processes are very dynamic, and while metal species are largely immobile once they adsorb onto the soil matrix, they are susceptible to some mobilization under certain conditions as described above. Metal species in roadside soils deposited over the past century have not had sufficient time to migrate far, as there is little evidence in the literature of metals migrating beyond depths of 1 m. Under normal circumstances, barring the use of deicing salts, the potential for transportation generated stormwater runoff to contaminate groundwater through infiltration is limited (Mikkelsen et al., 1997; Barraud et al., 1999).

Influence of winter conditions

Winter conditions invoke a number of factors that affect micro scale transport mechanisms. Frozen ground can inhibit infiltration of stormwater, and once saturated, can behave like an impervious surface, transporting contaminant laden stormwater further within the roadside environment or directly to receiving waters. If precipitation is in the form of snow, metals and sediment are transported to roadside environments via plowing to snow banks, which accumulate metals and particulates (Sansalone and Buchberger, 1996; Sansalone and Glenn III, 2002; Glenn III and Sansalone, 2002), deicing salts (Buttle and Labadia, 1999), and cyanide de-bulking agents (Paschka et al., 1999).

Soluble components are readily released with meltwater, while insoluble and particulate bound species are more slowly released (Marsalek et al., 2003) or remain once the snowbank disappears (Sansalone et al., 2003). With applications of deicing salts, aqueous phase metals concentrations increase in snowbank meltwater (Novotny et al., 1998). Regardless of phase, contaminant concentrations may be one to two orders of magnitude higher during the winter (Glenn III and Sansalone, 2002; Backstrom et al., 2003). Concentrations of metals and salts associated with snowbank meltwater may have the most significant impact on receiving waters (Lygren et al., 1983). As a result, there has been increased usage of calcium magnesium acetate (CMA), as it is less toxic to aquatic life and does not contain cyanide de-bulking agents (Amrhein et al., 1993). However, high concentrations of metals and salts that are not discharged to receiving waters infiltrate into the soil matrix as frozen grounds thaw, and both NaCl and CMA salts increase the mobility of metals in roadside soils (Amrhein et al., 1992; Bauske and Goetz, 1993; Norrstrom and Jacks, 1998; Norrstrom and Bergstedt, 2001; Backstrom et al., 2004). Increased metal mobility through roadside soils due to the use of deicing compounds may result in groundwater contamination (Granato et al., 1995).

Integrated metals budget

When developing an integrated budget for the transport of highway generated metals to roadside environments, one study found that ~8% of metals were removed in runoff, ~6% were removed via atmospheric deposition to environments within 50 m of the roadway, and the remaining ~86% were distributed away from the transportation corridor (Hewitt and Rashed, 1990). This has significant implications for the application of LIAs as a sink for the contaminants in stormwater runoff. If runoff contains ~8% of transportation generated metals, then natural dispersion of this stormwater roughly doubles the metal load to roadside environments (adding to the ~6% of metals deposited via atmospheric deposition).

 

Impact of metals on terrestrial organisms

If immediately discharged to receiving waters directly from the road surface, metals in stormwater may be predominantly in a dissolved form, and thus available for uptake by and subsequently toxic to aquatic organisms. But direct runoff can have little effect on rural stream invertebrates exposed to runoff from lightly traveled or rural highways (Smith and Kaster, 1983). In contrast, metals in runoff are quickly rendered relatively immobile in terrestrial environments, as they are readily adsorbed to organic matter and retained in soils. While this limits exposure of terrestrial organisms, metals are still bioavailable for uptake and biomagnification (Scanlon, 1986). Quantifiable tissue concentration gradients of metals with distance from the road have been documented for all manners of organisms, including plants (Cannon and Bowles, 1962; Lagerwerff and Specht, 1970; Dedolph, 1970; Motto et al., 1970;Chow, 1970; Ward et al., 1975; Ward et al., 1977; Haqus and Hameed, 1986; Ylaranta, 1995; Parkpian et al., 2003), invertebrates (Williamson and Evans, 1972; Gish and Christensen, 1973; Ash and Lee, 1980; Wade et al., 1980; Weigmann, 1991), and mammals (Jefferies and French, 1972; Welch and Dick, 1975; Scanlon, 1986; Nyangababo, 2001). Lead and Cd have also been recorded in the milk of dairy cows grazing on fertilized pastures with elevated Pb and Cd levels in forage and soils next to roads, but concentrations were well below health standards and did not differ from pastures not subjected to metals contamination (Parkpian et al., 2003). With regard to vegetation, generally one to two thirds of metals are associated with atmospheric deposition rather than uptake, and different species have widely varying abilities to uptake metals. Metals that are taken up by vegetation are re-released as vegetation dies and decomposes. It is then subject to continued biotic uptake, particulate adsorption, leaching or overland transport. Some plant species have shown an adaptive tolerance to both metals (Schmidt, 2003) and deicing salts (Beaton and Dudley, 2004), although both can inhibit seed germination (Hsu and Chou, 1992; Beaton and Dudley, 2004).

While increased accumulation of metals for most biota with road proximity has been amply demonstrated, toxicological research was not reviewed and would best be found in toxicology literature and not transportation research. Interestingly, while there has been no demonstrated significant population number or density differences for vegetation, mammals, and certain invertebrates with increasingly metal contaminated soils, some studies found that certain invertebrate populations increase with proximity to the road (Muskett and Jones, 1980; Braun and Fluckiger, 1984), speculatively due to decreased predation from organisms not capable of living at the road edge and from increased susceptibility of vegetation due to a decline in health. Finally, with regard to soil microorganisms, biomass has been shown to decrease with increasing metals concentration (Khan and Scullion, 2000).

 

Conclusion

As lead is no longer an additive to gasoline, the lead loading and therefore the known toxicity of transportation generated contaminants has been lessened (Helmers et al., 1995; Legret and Pagotto, 1999; Turer et al, 1991). However, the very presence of lead and its toxicity stimulated the extensive research reviewed in this paper of how automotive contaminants are transported and where they accumulate in the surrounding environment. The lessons learned are applicable to other metals and particulates, and serve to enrich our understanding of the fate and transport mechanisms of automotive contaminants as yet unknown.