FATE AND REMOVAL OF CONTAMINANTS IN URBAN ENVIRONMENT

Stormwater runoff from impermeable surfaces, such as roads, roof tops, and parking lots, is a cause of non-point source pollution problem. Sources such as fossil fuel combustion, automobiles, and other human and animal activities introduce high concentration of organic, inorganic and pathogenic contaminants in to environment. Stormwater runoff carrying these contaminants when untreated has the potential to impair water quality in the receiving water bodies such as surface and ground waters. The existing treatment systems are not reliable and needed improvements in order to meet water quality standards. Determining the levels and sources of contamination in an environment can help mitigate the pollutant and reduce risk to humans. In here we have developed materials for use in stromwater treatment and applied source apportionment techniques to determine the sources of polycyclic aromatic hydrocarbons in surface soils of San Mateo Ixtatán, Guatemala. In chapter 2 of this dissertation, we evaluated the pollutant removal capacity of pervious concrete pavement systems and propose a methodology to enhance the containment of polycyclic aromatic hydrocarbons (PAHs) through organic modification of Rhode Island soils. In chapter 3 we have proposed using nano and polymer based materials for treatment of stormwater runoff contaminants such as pathogens, organic and inorganic contaminants. The modification materials had successfully increased the contaminant removal capacities of the materials. These proposed materials can find applications in Best management practices such as tree filters. In chapter 4 we have provide the levels, sources of PAHs in surface soils of San Mateo Ixtatán, Guatemala. In addition cancer risk assessment due to exposure to surface soil was performed.

property damage, and human health hazards. In addition, pollutants deposited on impermeable surfaces are mobilized by storm runoff to collection systems, and discharged to aquatic ecosystems with little or no treatment [1,2,3] including discharges to groundwater.
The pollutants in stormwater include organic contaminants such as polycyclic aromatic hydrocarbons (PAH), inorganic contaminants such as heavy metals, nutrients such as nitrates and pathogens such as bacteria [1,3,4]. Most of the pollutants present in urban stormwater runoff originate from nonpoint or diffuse sources. These are notoriously difficult to locate and quantify and include wet and dry atmospheric deposition from industrial and domestic properties; traffic emissions; decomposed litter; de-icing salts; vegetative residues; pet feces; soil losses, among others [4][5][6]. The concentrations of these contaminants are often order of magnitude above the maximum contaminant level (MCL)for drinking water standards [1,3,4][7] which can impair the water bodies. Some of the contaminants of concern are PAHs, heavy metals and pathogens.
The fate and transport of some of these contaminants was studied herein, including PAHs, heavy metals and pathogens. Polycyclic aromatic hydrocarbons are class of compounds originating from the incomplete combustion of organic materials, including petroleum products and biomass. The US Environmental protection agency (US EPA) has listed 16 PAHs as priority pollutants of which seven PAHs are listed as probable human carcinogens or mutagens. PAHs are toxic, carcinogenic and mutagenic compounds with drinking water MCL for PAHs is of 0.2 µg/l [7].
Heavy metals are other group of contaminants found in urban runoff, this group includes but not limited to metals such as cadmium, lead, zinc, nickel and copper. The primary sources of these metals in urban runoff are automobiles, tires, and abrasion of road surfaces. The pathogens present in urban stormwater runoff are related to elevated waterborne illnesses after storm events.
Escherichia coli (E. coli) is bacterium that is commonly monitored in surface and drinking water as an indicator for pathogenic contamination. The concentration of E. coli in stormwater runoff often exceeds 10 4 colony forming units (CFU) /100ml [3]. The high microbial concentration in stormwater not only has negative impact on human health, but also is resulting in local economic impact through beach closing and fishing bans.
In order to reduce the volume and pollutant loads of stormwater runoff structural best management practices (BMPs) were adapted by agencies such as EPA, Department of environmental management (DEM) and Department of Transportation (DOT). Examples of these BMPs include pervious pavements, retention/detention ponds, tree filters, bio-swales and infiltration trenches. Previous research demonstrated that the installation of these BMPs has resulted in reduction of runoff volume and lower pollutant concentrations. Although BMPs such as porous pavements and tree filters have shown to remove total suspended solids and heavy metals the long term performance of these systems is unreliable [8][9][10]. Further the removal of contaminants such as PAHs and pathogens has not been fully tested. Modification of materials with quaternary ammonium compounds, or nanoparticles, for example have shown to significantly increase the contaminant removal of filter media [11][12][13] and therefore carry promise as amendments to existing BMP or to be the basis of an entirely new line of BMPs.
In urban and suburban environment PAHs are released to environment mainly due to anthropogenic sources such as oil spills, vehicle exhaust, industrial smoke stack emissions, combustion of fossil fuels and biomass burning (such as wood) [14][15][16]. PAHs emitted in to atmosphere through these sources can contaminate surface soil and water systems. Several studies have reported high concentrations of PAH in surface soils and sediment around the world. [4,[14][15][16][17][18] In developing countries nearly 90% of households living in rural areas rely on biomass for their energy needs. [19] Identification of PAH contamination and source apportionment therefore important for pollution control, remediation to minimize risk of exposure and can help plan intervention studies.
The objectives of this dissertation work were 1) to develop a methodology to enhance the organic contaminant removal capacity of Rhode Island soils and test the contaminant removal capacity of porous concrete pavement systems; 2) to develop a filter media amended with nanoparticles or quaternary ammonium polymer to enhance the simultaneous removal of PAHs, metals, and pathogens from stormwater runoff and, 3) to determine the concentration and sources of PAHs in surface soils in San Mateo Ixtatán, Guatemala, and perform a risk assessment References [1] P. Göbel

Introduction
Impermeable surfaces such as conventional concrete and asphalt have the potential to generate large volumes of contaminated storm-water runoff during precipitation events. Storm-water runoff is a source of organic and inorganic contaminants that can end up polluting natural bodies of water [1]. Of particular concern are polycyclic aromatic hydrocarbons (PAHs), which are a family of organic compounds containing two or more benzene rings.
PAHs are ubiquitous in the environment and present in the soil, water, and air. Most PAHs are acutely toxic and suspected human carcinogens [2]. The Safe Drinking Water Act enforces a maximum contaminant level of 0.0002 mg/L for PAHs [3]. Acute oral toxicity of PAHs range from 50 to 2000 mg/kg of body weight depending on the specific compound [2]. One of the main characteristics of PAH is that, as the number of benzene rings increases, the solubility in aqueous solutions decreases. Due to the hydrophobic properties of these compounds, PAHs are rapidly sorbed onto sediment particles and subsequently deposited. As previously demonstrated, the rerelease of PAHs into clean or less contaminated waters is possible [4,5]. After a rainfall event, PAHs are carried into the drainage infrastructure and are eventually transported into ground and surface waters. Hoffman et al. [4] reported that urban runoff entering Narragansett Bay was responsible for 71% percent of the total inputs of higher molecular weight PAHs and 36% of total PAHs. Another study, spanning from 1993 to 2001, showed that storm-water runoff contributes about 51% of all PAHs to the San Francisco Bay [6].
As vehicles produce the greatest concentration of PAHs during ignition and acceleration [7], roads and parking lots are ideal locations for containing the transport of these contaminants before they are discharged into natural ecosystems. In most major American cities, parking lots account for 10% to 15% of the total impervious area [8], and impervious areas in urban and suburban settings are expected to increase in the future. In addition to automotive-related releases, coats of pavement sealants and oil leakages are also important sources of PAHs [4,[9][10][11].
Engineered systems can mitigate the discharge of PAHs from parking lots, thereby drastically reducing the influx of contaminants into water resources. For example, pervious pavement has been used to reduce both the volume and contaminant load in storm runoff [12][13][14][15]. In addition, pervious concrete pavements can elevate pH of runoff from acidic nature [16]. However, conventional pervious pavements systems have a limited contaminant removal capacity [17]. The pollutant removal performance of the pervious pavement systems is not consistent over time and also depends on location [17].
Previous studies have demonstrated that, by increasing the fraction of carbon (f oc ) of a sorbent, the sorption of PAHs increases [18][19][20]. Several approaches have been suggested to increase the foc of soils or other engineered materials [18,21], such as activated carbon [22,23], which can be expensive, can be labor intensive, and may require frequent maintenance. The PAH compounds acenaphthene and flourene were used for the second part of the study.

2.1Overview of Experimental Methodology
The experiment was divided into three phases. The first phase involved the synthesis and characterization of the organically modified soil amendments. Second, the performance of the modified soil in terms of PAH sorption was assessed using naphthalene as the model PAH. This phase involved both batch isotherms and column experiments. Naphthalene was chosen due to its high aqueous solubility. The third phase focused on the assessment of PAH-retaining capabilities of conventional and organically modified pervious pavement elements. In these column experiments, acenaphthene and fluorene were used as model PAHs. These PAHs were selected as they have high aqueous solubility, have higher molecular weight, and are commonly found at contaminated sites.

Materials
Naphthalene, fluorine, and acenapthene (purity grade of 98% or higher) were obtained from Aldrich. Analytical grade solvents (dichloromethane and methanol) were obtained from Fisher Scientific. Aqueous solutions were prepared in deionized water free of detectable traces of PAH and with a pH near neutral. PAH are nonionic compounds, there is no influence of pH on the characteristics of PAH in the aqueous phase. Furthermore, the pH of runoff originating from traffic areas ranges between 6.4 and 7.9 [1]. The pH used for our experiments is 7.0, which is well within the narrow range of natural conditions. A gas chromatograph with flame ionization detector (Shimadzu GC-17 A/FID) and a gas chromatograph/mass spectrometer (Shimadzu GC-MS QP2010) were used to quantify PAH compounds. Only glass vessels were used, and care was taken to prevent PAH photo degradation. Hexadecyltrimethyl ammonium chloride (HDTMA) and benzyldimethylhexadecyl ammonium chloride (BHDH) were obtained from sigma Aldrich, with a purity of 99%; they were used as received.
The PAH-saturated solutions were prepared following a method described by Wang et al. [24]. In brief, PAHs were dissolved in 1 mL of dichloromethane in a 4 L Erlenmeyer flask. As the solvent volatilized, a thin film of PAH formed on the glass. The flask was then filled with deionized water, which was allowed to saturate and equilibrate with the PAH over the course of 1 week.
The initial concentration (C 0 ) of all three target PAHs in the aqueous phase was determined using the gas chromatograph-mass spectrometry.
A bulk sample (20 kg)  to 18%, which is typical for this type of material [26]. The porosity (Eq. (1)) of the materials was determined using the volume of voids which was determined by identifying the difference in weight between the dry sample and a water-saturated sample.

Synthesis and Characterization of Organically Modified Soil Amendments
The cation exchange capacity (CEC) of the soil was determined by the Soil, Water, and Plant Testing Laboratory at Colorado State University. A characterization of the soil particle size distribution was conducted on all modified and unmodified soils through sieve and hydrometer analysis using ASTM 136 and 152 H methods [27]. The hydraulic conductivity of the concrete, aggregate and soils was determined using the falling head permeameter method [28]. HDTMA and BHDH are long-chained cations with relatively high carbon content; therefore, they are considered to be an attractive choice for increasing the f oc of the porous matrices.
Equation (2) was developed by Boyd et al. [29] to determine the mass of quaternary ammonium cation (QAC) with respect to the mass of soil and CEC. f = M cation / (CEC * M soil * GMW cation * z) (2) where CEC is the cation exchange capacity in milli-equivalents per gram or soil, f is the fraction of CEC exchanged, M cation is the mass of cation used in grams, M soil is the mass of soil in grams, GMW cation is the formula weight of cation in grams per mole, and z represents the moles of charge per equivalent moles per milli-equivalent of exchange capacity.
The wet method developed by Breakwell et al. [30] was used to modify the soils. Briefly, the amount of QAC necessary to produce 100 g of 100% CEC modified soil was calculated using Boyd's equation. The QAC was dissolved into 400 ml of deionized water using a stir bar and a glass beaker. Soil was mixed into the QAC solution and agitated for 3 days at 150 rpm in an incubator at 20°C. The contents of the glass beaker were carefully decanted after settling, and the soil was placed into the oven for 6 h at 100°C to dry. The residual chloride ions were removed by rinsing with deionized water until electrical conductivities were below 1 µs/cm. Once dried, the modified soils were analyzed using a Carlo Erba EA1108 Carbon Hydrogen and Nitrogen analyzer to quantify the amount of carbon.
The second approach for introducing f oc involved mixing the glacial outwash soil with commercial organoclay (PM-199; CETCO Oil Field Services). To maintain a point of reference in isotherm and column experiments, the mass of the total organic carbon in the blend was calculated to match the mass of total organic carbon in the BHDH-modified soil. This resulted in a 1:18 ratio of commercial organoclay and Rhode Island glacial outwash.

Sorption Isotherms
A series of batch isotherms were performed for all sorbents to study the static interaction between sorbent and aqueous PAH. A constant mass (different for each soil) of the particular sorbent was

Column Experiments
Column experiments were conducted to study the sorption and desorption of PAHs under dynamic, flow-through conditions. For the concrete and aggregate materials, larger glass columns were performed to determine pore velocity and dispersion coefficient. After determining the column intrinsic parameters, an aqueous solution saturated with the target PAH was pumped through the columns. Effluent samples were collected until breakthrough, which defined as C/Co equaling one (i.e., the effluent concentration is equal to the influent concentration). The flow velocities were held constant at 2 mL/min for soil columns and 20 mL/min for concrete and aggregate columns. After breakthrough, the influent was switched to deionized water to study desorption. The pumping of deionized water continued until at least 90% of the PAH was recovered. This same procedure was repeated for concrete, aggregate, and all other soil media.

HYDRUS 1D
HYDRUS-1D (version 4.14) software was used to stimulate solute transport through unmodified and modified soils. Parameters obtained from sorption isotherms, soil characterization, and column experiments were used to predict the breakthrough curves (BTCs) for naphthalene in unmodified and modified soils. Langmuir and Freundlich isotherm coefficients were obtained from isotherm experiments. Dispersion coefficient and bulk densities were obtained from column tracer experiments. Hydraulic conductivities were obtained from falling head permeameter tests.
HYDRUS-1D was also used to estimate the adsorption coefficient from column experiments using the inverse solution to the Levenberg-Marquardt nonlinear parameter optimization method.

Synthesis and characterization of organically modified soil amendments
A CEC of 5.6 milli-equivalents of exchangeable cations per 100 g of glacial outwash soil (meq/100 g) was measured. These results were consistent with the CEC values determined by Wright and Sautter [31]. The results from soil modification, particle size analysis, and hydraulic conductivity measurements are summarized in Table 2-1. Also shown are the initial f oc and the percentage of f oc increase resulting from the soil modification. QACs HDTMA and BHDH increased the fraction of organic carbon in the Rhode Island glacial outwash soil by 70.6% and 72.9%, respectively. The f oc increase in the blend was calculated to match the organic carbon in the BHDH-modified soil. This resulted in mix of 1:18 commercial organoclay and glacial outwash. The hydraulic conductivities of the modified soils were an order of magnitude smaller than the unmodified soil. Decreases in effective grain size (D 10 ) and porosity were also observed, although the reasons for these decreases were not investigated.
Conversely, adding the PM-199 organoclay to the unmodified soil increased the effective grain size D 10 of the blend while decreasing the uniformity coefficient and increasing the hydraulic conductivity. Thus, a material with increased hydraulic conductivity, such as this organoclayglacial outwash blend, will be best suited for application below permeable pavement because it will aid in faster drainage of storm water.

Sorption Isotherms
Kinetic investigation revealed an equilibrium time of 30 h (data not shown) between the soils and naphthalene, which is consistent with prior research [32,33].

Column experiments with unmodified, organically modified, and blended Rhode Island glacial outwash soils
Parameters for the four column experiments, including pore volume, pore water velocity, and dispersion coefficients, are shown in Table 2   The CXT fit model [36] was used to determine column intrinsic parameters and the retardation factors for naphthalene. CXT fit allows for inverse estimation of transport parameters from a laboratory study. The program uses convection dispersion equation with a data set, using the nonlinear least-squares parameter optimization method. Together with experimental data,

Desorption Desorption
Desorption Desorption including hydraulic conductivity, bulk density, and isotherm coefficients from sorption studies, parameters determined with CXT fit were used as inputs to predict BTCs with HYDRUS. The predicted BTC (Figure 2-2) compared poorly with the measured batch experimental data.
Similarly, the predicted adsorption isotherm coefficients ( Table 2-2) for column experiment data were lower than the values from batch experiments. Maraqa et al. [37] also reported that their batch study overestimated adsorption coefficients of benzene and dimethylphthalate compared with the results from the column study. Similarly, Lee et al. [38] reported a poor match of their predicted BTCs relative to the ones measured for naphthalene. However, the Langmuir coefficients predicted from column experiments for glacial outwash are in close agreement with isotherm coefficients. The fraction of Type 1 adsorption sites at which sorption is assumed to be instantaneous was set to one, and the first-order rate constant was set to zero for the calculation of BTCs and the inverse solution. HYDRUS permits chemical nonequilibrium (adsorptiondesorption process) and physical nonequilibrium (possible heterogeneity of soil). As none of these processes were considered in the modeling of the results, this could be a reason for the poor prediction of BTCs. Additional experiments are required to obtain the parameters needed to include chemical and physical nonequilibrium processes. The HYDRUS modeling was test if the results can be replicated using this computational tool. The results suggest that HYDRUS and these results will support a more thorough future modeling study regarding determining the life time and optimizing the ratio of organoclay to glacial outwash. More data and modeling effort is needed to obtain an optimal fit. The BTCs obtained using the isotherm data was used to compare the experimental data and modeling data. This allowed us to determine how effective the modeling data is when compared to laboratory scale/field scale results. Retardation factor obtained from breakthrough curve plotted in the unitless terms of relative concentration (C/Co) and pore volumes (PV). * G.O unmodified glacial outwash

Performance of convention and organically modified pervious concrete pavements
As potential parts of a pervious pavement system, the capacities of porous concrete and aggregate for retaining PAH of different molecular weights were determined through column experiments ( and R f = 7 for acenaphthene and flourene, respectively. Due to the very low organic matter, the retardation of PAHs in concrete and aggregate is significantly lower compared to in soils [18][19][20].
Similarly, low retardation was found in the case of the glacial outwash soil (R f = 16.5; Figure    ⍴ is bulk density, PV is pore volume, Ace is Acenaphthene, Flu is Fluorene, R f is Retardation factor * The column volume of concrete and aggregate columns were 850 cm3 ** Column volume is 24.5 cm3 Commercial organoclay ranges in cost between one to two dollars per pound. If the organoclayglacial outwash blending ratio and bulk density used in the study was kept constant, the cost of the organoclay material would average about $200.00 per cubic meter. However, the thickness of the modified section could be reduced to decrease the cost. Also, the QAC loading on the soil could be optimized to further reduce prices. The cost-benefit analysis and optimization of the enhanced sorption section is beyond the context of this study but will be a subject of our future studies.

Conclusions
The adverse effects of contaminants commonly present in storm-water runoff can be minimized using BMPs, such as pervious pavement systems. The results of this study demonstrated that conventional pervious pavement components, such as porous concrete, aggregate, and unmodified soil, have little capacity to retain PAHs from the aqueous phase. The modification of glacial outwash using QAC increased the fraction of organic carbon in the soil and greatly enhanced the PAH sorption capacity of the soils. After flushing several 100 pore volumes of PAH saturated water, the modified soil media removed up to 74% of PAHs (i.e., irreversible sorption).
Thus, modified soil media, such as a blend of organoclay and glacial outwash, could find applications in BMPs to reduce contaminant flux into surface water or groundwater. These modified organoclay-glacial outwash blends can potentially be incorporated as a layer beneath pervious pavement. In addition to retaining PAHs, their higher hydraulic conductivity relative to the unmodified soil is an added advantage. However, cost is an important factor when incorporating modified materials into permeable pavement design. Currently, the cost of commercial organoclay ranges from 1 to 2 dollars per pound, which may prevent the widespread application of this promising amendment.

Abstract
The objective of this study was to develop and test nanoparticle and polymer based bioactive amended sorbents to enhance stormwater runoff treatment in best management practices (BMPs). Red cedar wood and expanded shale were the sorbents tested. Red cedar wood chips (RC) were modified with 3-(trihydroxysilyl) propyldimethyloctadecyl ammonium chloride (TPA) and silver nanoparticles (AgNPs) at different mass loadings (3.6 mg/g, 6.7 mg/g and 9.3 mg/g for TPA and 0.33 mg/g and 0.68 mg/g for AgNPs) to simultaneously improve the sorption of organic and inorganic contaminants and pathogenic deactivation in BMPs treating stormwater runoff.
Unmodified expanded shale is often used as a filter material for stormwater treatment and was used as a base comparison. The results showed that TPA and AgNP loading onto red cedar increased the Langmuir maximum sorption coefficient (Q) for polycyclic aromatic hydrocarbons, up to 35 fold and 29 fold, respectively compared to unmodified red cedar. In case of heavy metals, Q for lead increased with increased loading of TPA and AgNPs, whereas no significant change in the Q value for cadmium was observed, while zinc and nickel sorption slightly decreased. The Langmuir maximum sorption coefficient of copper was higher for modified red cedar, however no correlation was observed with TPA or AgNP loadings. The log reduction value (LRV) for Escherichia coli using unmodified red cedar was <1 log, while modified red cedar exhibited LRV up to 2.90±0.50 log for 6.7 mg/g TPA-RC and up to 2.10±0.90 log for 0.68 mg/g AgNP-RC. Although AgNP modified red cedar shows a comparable performance to TPA-RC, the high cost of production may limit the use of AgNP amended materials. While TPA modified red cedar has advantages of lower cost and lower toxicity, the fate, transport and environmental implications of TPA in natural environments has not been fully evaluated. The findings from this study show that if BMPs were to incorporate the modified red cedar, stormwater treatment of PAH and E.coli could be enhanced and the quality of the treated water will improve.

Introduction
Stormwater runoff contains polycyclic aromatic hydrocarbons (PAH), heavy metals, and pathogens that are discharged into natural surface and groundwater bodies, impairing ecosystems and compromising human health1-3. During and after precipitation events, these contaminants commonly exceed the maximum contaminant level (MCL) standards in runoff and receiving water bodies 1, 4, 5 . High concentrations of heavy metals and petroleum hydrocarbons, such as PAHs, 3 can compromise the ability to use stormwater for recharging aquifers or apply it in greywater operations. The concentrations of Escherichia coli (E.coli) in runoff can exceed 104 colony-forming units (CFU) per 100 ml1, 6 making stormwater runoff one of the major contributors of pathogens into surface and coastal waters 5,7 . The sources of pathogens in stormwater runoff are attributed to wildlife or pets 8 and, to some degree to human fecal contamination. 7 Exposure to pathogens can lead to serious illness, such as gastroenteritis or cholera. 5,9 Ideally, structural BMPs should simultaneously attenuate organic, inorganic, and microbiological contaminants. However, most stormwater BMPs are only effective in treating heavy metals and petroleum hydrocarbons 10,11,12 through filtration and sorption 13,11,12 . They are largely ineffective in treating pathogens 14, 15 , 16 , 17 . Previous studies showed that materials such as organoclays amended with quaternary ammonium compounds exhibit higher sorption for PAHs and metals. [18][19][20][21] The quaternary ammonium polymer 3-(trihydroxysilyl) propyldimethyloctadecyl ammonium chloride (TPA) is a used as a disinfectant material in environmental and medical applications such as ceramic filters 22 and prosthetic devices 23 . Silver nanoparticles (AgNPs) are another well-known antimicrobial agent. Impregnating filters media with AgNPs effectively removes pathogens from aqueous solutions. 24,25 To our knowledge these nanoparticles and the TPA polymer have never been used for stormwater runoff treatment.
One commonly used BMP is a tree filter, which consists of a subsurface biofiltration system that combines filtration, sorption, and phytoremediation processes for contaminant removal. However, previous studies have shown that contaminant removal is limited. 13

Materials
Untreated red cedar wood was obtained locally (Liberty Cedar, West Kingston, RI). The wood was chipped, and the fraction of RC chips retained between 10 and 3. No. 83019-2) was obtained from Biosafe (Pittsburgh, PA). AgNPs were synthesized via Tollens method as described elsewhere, 27 using polyvinylpyrrolidone (PVP, average molecular weight 29,000 g/mol, Sigma-Aldrich) as a stabilizer.
Two common stormwater PAH compounds, acenaphthene and fluorene (purity grade of 98% or higher) were obtained from Sigma-Aldrich. The PAH solutions were prepared as described elsewhere. 21 PAH standards and deuterated PAH standards were obtained from Ultra Scientific, U.S.A. Metal reference standards containing 1000 mg/L ± 1% certified cadmium, copper, lead, nickel, and zinc were obtained from Fisher Chemical. Sodium sulfate (Na2SO4), disodium phosphate (Na2HPO4), and sodium nitrate (NaNO3) were obtained from Sigma-Aldrich. A nonpathogenic wild strain of Escherichia coli (E. coli) was obtained from IDEXX laboratories.

Material Characterization
The

Experimental Methodology
The laboratory experiments were divided into two phases. During the first phase, the loading capacity and stability of TPA and AgNP amendments on RC and ES were tested. This was achieved through batch sorption isotherms (loading capacity) followed by sequential desorption for a week (stability). In the second phase, building on the isotherm results obtained during the first phase, RC was modified with different TPA and AgNP loadings and evaluated in batch experiments for sorption capacity of PAHs and heavy metals and also E. coli disinfection performance.

Phase I: Nanoparticle and Polymer Loading Capacity of Sorbents
Batch isotherms were carried out in triplicates to determine the loading capacity of TPA and AgNPs onto the sorbent materials. For all experiments, a 1.3 mmol/L ionic strength medium using sodium chloride (NaCl) was used as a background solution in order to mimic surface water conditions in Rhode Island. 27 Detailed experimental procedures are in Text S1 of the Supporting Information. After conducting the sorption experiment, the samples were decanted and dried at 60 °C. Desorption of TPA and AgNPs from modified sorbent materials was determined through sequential desorption in DI water for a minimum of 24 h and up to 168 h, i.e., until aqueous concentrations were below the method detection limit of 0.2 mg/L for TPA and 0.01 mg/L for Ag. The method detection limit for Ag is below EPA MCL of 0.1 mg/L, while for TPA, no EPA MCL limit is defined. TPA is categorized as slightly toxic with oral LD 50 above 5000 mg/L. The results from the desorption experiment were reported as total silver and TPA, respectively. The fractions of desorbed AgNPs and Ag ions are also measured and reported in Table S2 of the Supporting Information.

Phase II: Organic and Inorganic Contaminant Removal Efficiency of Modified Sorbents
Using the results obtained in phase I, RC was modified at different loadings of TPA and AgNPs (Table 1) to assess the ability to remove organic, inorganic, and microbiological contaminants as a function of loading. This process was required to adapt the lab-scale amending procedure to a large scale due to the large amount of the material required to perform all tests; details of this procedure are presented in Text S2 of the Supporting Information. Due to disintegration of ES when agitated for modification, this material was no longer tested. However, unmodified ES was used as baseline comparison for modified RC.
Batch tests were carried out to determine the sorption capacity of the modified and unmodified sorbent materials for PAHs and heavy metals. A synthetic stormwater runoff stock solution, containing PAHs, metals, and inorganic salts, was prepared (Table S3, Supporting Information).
All solutions were adjusted to a pH of 5.5. Details of the batch tests are described in Text S3 of the Supporting Information. Visual MINTEQ Ver.3.0, an equilibrium speciation software, was used to determine the speciation of metals in the synthetic runoff.

Microbiological Contaminant Removal Efficiency by Modified Sorbents
Two of the modified RC sorbents (6 mg/g TPA-RC and 0.6 mg/g AgNP-RC) along with unmodified RC and ES were selected to test bacteria deactivation efficiency. Five different initial E. coli concentrations relevant to stormwater runoff, ranging from 10 2 -10 6 CFU/100 mL, were used in the experiments. 6,7 Detailed experimental procedures are provided in Text S4 of the Supporting.

Material Characterization
The average hydrodynamic size of the AgNPs was 38.5 ± 3.5 nm. The surface area of the modified RC increased with increased loading of AgNPs and TPA, for example, the surface area increased from 4.89 ± 0.092 m2/g for UM-RC to 7.98 ± 0.42 m2/g and 6.01 ± 0.23 m2/g for 9TPA-RC and 0.6AgNP-RC. The contact angle of water for modified RC increased with loading of TPA and AgNPs (Table S4, Supporting Information). The contact angle was greater than 90° for 9TPA-RC and was greater than 70° for 6TPA-RC and 0.6AgNP-RC, indicating that modification with TPA and AgNPs made the material more hydrophobic.

Phase I: Nanoparticle and Polymer Loading Capacity of Sorbents
The TPA batch sorption on RC showed that between 37% and 98% of the initial mass of TPA in the aqueous phase was sorbed to the RC, with a maximum loading of 0.93 ± 0.03 mg/g. For expanded shale (ES), the total mass of TPA sorbed was in the range of 30% to 75%, with a maximum loading of 0.30 ± 0.03 mg/g. During the sequential desorption study, only 0.54 ± 0.30% of the TPA initial mass sorbed was released from RC, while for ES, it was 1.13 ± 0.60%.
The higher sorption of TPA to RC is likely due to stronger interaction with the wood's molecules, such as lignin and cellulose. 28 The nonlinear shape of the sorption isotherm model (Figure 3-S1a, Supporting Information) supports that TPA predominantly sorbs onto RC rather than diffusing into it, which would have been by linear sorption isotherm, indicative of Fickian transport processes. In comparison, ES contains little or no organic matter due to the extreme heating during expansion process. 29 This likely resulted in fewer sorption sites for TPA. The linearity of the ES isotherm ( Figure 3-S1a, Supporting Information) indicates that TPA may be partitioning into porous spaces in the shale.
The amount of AgNPs sorption to RC depended on the initial concentration of the AgNP solution.
That is, RC sorbed up to 97% when the initial concentrations of AgNPs were below 21 mg/L. were desorbed if the AgNP loading was low (0.33 mg/g) and 14.4 ± 9.28% if the loading was high (0.63 mg/g) (Figure 3-1b). Depending on the initial aqueous phase concentration of AgNPs, between 88% and 98% of the mass desorbed is in the form AgNPs, while the remainder is in form of Ag ions (Table 3-

Phase II: Organic and Inorganic Contaminant Removal Efficiency of Modified Sorbents
The results of the batch isotherm experiments for PAHs and heavy metals onto unmodified ES as well as unmodified and modified RC are nonlinear (Figure 3-2). Assuming that the number of sorption sites is limited, the results were fitted to the Langmuir model Equation 1 where q is the amount of solute sorbed (μg/g), Q is the Langmuir maximum amount of solute that can be absorbed by the sorbent (μg/g), b is the Langmuir adsorption coefficient, and Ce is the for all sorbents were obtained using SigmaPlot 11 (Systat Software, Inc.) ( Table 3-2). Except for ES, the goodness-of-fit using the Langmuir model was high (R2 >0.90) for all isotherms.

Polycyclic Aromatic Hydrocarbons
The Langmuir maximum sorption capacity of PAH for ES is much lower compared to both unmodified and modified RC (Table 3-2), which may be a result of the vitrification process the expanded shale has undergone during thermal expansion, making it more inert by removing potential organic sorption sites. 29 In the case of RC, the value of Q for acenaphthene and fluorene increased with increased loadings of TPA and AgNPs (Figure 3-2). For example, for acenaphthene, Q increased from 48.6 μg/g for UM-RC to 1703.5 μg/g for 9TPA-RC and to 1429.2 μg/g for 0.6AgNP-RC (Table 3-2).
Finally, the modification with TPA and AgNPs increased the hydrophobicity of the RC as confirmed by contact angle measurements (  30 The increase in Pb and Cu sorption could be due to bonding with hydroxyl groups on TPA and formation of complexes with PVP present on AgNPs. In addition, the increase in surface area of modified RC (Table 3-S4, Supporting Information) likely provided more sorption sites for these metals. However, the reasons for the slight decrease in Q for Zn and Ni (Table 3-2) is unclear. No correlations with physiochemical properties or interactions with the amendments were found to explain the different sorption behavior of these metals to modified RC. Therefore, further studies are required to gain insight on the sorption behavior of these metals onto modified RC.

Microbiological Contaminant Removal Efficiency of Modified Sorbents
Two modified sorbents (6TPA-RC and 0.6AgNP-RC) were tested for bacteria deactivation efficiency and were compared to unmodified RC and ES. The disinfection performance of the materials was calculated as log removal value (LRV) LRV = log (initial E. coli concentration)log (final E. coli concentration) Equation2 Final E. coli concentration is the total E. coli. in aqueous phase and sorbed to the sorbents. The average LRV for the unmodified RC and ES was always below 1 (Figure 3-3). For the modified materials, the average LRV values ranged from 1.74 ± 1.04 log to 2.90 ± 0.50 log for 6TPA-RC.
In the case of 0.6AgNP-RC, the average LRV value ranged from 2.03 ± 0.60 log to 2.10 ± 0.90 log. Overall, the deactivation performance shows that the modified materials are significantly more effective at deactivating bacteria compared to unmodified ES and RC (p < 0.001) (Figure 3-3). Previous studies showed that the impregnation of TPA and AgNPs onto surfaces enhanced the deactivation of E. coli. 22 Several mechanisms have been suggested for disinfection using AgNPs.
These include damage to bacteria by pitting the cell membrane, lysis of cells caused by silver ion release, or damage of the cell by the reactive oxygen species formed on the surface of the AgNPs.
The biotoxicity of TPA has been explained mainly by the positively charged quaternary amine groups attracting E. coli and C18 groups the piercing membrane and causing cell disruption 31

Cost Analysis and Limitations
The contaminant removal performance and bacterial deactivation of both TPA-and AgNPmodified RC compare well and significantly enhanced water quality compared to unmodified RC.
However, the selection of which antimicrobial agent to use for modification will be driven by the costs of amending the materials. The initial modification cost for RC using AgNPs is higher compared to TPA (      Instead, unmodified ES was used a baseline for comparison with modified RC.

Text 3-S3: Organic and inorganic contaminant removal efficiency of the modified sorbents
The sorption capacity of modified and unmodified sorbent materials for PAHs and heavy metals were determined through batch tests. A synthetic stormwater runoff stock solution, containing PAHs, metals, and inorganic salts was prepared (Table 3-S2). Acenaphthene and Fluorene were chosen as representative PAHs due to their high solubility and consistent presence in stormwater runoff 4, while cadmium, copper, lead, nickel, and zinc were chosen as representative heavy metals due to the high concentrations found in runoff 5,6 . To prevent heavy metal precipitation the pH of the solution was adjusted to 5.5 using sodium hydroxide. The study was conducted by combining water saturated sorbent material (1.52 ± 0.052 g for RC and 3.50 ± 0.076 g for ES) with the synthetic runoff solution into 40 ml amber glass tubes with Teflon (PTFE) lined caps.
Dilutions over three orders of magnitude were prepared. Triplicates and appropriate controls were set for QC/QA. All samples were placed in a rotisserie incubator at five rpm and 25 °C for six days, as determined by kinetic study to determine equilibrium time (data not shown). Visual MINTEQ Ver.3.0 an equilibrium speciation model was used to determine the speciation of metals in the synthetic runoff.

Text 3-S4: Microbiological contaminant removal efficiency of the modified sorbents
The water saturated sorbent materials were weighed (1.59 ± 0.181 g for RC and 3.48 ± 0.059 g for unmodified ES) and exposed to E. coli prepared in 1.3 mmol/l ionic strength solution 1 for three hours at 25 °C and five rpm in a rotisserie incubator. The three hour exposure time was based on a kinetic study where a ~50% reduction of E. coli was observed within that time when         Wood combustion contributes more than 70.6% and 75.5% in town samples and agricultural soils, respectively. The estimated accumulation of ∑16PAH in corn grains cultivated in contaminated agricultural soils ranged between 0.73 µg/kg and 11.82 µg/kg of corn. The cancer risk assessment showed that incremental life time cancer risks (ILTCR) were greater than the acceptable level of 10 -6 . Main routes of exposure are through soil ingestion and dermal contact.
The dietary uptake of PAH through ingestion of corn resulted in a significantly higher cancer risk (ILTCR corn ) of 1.95x10 -3 and 4.54x10 -3 for adults and children, respectively. Analysis of sediment samples revealed that the ∑17PAH concentrations in the river sediments were three times higher for samples taken in the town compared to the upstream sediment samples. The study concludes that the contamination of surface soils and the high cancer health risk to local residents is due to the usage of wood as domestic fuel source. Alternative fuels sources and improvements to existing stoves should be considered to reduce the exposure.

Introduction
Polycyclic aromatic hydrocarbons (PAHs) are a class of persistent organic pollutants in the environment that possess carcinogenic and mutagenic properties. Atmospheric deposition of PAHs on soils occurred via solid, gaseous, and gaseous processes. Soil ecosystems are therefore considered major reservoirs and sinks for these hydrophobic organic contaminants [4]. PAHs accumulated in surface soils bond strongly to the soil's organic matter.
PAHs with three or more rings (i.e. higher molecular weight, HMW) have low water solubility, very low vapor pressures and high octanol/water partition coefficients, therefore they can strongly sorb to soil. Because of that, most PAH are persistent in the environment. High concentrations of PAH in soils and sediments have been reported around the world in a number of studies Occurrence of PAHs in agricultural soils is a public health concern, because they can be introduced into fruits and vegetables [10] and humans can get exposed through direct contact and ingestion. Assessment of PAH contamination and source apportionment are important for pollution control, remediation to minimize risk of exposure and can help plan intervention studies.
In developing countries nearly 90% of rural households rely on biomass for their energy needs [11]. Combustion of wood and other biomass fuels can release many pollutants including particulate matter, carbon monoxide and PAHs [12,13]. In developing countries, the pollution generated combustion of wood and biomass for domestic purposes has been linked to respiratory diseases [14,15]. In several developing countries traditional wood stoves have been modified to reduce in door emissions through the venting of the smoke outside the buildings [15][16][17][18]. This can result in increasing PAH atmospheric deposition on soils posing a health risk to humans.
The objectives of this study were to 1) to investigate PAH contamination in surface soils and sediments in the town of San Mateo Ixtatán, Guatemala, 2) determine the sources of PAH contamination, 3) estimate life time health risk for adults and children due to soil contamination.

Materials and methods
PAH standards and deuterated PAH standards were obtained from Ultra Scientific, USA. All solvents (Dichloromethane, n-Hexane, Acetone) were of HPLC grade and were obtained from Fisher Scientific.

Description of study area and sampling sites
The town of San Mateo Ixtatán (SMI) is located in Huehuetenango department, Guatemala (elevation 2540m) with a population of approximately 10,000. About 98% of the households use wood as a primary source of fuel for cooking and heating purposes [19]. In August 2012, surface soil samples (0-5cm) were collected at 24 sampling sites (18 town sites and 6 agricultural soils).
The 18 town sampling sites (T1 to T18) include residential areas with high and low housing density and commercial areas (Figure 4-1). Six agricultural soil samples (A1 to A6) three samples (A2 to A4) were taken in rural areas while the rest were collected in urban agricultural settings.
In addition, six sediment samples (SD1 to SD6) were collected along the river passing through the town in order to determine the contamination caused by PAHs from the town. The sediment samples SD1 and SD2 were taken at the beginning of the river and 0.5 km from SD1 to establish a baseline. Sample SD3, SD4, and SD5 were taken in the town. Locations SD4 and SD5 are where tributaries running through the town carrying household waste water and rainwater runoff merge into the river (Figure 4-1). The sampling of surface soil and sediments were done following EPA standard operating procedures [20][21].

Sample preparation, extraction, cleanup and analysis
All samples were air dried for one week at room temperature and sieved through a 2 mm mesh to remove stones and residual roots. The extraction of PAHs from soil and sediment samples was  (Table 4-1) was performed using gas chromatography coupled with mass spectrometry (Shimadzu QP2010S GC-MS).

Quality assurance/quality control (QA/QC) and statistical analysis
A five point calibration was used for the GC-MS with a detection limit of 10 ng/µl for PAH and D-PAH. Procedural blanks were run periodically to detect false positives. The average recoveries for deuterated standards ranged from 44.5% to 94% for acenaphthene-D10, and phenanthrene-D10 while it was between 51% to 132% for benz[a]anthracene-D12 and benzo[a]pyrene-D12.
All statistical analyses were carried out using R version 3.0.3 or SigmaPlot 11 (Systat Software Inc.).

Estimation of carcinogenic potency
Benzo[a]pyrene (BaP) is the only PAH which has enough toxicological data to estimate carcinogenic potency. The carcinogenic potency due to contamination of soil samples with PAHs can be estimated using the BaP equivalent concentrations of all PAHs using toxicity equivalency factors (TEFs) (  [22,23]. The BaP toxic equivalent concentration (TBaP eq ) for each site was calculated using equation 1.
Where, C i is the concentration of the individual PAH and TEF i is the corresponding toxicity equivalency factor. As a TEF value for Retene was unavailable, it was not considered for toxicity and cancer risk calculations.

Estimation of PAH accumulation in corn grains and dietary intake
Corn is a widely used staple of the local cuisine and was therefore selected to evaluate the potential risk associated with the intake of the food staple. The accumulation of PAH in corn grains were estimated based on the model developed by Paraíba et al 2010 [24]. The assumptions for developing the model are 1) in the soil-plant system PAH are degraded, dissipated, and transformed according to first order kinetics and 2) the uptake, transport, and accumulation of PAH in corn plants is via transpiration stream from contaminated soil solution. Equations 2 to 5 were used to estimate the concentration of PAH in corn and the definitions and the values of parameters used are listed in Table 4-2 Where the constants A and B represents the PAH total uptake rate by corn plants and the PAH total dissipation rate, both in the soil-plant system and are given by C w is the concentration of PAH in soil solution and is calculated using equation 5

Cancer health risk assessment
The lifetime average daily exposure (LADD) due to exposure to 16 PAHs through dietary uptake of corn, soil ingestion, and dermal contact was calculated for each exposure pathway according to the USEPA framework [44] using the equations 6 through 8 LADD corn ,LADD ing and LADD der are lifetime average daily doses associated with uptake of PAH by corn, ingestion of soil, and dermal exposure (mg/kg/d), C c and C s are concentrations of TBaP eq PAH in corn and soil (mg/kg), ET is exposure time (hrs/day), EF is exposure frequency (days/year), IR is ingestion rate (mg/kd/day), BW is body weight (kg), AT is average time (days), C f is conversion factor, ABS is dermal adsorption fraction (-), SA is surface area (cm 2 ) and AF d is soil to skin adherence factor (mg/cm 2 /event) The average incremental lifetime cancer risks (ILTCR i ) is cancer risk for an exposure route (i) and SF i = cancer slope factor for an exposure route (i).

Correlation among individual PAH and with soil organic carbon (TOC)
The relationship between individual PAHs and soil organic carbon (TOC) was examined using

Isomer ratios of PAHs
The PAH isomer pair ratios for Ant/(Ant+Phn), Fln/(Fln+Pyr), BaA/(BaA+Chr), and IP/(IP+BgP) can be used as tracers to identify sources of PAH contamination [1,37,38]. In the present study, the ratio of Ant/(Ant+Phn) was higher than 0.1 for all the sampling sites (Figure 4-4a). This indicates a dominance of combustion source rather than petroleum source [38]. and IP/ (IP+BgP) ratios greater than 0.5 which is indicative for biomass and/or coal combustion. [38,39]. In SMI, coal is not commonly used, but wood and other biomass are the predominant domestic fuels. Thus, combustion of wood and biomass likely is the primary sources of surface soil contamination.

Principal component analysis/Multiple linear regression
Principal component analysis (PCA) as a multivariate analytical tool has been widely used in environmental studies for source identification [1,2]. PCA reduces the total variability in the original data to a minimum number of factors. The factor loadings indicate the correlations of among the contaminants and related source emission composition [40]. The PCA was performed using R software R version 3.0.3. The principal components were identified by varimax rotation and Eigen values greater than one. The percentage mean contribution of each source was determined using multiple linear regression as described by [41]. PCA

Estimation of carcinogenic potency
The calculated TBaP eq for 18 town sample sites ranged from 9.0 µg/kg to 257. The TBaP eq concentrations in this study are higher compared to previous reported mean concentrations around the world. For example, TBaP eq in natural reserve soil samples in Italy were (18 µg/kg) [28], rural soils in Norway (14.3 µg/kg [30]), and soils of Tarragona, Spain (64 µg/kg) [29]. Our data are in range with agricultural soils in Delhi, India (45.6 µg/kg to 387.13 µg/kg, [1]. The suggest that concentrations of PAH in the surface soils of San Mateo Ixtatán are high and carcinogenic potency of PAH needs to be considered.

Estimation of PAH accumulation in corn grains and dietary intake
In SMI corn or maize is the primary crop cultivated at the sites where agricultural soil samples ∑16PAH uptake based on daily corn intake was set at 454 g for adults and 227g for children [47]. This is equivalent to an estimated PAH uptake of 2.26 µg/day for adults and 1.13 µg/day for children (below 6 years) ( Table 4-8).

Risk assessment
The aim of the risk assessment was to investigate the possibility of cancer development in adults and children (below 6 years) in San Mateo Ixtatán as a result of dietary uptake of PAH through corn consumption and exposure to PAHs in soil via ingestion and dermal exposure routes. The average incremental lifetime cancer risks (ILTCR) due to soil ingestion (ILTCR ing ) and dermal contact (ILTCR der ) are higher (figure 4-6) than US EPA acceptable cancer risk level of 1x10 -6 [25]. Children are at higher risk than adults with average ILTCR ing and ILTCR der at 1.61x10 -5 and 1.49x10 -5 respectively. The average incremental lifetime cancer risks due to dietary uptake (ILTCR corn ) of PAHs are 1.93x10 -3 for adults and 4.54x10 -3 for children. All the calculated ILTCRs exceed the US EPA acceptable cancer risk level of 1x10 -6 indicating high potential carcinogenic health risk [48]. Dietary uptake of PAH through corn consumption was the major contributor for total incremental lifetime cancer risk (ILTCR total ) accounting for >99% of ILTCR total . The individual PAH contributions to ILTCR corn are dominated by 2 and 3-ring PAHs (Table 4-7), which are associated with low temperature combustion sources, such as wood burning. From source apportionment it was established that more than 75 % of the PAH in agricultural soils have wood combustion as a source. Therefore it can be concluded that use of wood as domestic fuel is primary contributor to higher cancer risk for the people of SMI and should be taken into account for health protection.

Contamination of river sediments
The Ʃ17PAH concentrations at the starting of the river (SD1) and upstream of the town (SD2) PAHs in to the river, and household waste water discharged into the river are likely sources for higher sediment PAH concentrations at SD4 and SD5. However, it has not been studied which of these contamination sources might be more dominant. In comparison, the concentrations of PAH found in these sediments are higher than in sediments from Galveston Bay and the Mississippi River Delta [49]. They are in comparable range with sediment concentrations in Tonghui River of Beijing, China [50]. Finally, the ratios of Ant/ (Ant+Phn), Fln/(Fln+Pyr), BaA/(BaA+Chr), IP/(IP+BgP) suggest that most of the PAH contamination in river sediments is due to the combustion of wood and biomass (Table 4-9).

Conclusions
The data from this study suggest that in the rural town of San Mateo Ixtatán in the northwestern highlands of Guatemala the PAH contamination in surface soils is higher or are in range with urban and industrial regions around the world. The source apportionment concludes that the use of wood as domestic fuel is the primary source of this contamination in surface soils. The results of a cancer risk assessment indicate that both, children and adults face a high potential risk of cancer development as a result of the exposure to PAH via soil ingestion, dermal contact, and dietary uptake through corn cultivated in the contaminated soils. Alternative domestic fuel sources or improving existing stove models should be considered to reduce the health risk resulting from high PAH exposure.           This study demonstrated that cedar wood chips amended to have antimicrobial properties could be applied as a filter media in structural BMP systems, such as tree filters.

INTRODUCTION
Stormwater runoff mobilizes microorganisms into surface and ground water bodies,  Table 1).
Briefly, wood chips were submerged in TPA or AgNP solutions of 1.3 mmol/L ionic strength (NaCl) and were allowed to equilibrate for four to six days on a shaker to achieve a range of loadings that include the maximum possible loading of each antimicrobial (Table 1)

Experimental Methodology
The first part of the study involved batch experiments to test the inactivation kinetics of TPA and AgNP modified RC at 25°C and establishing whether exposure time or disinfectant loading had the main contribution on bacteria inactivation. This was followed by testing the disinfection performance of each material at 25°C and establishing bacteria inactivation efficiency at three different temperatures (25°C, 17.5 °C and 10°C) for the most effective materials. In the last part of the study, the amended materials were submerged in water to test for possible changes in bacteria inactivation efficiency over 120 day period.
During all experiments the solution ratio was 1 g of sorbent to 25 ml of solution containing the desired E. coli concentration. This ratio was based on a previous study that tested the performance of contaminant removal from an aqueous solution using the wood chips. All experiments were carried out in triplicates. Controls, buffer blanks, and NaCl blanks were analyzed for quality control purposes and to establish natural decay of bacteria over the experimental period. All solutions, glassware and other materials used for the experiments were sterilized through autoclaving.

Bacteria Culturing and Enumeration
Bacteria were cultured in LB broth (10 g/L Sodium Chloride, 10 g/L Tryptone, and 5 g/L Yeast

Inactivation Kinetics
The experiments were conducted with solutions containing E. coli at 10 6 CFU/100 ml and 10 5 The parameterized model (details in SI) was tested against the experimental data and a percentage difference was calculated to determine whether the model over or underestimated inactivation kinetics.

Inactivation Efficiency of Amended Red Cedar
To test the bacteria inactivation efficiency as a function of antimicrobial amendment loading, batch experiments with E. coli solutions at five different concentrations ranging from 10 2 to 10 6 CFU/100ml were prepared in 1.3 mM NaCl solution. These solutions were exposed to unmodified, TPA modified (loading range: 3.6 mg/g to a maximum lading of 9.3 mg/g TPA) and AgNP modified (0.33 mg/g to a maximum loading of 0.68 mg/g AgNP) red cedar (from now on test materials are abbreviated as shown in Table 1) at 25 °C in a rotisserie incubator set to 5 rpm (Major Science, Saratoga, CA, USA). The exposure time was 180 minutes. This time was based on a preliminary kinetic study in which a ~50% reduction of E. coli was measured when exposed to unmodified red cedar (UM) at 25 °C (data not shown). The inactivation efficiency of the materials was calculated as log 10 removal value (LRV): LRV = log 10 (C 0 )log 10 (C t ) Equation 3 Where C 0 is the initial E. coli concentration and C t is the final E. coli concentration, which is the sum of the E. coli that survived in both the aqueous phase and the E. coli attached to the unmodified or modified RC material.
To observe the differences in active E. coli present on the surface of unmodified and 6mg/g TPA modified red cedar, scanning electron microscope (SEM) images were taken. Each of the wood chips were put in contact with 10 6 CFU/100 ml for three hours in a rotary incubator at 25°C set to 5 rpm (Major Science, Saratoga, CA, USA) and prepared for SEM analysis (details in SI and Pathan et al. (2013)). Although all the materials were test in the same set of experiments, the inactivation efficiency data for unmodified red cedar, 6TPA and 0.6 Ag were published as part of a proof of concept paper (Kasaraneni et al. 2014) and the data for these three materials was supplemented with four additional amended materials to compare the whole spectrum of materials (Table 1)

Temperature Effect on Inactivation Performance
From the seven materials tested, the three best performing amended materials (6TPA, 9TPA, and 0.6AgNP; abbreviations explained in Table 1) along with UM were selected for testing possible effects of temperature on bacteria inactivation. The materials were exposed to an E. coli solution

Long-term storage of modified materials and its impact on inactivation efficiency
Tests were conducted to determine how long term storage impacts the inactivation performance of the TPA and AgNP modified materials. where LRV 0 is the initial LRV that was achieved by the freshly amended material, LRV Final is the final LRV achieved after the material has been stored for 120 days, k is the disinfection efficiency rate loss constant, and t is the time of storage.

Statistical Analysis
To compare the overall inactivation efficiency of the different materials for all bacteria concentrations tested, a one-way ANOVA was performed. To compare the inactivation kinetics of the different materials, an ANCOVA was carried out where LRV was the response variable and exposure time was the explanatory variable for each different type of red cedar treatment. The comparison between the LRV of the amended materials before and after long term storage was tested using a two-tailed t-test for each material. All statistical analyses had the alpha value set to 0.05 and were carried out using R version 3.0.3.

Inactivation Kinetics
The kinetic studies data were evaluated with the Chick-Watson model (Haas and Karra 1984) (Equation 1) to determine the concentration-time dependency necessary to achieve a 2 log 10 (99%) reduction of total culturable E. coli. In addition, the Chick-Watson model parameters differentiate whether the exposure time or antimicrobial loading had a greater impact on the reduction of the number of total culturable E. coli.
Over the 180 minute sampling period, the E. coli LRV increased with increased exposure time, following a generally log-linear trend for all materials (Figure 1). The order of calculated inactivation rates was as follows: 9TPA > 6TPA > 0.6AgNP > 0.3AgNP > 4TPA ~ 3TPA > UM.
Except for 4TPA and 3TPA, the inactivation rates of 9TPA, 6TPA, 0.3AgNP, and 0.6AgNP are statistically different from the unmodified red cedar inactivation kinetics, while 3TPA and 4TPA are similar to unmodified red cedar (  Figure S1) that pierces the bacteria cell membrane. Therefore, reduction of total number of culturable organisms by TPA relies heavily on the TPA loading, i.e. the more TPA molecules, the better the chance of contact (Kim, Kim and Rhee 2010).

Inactivation Efficiency of Amended Red Cedar
Comparing all seven amended materials demonstrates that highly modified materials (0.6AgNP, 6TPA, 9TPA) are overall significantly more effective at deactivating E. coli over a three hour time period compared to unmodified and low modified materials (0.3AgNP, 3TPA, 4TPA; Figure   1).
The LRV of unmodified RC was less than 1 and similar results were obtained using low loading on RC (3TPA, 4TPA and 0.3 AgNP; Table 1, Figure 2). At high loadings (6TPA, 9TPA and 0.6AgNP), E. coli reduction increased significantly (p<0.001, DF=6, F=32.8, ANOVA table in SI Table S1a&b) and the LRV was measured to be up to 3.71±0.38 (>99.9%; Figure 2). The LRV for TPA or AgNP modified RC was independent of the initial E. coli concentration, which ranged from 10 2 CFU/100 ml to 10 6 CFU/100 ml. Instead the LRV was influenced by the antimicrobial amendment loading on the material. The natural decay in Figure 2 depicts the natural die-off of bacteria in the control solution containing only NaCl and cannot be attributed to the antimicrobial.
The SEM images support these findings, as the unmodified red cedar shows much higher numbers of E. coli and denser clusters of live bacteria present (Figure 3 a&b) on the surface of the wood compared to 6TPA red cedar, which shows fewer bacteria (Figure 3 c&d). At higher magnification, the bacteria cells that are present on the 6TPA material are damaged and/or ruptured (Figure 3d), compared to those on the unmodified red cedar (Figure 3b). The structure of the bacteria cells in the SEM images from Figure 3 can be compared to those in Figure S2 in the As the more highly modified materials have a greater concentration of antimicrobial agent on the surface of the red cedar, the likelihood of the bacteria coming in contact with an antimicrobial agent during attachment were greater than for unmodified and low modified materials. As the temperature increased, there are significantly fewer bacteria that survive once they attach to the modified red cedar surface (p<0.001, DF=2, F-value=49.92, ANOVA table in Table S2b). This trend was clearly detected when comparing results at 17.5°C and 25°C (Figure 4b), whereas no difference in bacteria survival after attachment to the three amended materials was detected at 10°C. However, at 10°C survival after attachment for unmodified red cedar was two orders of magnitude higher compared to modified materials ( Figure 4b). These results are consistent with previously reported studies, where lower temperatures, such as those used in this study, allowed

Long-term reduction of the disinfection efficiency of modified materials
The three most effective sorbent materials (0.6AgNP, 6TPA, and 9TPA) were tested to quantify if and how saturated bench storage impacts the inactivation performance of the materials. These materials showed changes in efficiency over time that resulted in a 26% to 38% decrease in LRV (Table 3, Figure 5) compared to freshly prepared material. The decay in removal efficiency of these materials fell into the following order: 9TPA>0.6AgNP>6TPA, where there decrease in inactivation efficiency was significantly different from the original, freshly prepared inactivation efficiency ( Figure 5, Table S4). All models were fitted with r 2 > 97%.
Possible explanations for reduction in performance are: 1) leaching /desorption of TPA and AgNP during storage time, 2) fines clogging the pores of the RC making the antimicrobial inaccessible, 3) degradation of antimicrobial agent, 4) diffusion of TPA and AgNP in to RC thereby becoming inaccessible, and 5) surface aggregation and bridging among TPA polymer molecules or AgNP, depending on the material, making the antimicrobial less accessible. Even though all materials exhibit a decrease in inactivation performance over time (Table 3 and Figure   5), there was non-significant leaching of either TPA (<0.02%) or AgNP (0%) and Ag + ions (0%) from the RC sorbent, even after the full 120 days of storage. These findings are in agreement with To test whether eroded red cedar fines might have settled on the wood surface after storage and thus limited access to the antimicrobial coating, all materials were agitated by ultrasound for 10 minutes. Subsequently, the inactivation efficiency of all material was tested again and compared to the non-sonicated materials' performances. Because the LRV for the sonicated and nonsonicated materials was the same, covering the actives sites by wood-derived fines could be ruled out as a reason for decreased inactivation efficiency after storage. While aggregated/bridged TPA and AgNP amendments can explain the decrease in antimicrobial effectiveness, the overall antimicrobial effectiveness of the filter material is not entirely compromised as the amended materials still outperforms the UM red cedar by 3.45 times after 120 days of storage. Further, much of the decrease in inactivation performance occurs after 60 days of storage, whereas the inactivation performance remains essentially stable for the remainder of the time (days 60-120) ( Figure 5). Therefore, it can be postulated that the bridging/aggregating occurs within the first 60 days of storage and thereafter the TPA polymer and PVP-stabilized AgNP have reached a stable state of continued inactivation performance.

Implications and Benefits to Using Antimicrobial Filter Materials
Red cedar wood chips amended with either TPA or AgNP have been shown to increase the removal of bacteria from water. Even though similar results could be achieved with chlorine or Based on these properties, the red cedar wood chips amended with either TPA or AgNP could find use in stormwater treatment systems. An example is a Tree Filter BMP, which treats conventional stormwater pollutants, such as heavy metals or petroleum hydrocarbons, but could be outfitted with a layer of amended wood chips to provide additional antimicrobial treatment functions. Such a BMP system could potentially remove a much higher percentage of bacteria from stormwater runoff than most currently available treatment systems such as BactoLoxx (Filtrexx, Goffstown, NH) or Bacterra (Filterra Bioretention Designs, Ashland, VA). These currently available systems rely on sorption processes, pH changes, and natural predation by other organisms to achieve the reported removal of up to two log 10 units (99%) in E. coli, whereas up to 3-log 10 (99.9% removal) of E. coli can be achieved using the amended red cedar filter materials investigated herein. This removal far exceeds the 60% reduction in stormwater bacteria levels required in the more recently formulated stormwater treatment manuals (Rhode Island Department of Environmental Management 2010).
More importantly, our laboratory study showed that once the wood has been amended, neither TPA nor AgNP showed significant desorption from red cedar, i.e. no detectable leachate for 0.6AgNP and 6TPA, and less than 0.02% of the mass loaded onto 9TPA. Therefore, introducing these materials into BMPs should not disturb native microbial communities such as denitrifiers If the proposed amended red cedar were to be implemented as a filtration matrix in structural best management practices, the 6TPA modified red cedar would be favorable compared to the currently used matrix that consists of a shale-sand mix, or even unmodified red cedar. This is because not only does 1 gram of this material (i.e. 6 mg of TPA) achieve >2 log 10 removal after three hours of exposure at room temperature, but it also is effective at removing bacteria at decreased temperatures (17.5°C). These lower temperatures are closer to a natural BMP operating temperature in most temperate climates. However, introducing any type of bioactive material into the environment will require additional field studies to ensure that no unanticipated consequences result from the well-intended use of antimicrobial wood for stormwater treatment.

Limitations
The results in this study indicate that the inactivation performance of amended materials is significantly higher than unmodified materials. However, further testing of these materials under dynamic flow conditions is required to confirm these results. Stormwater is a complex mixture which can include organic contaminants, heavy metals, inorganic salts, humic substances etc.
along with several species of microorganisms. A study determining the impact of the complex stormwater chemistry conditions on the inactivation performance of the amended materials is necessary prior to use of these materials in the field.   Figure 2. Comparison of log 10 removal values (LRV) (average ± standard error) for the unmodified red cedar wood chips and the six modifications at different E. coli concentrations ranging from 10 2 to 10 6 CFU/100ml. All materials were exposed for 180 minutes. Natural decay data shows the fraction of E. coli die-off that is not due to antimicrobials or removal by attachment. In order to compare the performance of all the materials data for unmodified red cedar, 6TPA, and 0.6 AgNP was obtained from our previously published work 25 .  A B Figure 5. Average log 10 removal value ± standard error for the three most effective materials (6TPA, 9TPA, and 0.6Ag) compared to unmodified red cedar at 25°C. The LRV was determined immediately after modification and at 60, 90, and 120 days after being stored in water. An initial drop in the antimicrobial performance over the first 60 days was followed by a much slower decrease in inactivation performance afterwards.

Supplemental Text Chick-Watson Model Parameterization
To parameterize the model constants k' and n, a two-step approach is required as described in Tchobanoglous et al. 2003. (Tchobanoglous, Burton and Stensel ) Briefly, a linear regression is fitted to a plot of the natural-log normalized ratio ln(Nt/No) versus the experimental exposure time t for each antimicrobial material. The linear regression equation of each data set is utilized to solve for the theoretical time required to achieve a 99% removal of E. coli. Then, the theoretical time to achieve 99% removal of bacteria is plotted against the disinfectant concentration of the material and a linear regression equation for each type of disinfectant system is fitted to the calculated C. The negative inverse of the slope of that linear regression equation equals the coefficient of dilution n. The die-off constant k', is then calculated using Equation 2 where y is the y-intercept and Nt/No is the desired ratio of E. coli concentration at time t relative to the initial E. coli concentration, in this example 99%.
Sample Preparation for SEM analysis Briefly, the samples were fixed in 2.5% glutaraldehyde solution for 30 minutes. After fixation, the samples were washed three times with phosphate buffer and then dehydrated for 20 minutes using sequentially increasing ethanol solutions (increments of 10%) containing 40% to 100% ethanol. The samples were exposed to 100% ethanol three times before being chemically dried using Hexamethydisilazane (HMDS). The samples were transferred to a 1:2 HMDS:Ethanol solution for 20 minutes, then a 2:1 HMDS:Ethanol solution, and finally into 100% HMDS. The last step was repeated and the HMDS was left to evaporate in the fume hood overnight. The samples were sputter-coated with a thin layer of gold for 90 seconds under vacuum before the imaging was done in a SIGMA VP Field Emission-Scanning Electron Microscope (Zeiss, USA) with an acceleration voltage of 5kV and a SE2 detector.
Supplemental Tables  Table S1a -Performance -ANOVA Table. ANOVA table for one-way comparison of log10  removal values for the six modified red cedar wood chips and unmodified red cedar wood chips  using LRV as the response variable and material Table S2a -Attachment -ANOVA Table. ANOVA table for analysis of covariance comparing culturable E. coli after attachment for the three most effective materials (0.6 mg/g AgNP, 6 mg/g TPA, 9 mg/g TPA) compared to unmodified red cedar at three different exposure temperatures (10°C, 17.5°C and 25°C) with the final model being: E. coli removed by attachment ~ Temperature + Material. There is no interaction term in this model as it went through a stepwise regression and the final model as stated above had a lower AIC value compared to the initial model, which was attachment ~ Temperature * Material.
Degrees of Freedom  Figure S1. Structure of (A) 3-(trihydroxysilyl) propyldimethyloctadecyl ammonium chloride (TPA) and (B) 3-trimethoxysilyl propyldimethyloctadecyl ammonium chloride. Figure S2. SEM images of E. coli on a silicon wafers. In A) the E. coli at 104 CFU/100 ml were not exposed to TPA and depict live organisms, whereas in B) E. coli at 106 CFU/100ml were exposed to TPA for three hours before SEM imaging, thus depicting damaged organisms.