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Impact of chloride on denitrification potential in roadside wetlands
Environmental Pollution 212 (2016) 216e223
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Impact of chloride on denitrification potential in roadside wetlands* Nakita A. Lancaster a, Joseph T. Bushey a, Craig R. Tobias b, Bongkeun Song c, 1, Timothy M. Vadas a, d, * a
Department of Civil and Environmental Engineering, University of Connecticut, Storrs, CT 06269, USA Department of Marine Sciences, University of Connecticut Avery Point, Groton, CT 06340, USA c Department of Biology and Marine Biology, University of North Carolina Wilmington, Wilmington, NC 28403, USA d Center for Environmental Science and Engineering, University of Connecticut, Storrs, CT 06269, USA b
a r t i c l e i n f o
a b s t r a c t
Article history: Received 1 September 2015 Received in revised form 10 January 2016 Accepted 24 January 2016 Available online xxx
Developed landscapes are exposed to changes in hydrology and water chemistry that limit their ability to mitigate detrimental impacts to coastal water bodies, particularly those that result from stormwater runoff. The elevated level of impervious cover increases not only runoff but also contaminant loading of nutrients, metals, and road salt used for deicing to water bodies. Here we investigate the impact that road salt has on denitrification in roadside environments. Sediments were collected from a series of forested and roadside wetlands and acclimated with a range of Cl concentrations from 0 to 5000 mg L1 for 96 h. Denitrification rates were measured by the isotope pairing technique using 15NeNO 3 , while denitrifying community structures were compared using terminal restriction fragment length polymorphism (TRFLP) of nitrous oxide reductase genes (nosZ). Chloride significantly (p < 0.05) inhibited denitrification in forested wetlands at a Cl dosage of 2500 or 5000 mg L1, but the decrease in denitrification rates was less and not significant for the roadside wetlands historically exposed to elevated concentrations of Cl. The difference could not be attributed to other significant changes in conditions, such as DOC concentrations, N species concentrations, or pH levels. Denitrifying communities, as measured by T-RFs of the nosZ gene, in the roadside wetlands with elevated concentration of Cl were distinctly different and more diverse compared to forested wetlands, and also different in roadside wetlands after 96 h exposures to Cl. The shifts in denitrifying communities seem to minimize the decrease in denitrification rates in the wetlands previously exposed to Cl. As development results in more Cl use and exposure to a broad range of natural or manmade wetland structures, an understanding of the seasonal effect of Cl on denitrification processes in these systems would aid in design or mitigation of the effects on N removal rates. © 2016 Elsevier Ltd. All rights reserved.
Keywords: Denitrification Chloride Microbial communities nosZ T-RFLP
1. Introduction Anthropogenic activities have increased the nitrogen (N) loading in rural, suburban, urban, and mixed land use watersheds. One of the most important consequences of higher N loading is the increase in N exports from watersheds to coastal ecosystems. Excessive N exports to receiving water bodies can result in eutrophication, hypoxic conditions, and/or loss of biodiversity (Paerl,
*
This paper has been recommended for acceptance by Charles Wong. * Corresponding author. Glenbrook Rd., Unit 3037, Storrs, CT 06269, USA. E-mail address: [email protected] (T.M. Vadas). 1 Present address of B.K. Song: Department of Biological Sciences, Virginia Institute of Marine Science, Gloucester Point, VA 23062, USA. http://dx.doi.org/10.1016/j.envpol.2016.01.068 0269-7491/© 2016 Elsevier Ltd. All rights reserved.
1997). The input rate of N to coastal ecosystems has doubled since the 1940's, mostly due to agriculture and the burning of fossil fuels (Vitousek et al., 1997). As N inputs increase, the ability of freshwater systems to act as a buffer for N processing between the terrestrial landscape and coastal ecosystems is more important. In forested streams networks, for example, up to 95% of N inputs are retained in the watershed (Groffman et al., 2004). However, retention of N in suburban/urban watersheds has been observed to range from 25% to 95% (Groffman et al., 2004), and the reasons for the variability are not well understood. A number of hydrological, geomorphological, and biogeochemical changes in developed watersheds as well as elevated N loading may alter N removal. Changes in hydrology are driven by the increase in impervious cover in developed areas which generates high volumes of stormwater runoff flowing into streams where
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N removal and uptake is variable (Groffman et al., 2002; Paul and Meyer, 2001; Sweeney et al., 2004). At the interfaces, in wetlands and riparian zones which typically have higher organic matter and longer retention times, denitrification rates are often higher. There are many instances of distributed wetlands such as vernal pools in developed landscapes, often adjacent to paved areas that could be active sites for denitrification. Capps et al. (2014) found that vernal pools are hotspots of leaf litter turnover and denitrification, playing a significant role in nutrient dynamics across the forested landscape. However, the stormwater runoff that may enter these sites has elevated concentrations of nutrients and inorganic contaminants (Bannerman et al., 1993). High concentrations of N in particular can lead to biological saturation where the available N exceeds the processing ability of organisms (O'Brien et al., 2007). N processing is further impacted by increased pollutants associated with stormwater including metals (Sakadevan et al., 1999) or chloride from road salt (Green and Cresser, 2008; Green et al., 2008; Groffman et al., 1995; Hale and Groffman, 2006). Road salt usage is high in developed landscapes contributing to high concentrations of Cl in surrounding water bodies, the negative environmental impacts of which have been documented (Findlay and Kelly, 2011). As Cl concentrations increase in urban and rural environments, reaching levels as high as 4600 mg L1 (Kaushal et al., 2005), determining the impacts that the combined increase in N and Cl will have on N processing within watersheds is critical. Elevated Cl concentrations are known to have negative impacts on aquatic plants and animals (Kaushal et al., 2005) and cause changes to microbial community structure and respiration rates (Baldwin et al., 2006). The impact of Cl on N cycling has also been investigated; however, the results vary. In forest soils treated with 100 mg L1 Cl, nitrification and N mineralization rates decreased, but denitrification was not inhibited (Groffman et al., 1995). Sodium chloride exposure has also been observed to lead to a release of ammonia N, an increase in pH and an increase in n et al., rates of nitrification and mineralization of organic N (Ardo 2013; Compton and Church, 2011; Green and Cresser, 2008; Green et al., 2008). Denitrification potential decreased in freshwater stream sediments from forested watersheds exposed to 2500 mg L1, but was unchanged in stream sediments from suburban watersheds (Hale and Groffman, 2006). Upon exposure to much higher Cl levels, 2e10% salinity, sediment denitrification rates typically did not vary (Laverman et al., 2007; Marton et al., 2012). These studies independently used a wide variety of salinity levels with differing effects on denitrification, but the Hale and Groffman (2006) study in particular, suggests that microbial communities that are exposed to elevated concentrations of Cl for prolonged periods of time may adapt to the physiological stress associated with an increase in salinity, and thus observed no significant change in denitrification rates in suburban streams. We hypothesize that denitrification rates in distributed freshwater wetlands that are continuously exposed to Cl concentrations from road salt applications will not vary with increased exposure to Cl, while wetlands not previously exposed to Cl will have decreased rates. The purpose of this study was to determine how Cl addition impacted denitrification rates and denitrifying communities in sediments historically impacted by Cl relative to unimpacted sediments. The study compares roadside wetland areas historically exposed to elevated concentrations of Cl to unexposed forested wetland areas. In conjunction with varying Cl exposure, changes in water chemistry and denitrification rates were measured. In addition, denitrifying community structure in the soils were compared based on Terminal Restriction Fragment Length Polymorphism (T-RFLP) of the gene encoding nitrous oxide reductase (nosZ), which is an enzyme that reduces N2O to N2. Here, we
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evaluate the impact of Cl on denitrification rates, DOC, concentrations of N species, and denitrifying community structures to demonstrate the potential impact that road salt use will have on N removal and retention in developed landscapes. 2. Materials and methods 2.1. Sample collection Sediment slurries were collected from six wetland locations in Northeast Connecticut, within the Yale Myers Experimental Forest (YEF) (Fig. S1). The majority of YEF is undeveloped with some surrounding suburban areas, and with routinely salted local roads passing through the area. The soil in YEF and the surrounding area is classified as glacial till and the region is forested by a mix of hardwoods (YEF, 2009). This region was chosen to isolate the role of Cl used in road salt applications (a mixture of sand and NaCl) in the area and minimize the potential effect of other urban development impacts (e.g. other contaminants or heat). The specific sampling locations were selected based on observed Cl levels and proximity to roadways, with three of the locations within 10 m of a roadway and three located more than 200 m from a roadway, each with similar canopy cover and size. These particular wetlands are vernal pools and were a subset of paired wetlands from prior research in the area (Brady, 2012). Pools were paired based on leaf-off hemispherical photographs, had maximum depths at the center of typically about 70 cm, and areas of about 500e1500 m2 (Brady, 2013). The soils collected were from areas with no vegetation, though emergent vegetation was present in some parts of the pool. The roadside pools receive salt inputs through either surface runoff or salt spray deposition. Historically, the elevated conductivity in these pools persists through the spring and early summer. Sediment and water samples were collected during the snowfall season in November and January. All sampling events occurred after snowstorms during which road salt was applied. During the month of January, all of the wetlands were covered with a thick layer of ice necessitating the use of an axe to break through the ice. Sediments were collected using a 200 diameter sediment corer. Three 600 length cores were collected from each location, large leaf litter was removed and sediments were homogenized to form a composite sample. Grab samples of the surface waters were collected into acid-washed (3% HCl v/v) HDPE bottles. Sediments for Cl exposure experiments were stored at 4 C in the dark and used within 96 h. Water samples were filtered (0.45 mm) and frozen (20 C) in the dark prior to analysis of dissolved organic carbon (DOC), Cl, pH, alkalinity, and the concentration of different N species. 2.2. Chloride exposure microcosms Sediments samples were exposed to a range of Cl concentrations, diluted from NaCl stock, the most common road salt (Findlay and Kelly, 2011), for 96 h at a ratio of 20 g wet sediment to 40 mL salt solution. Salt exposure concentrations were 0, 500, 1000, 2500, or 5000 mg L1 Cl. After acclimating for 96 h, 2 g of wet sediment from each slurry was divided into a 12 mL Exetainer® tube (Labco, High Wycombe, United Kingdom), sealed with caps containing gastight septa and flushed with He gas according to published methods (Long et al., 2013). The vials were incubated overnight at 20 C to reduce the background concentrations of nitrate in the sample prior to laboratory assessment of the denitrification rates using isotope pairing. Sediment sub-samples for microbial community analysis were frozen at 80 C prior to extractions for denitrifying community analysis described below. Additional microcosms were established to measure changes in
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water chemistry parameters as a function of Cl exposure that might alter microbial processes. Microcosms were prepared in HDPE bottles by mixing 150 g of wet sediment with 300 mL of filtered site water. Sodium chloride was added to establish Cl concentrations of 0, 500, 1000, 2500, or 5000 mg L1, and samples were placed sideways on a shaker table rotating at 60 rpm for 96 h. The slurries were first centrifuged and the supernatant filtered (0.45 mm) prior to analysis for DOC, Cl, pH, alkalinity, and concentrations of different N species. 2.3. Water and soil chemistry analysis Filtered water samples from the wetlands and Cl microcosms were analyzed for DOC, Cl, pH, alkalinity, Total Nitrogen (TN), þ NO 3 þ NO2 , NH4 , and dissolved organic N concentrations via standard methods (APHA, 2000). The NO 3 þ NO2 data are referred to as NO3 in the results. DOC concentrations were analyzed via thermal oxidation (ASTM, 1994b) using an Apollo 9000 TOC analyzer (Teledyne Tekmar Company, Mason, OH). Chloride concentrations and pH were analyzed via ion-selective electrodes (Fisher Scientific Accumet probes, Pittsburgh, PA). Alkalinity was measured via titration to an endpoint of pH 4.3 (ASTM, 1994a). Nitrogen species concentrations were analyzed via a Lachat QuikChem 8500 Flow Injection Ion Auto Analyzer using Lachat Instruments (Loveland, CO) methods 10-107-04-1-A for nitrate þ nitrite, 12-107-06-1-A for ammonia, and 10-107-04-3-D for total N. Detection limits were 0.003, 0.003, and 0.008 mg L1 for ammonia, nitrate and total N, respectively. Percent organic matter was measured via loss on ignition at 450 C (Robertson et al., 1999). Total organic carbon and total nitrogen in soils were measured with a Perkin Elmer® 2400 CHN analyzer according to EPA Method 440. Soil pH was measured in a 1:1 (v/v) soil slurry according to EPA Method 9045D. Quality control was assessed via continuous calibration verification, lab duplicates and blanks, field duplicates and blanks, and matrix spikes. Recovery of spikes and quality control samples were within 10% of expected values. Differences in concentrations between forested and roadside sediments were evaluated using a student's t-test at the 95% confidence interval using Sigmaplot 13.0 software. 2.4. Denitrification rate analysis Denitrification and anammox rates were determined based on 30 the transformation of 15N-labeled NO N2 and 29N2, respec3 to tively, at 20 C (Thamdrup and Dalsgaard, 2002). This method yields a metabolic potential without N limitation that standardizes for NO 3 availability under anaerobic conditions (Long et al., 2013). After anoxic incubation under a He headspace to remove existing nitrate as stated above, the exetainers® were re-flushed with He for 14 5 min and spiked with 0.1 mL of 550 mM 15NeNO NeNHþ 3 and 4 using a He-flushed syringe. Following spiking, samples were vortexed and allowed to equilibrate for 30 min. A continuous flow isotope ratio mass spectrometer (IRMS; Thermo Delta V, Thermo Scientific, Waltham, MA) was programmed to collect gas samples from the vials four different times at 15 min intervals. Helium blanks were included at the beginning and end of each run to detect any potential leaks from the vials during the experiment. The peak area measurements for 30N2 and 29N2 were converted to mass using air calibration standards (Scotts Specialty Gases, Philadelphia, PA). Denitrification rates were determined from the slope of the linear regression of the time series data. In order to compare rates to other studies, rates on a per gram sediment basis were converted to a per area (m2) basis using a sediment bulk density of 2 g/cm3 and a conservative estimate of the denitrification zone thickness of 0.5 cm (David et al., 2006). Anammox rates were below detection in
these samples. Differences in rates between forested and roadside sediments were evaluated using a student's t-test at the 95% confidence interval using Sigmaplot 13.0 software. 2.5. Denitrifying community analysis DNA was extracted from 0.25 g of the Cl exposure microcosm sediments described above following a 96 h exposure. A PowerSoil DNA Isolation Kit (MO BIO Laboratories, Inc.) was used, following the manufacturer's protocol. The nosZ genes were amplified using the GoTaq Master mix (Promega, Fitchburg WI) with the primers of 6-FAM-labled nosZ1F and nosZ2R as described in Henry et al. (2006). The amplicons were purified using the Wizard SV Gel and PCR Clean-Up System (Promega, Fitchburg WI). DNA concentration of purified amplicons was measured using the Quant-iT ds DNA Assay Kit (Life Technologies, Carlsbad, CA). A total of 20 ng of PCR product was digested overnight at 37 C with 5 units of Cfo1 restriction endonuclease (Promega, Fitchburg WI). The digested products were precipitated with isopropenol and run on a 3130x/ Genetic Analyzer (Life Technologies, Carlsbad CA). Fragment analysis was conducted using the Gene Mapper 4.0 (Life Technologies, Carlsbad CA) and T-RFLP Analysis Expedited (T-REX) programs (Culman et al., 2009). Variations in nosZ gene fingerprints were assessed using a BrayeCurtis similarity matrix and cluster analysis in the Primer-5 software package (Primer-E Ltd, Lutton, UK). Sites CF and CR did not yield nosZ gene amplifications. 3. Results and discussion 3.1. Water and soil chemistry Dissolved organic carbon concentrations in samples collected during November were higher at all but one sampling site, but not significantly different than samples collected in January (Table 1). Forested wetlands had, on average, a DOC concentration of 11 mg L1 in the fall and 5.8 mg L1 in winter. The average DOC concentration in roadside wetlands were not significantly different between November and January, although site C had much higher concentrations of between 6.1 and 8.7 mg L1. Higher concentrations were expected during fall months since it is at the end of the increases in primary productivity generally observed during the summer and also includes fresh allocthonous inputs (Mann and Wetzel, 1995). With the exception of CR, the results presented here are consistent with the expected seasonal trend. CR was the only wetland that was not covered with ice during the January sampling event and the higher DOC concentration at that time might have been influenced by runoff, snowmelt, or the warmer wetland temperature (Westerhoff and Anning, 2000). The concentrations of Cl measured in the roadside wetlands were significantly higher (p < 0.05) than concentrations measured in forested wetlands (Table 1). On average the roadside wetlands were about 30-fold higher in Cl concentration compared to the forest wetlands, and concentrations were slightly lower in January. This was a unique year in that a heavy snowstorm hit New England at the end of October, just before sampling, resulting in the increase of road salt application causing high Cl concentrations prior to the typical winter season. The elevated concentrations of Cl in roadside wetlands are consistent with previous observed measurements which showed Cl concentrations ranging from 80e300 mg L1 in the roadside wetlands and 2e6 mg L1 in forested wetlands during late spring and summer months that helped identify potential sampling locations. Although observed concentrations were lower in the spring and summer due to an increase in rainfall and the termination of road salt application, roadside wetlands still held elevated concentrations throughout
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Table 1 Average water quality dissolved concentrations measured for forested and roadside wetlands in November and January. All concentrations are in mg L1 ± 1 SD. DOC e dissolved organic carbon; Cl e chloride; TN e total nitrogen (N); NO3 e total nitrate plus nitrite; NH4 e total ammonia species; DON e dissolved organic N (calculated by difference).
November January
mg L1
DOC
Forested Roadside Forested Roadside
11 3.8 5.8 3.7
± ± ± ±
Cl 4.2 2.0 4.0 4.3
NO3eN
TN
160 4800 110 3600
± ± ± ±
37a 550b 18a 2200b
0.26 0.10 0.27 0.25
± ± ± ±
0.052a 0.06b 0.20ac 0.27bc
0.008 0.006 0.027 0.036
± ± ± ±
NH4eN 0.004ac 0.003a 0.015bc 0.015b
0.019 0.016 0.11 0.14
± ± ± ±
DON 0.001 0.017 0.11 0.20
0.24 0.078 0.13 0.073
± ± ± ±
0.06a 0.045b 0.10ac 0.087bc
A difference in the letter after the value indicates significance at p < 0.05 between water quality parameters.
the year compared to forested wetlands. It is not uncommon for Cl concentrations to remain elevated during the year due to storage in the watershed and aquifers (Findlay and Kelly, 2011). Concentrations of different N species were all less than 0.3 mg L1 and DON was in the highest proportion (Table 1). Only a few comparisons showed significant differences (p < 0.05), including the roadside and forested wetland concentrations of TN and DON in November, and NO3 concentrations between November and January. The concentration of NO3 in both sets of wetlands was higher during the January sampling event. Typically roadside environments have higher concentrations of NH4 and NO3 compared to forested environments due to near source mobile emissions within 10 m of the roadway (Bettez et al., 2013), but they were not significantly different in this case. The significantly higher DON in the forested wetland compared to the roadside wetland during November likely reflects a combination of increased microbial activity and leaf litter input in the interior of the forest. By January, a larger portion of TN was in the form of NH4, possibly from the mineralization of biomass during a relatively mild winter, but there was no significant difference between forested and roadside wetlands. The soils used in this study had varying organic matter content, but a similar and relatively low C:N ratio (Table 2). There was not a clear trend among the sites, i.e. not all roadside ponds had the lowest organic matter, and TOC and TN were actually not significantly different between any sites. The C:N ratio was relatively low, less than 17 at all sites, and that is below the level that is considered N limiting in the literature (Palta et al., 2014). Soil pH was over 5 in both roadside and forested wetlands, but slightly lower on average in the roadside wetlands. Acidification has been observed in watersheds exposed to salt solutions and could be due to cation exchange at those sites (Wright et al., 1988). In the wetlands that were selected for this study, excess N inputs are limited to atmospheric deposition and stormwater runoff from impervious surfaces. Direct atmospheric deposition likely influenced the forested wetlands in the study area to the same extent that it impacted the roadside wetlands since wetlands were selected based on similar canopy cover. Runoff from impervious surfaces was limited to the adjacent roadway and did not appear to carry higher concentrations of NO3 into the roadside wetlands. While the measured concentrations reflect the winter season when
biological activity is greatly reduced, concentrations in roadside wetlands here or in similar developed landscapes in the spring or summer are expected to increase from higher primary productivity or fertilizer application. Overall, our initial assessment of aqueous water quality parameters indicates that the most distinct difference between the forested and roadside sites was Cl concentrations. Chloride exposure microcosms were established to assess the difference in denitrification rates in previously exposed (roadside) or unexposed (forested) sediments at a range of Cl concentrations. 3.2. Denitrification rates The rate of denitrification decreased sharply at a Cl dosage of 2500 mg L1 and above in both the forested and roadside wetlands, but was only significantly different (p < 0.05) from the control (zero Cl dose) in the forested wetlands (Fig. 1). The average denitrification rates were similar and not significantly different between the sites at 1000 mg L1 and below and were similar to those typically observed in freshwater sediments. Aerial rates in these wetlands ranged from 14 to 53 mmol N/m2/hr, which is near the low end of the rates observed in the sediments of freshwater lakes (Pina-Ochoa and Alvarez-Cobelas, 2006). Some studies have observed stimulatory effects of increasing salinity on denitrification rates in tidal forested wetlands (Marton et al., 2012), but that was not observed here. At a Cl dosage of 2500 mg L1, the rate of denitrification dropped significantly (p < 0.05) by 99% in the forested wetlands. Rates in roadside wetlands dropped by 90%, but the decrease was not statistically significant. At 5000 mg L1 the rates decreased slightly more, by 99% and 94% for the forested and roadside wetlands, respectively. Cl has been demonstrated to negatively impact denitrification rates in an estuary, with rates
Table 2 Soil characteristics measured for forested and roadside wetlands. Site
Organic matter (%)
AF AR BF BR CF CR
27.5 3.4 14.0 10.4 9.9 33.3
± ± ± ± ± ±
11.7 0.1a 2.7 1.5b 5.1 8.6b
TOC (%) 13.7 1.6 4.6 2.5 4.2 12.5
± ± ± ± ± ±
6.9 0.6 1.5 0.5 0.6 3.6
TN (%) 0.88 0.14 0.34 0.2 0.33 0.76
± ± ± ± ± ±
C:N ratio 0.42 0.04 0.14 0.04 0.06 0.25
15.5 11.6 13.9 12.9 12.7 16.7
± ± ± ± ± ±
0.4a 0.5b 1.4acd 0.2bc 0.5d 0.7e
Soil pH 5.57 5.29 5.53 5.23 5.62 5.45
A difference in the letter after the value indicates significance at p < 0.05 between parameters.
Fig. 1. Average denitrification rates for forested and roadside wetlands as a function of Cl dosage to microcosms. Error bars reflect the standard deviation of three replicate wetlands. A * above the data indicates a significant difference between roadside and forested wetlands (p < 0.05).
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measured at the more saline end of a salinity gradient being lower relative to those at the freshwater end (Fear et al., 2005; Tobias et al., 2001), but this could be a function of DOC availability (Wallenstein et al., 2006) or other water quality parameters as well. In order to assess any impact water quality parameters might have had on denitrification rates, we also assessed the changes in these same parameters during Cl dosage experiments. While the average DOC concentrations did decrease upon Cl exposure, the decreases were not significantly different (p < 0.05) from the control (Fig. 2). There was only one significant difference in DOC observed between forested and roadside wetlands at an exposure concentration of 5000 mg L1 Cl. There was a general trend of decreasing DOC concentrations as chloride dosage increased, from starting concentrations of 14 or 5 mg L1 in the forested and roadside wetlands, respectively, to around 6 and 2.3 mg L1, respectively, at 5000 mg L1 Cl, though the roadside wetlands always had lower average DOC concentrations compared to the forested wetlands. This was at least in part due to the fact that at least two of the roadside wetland soils had lower C content relative to forested wetland soils (Table 2). Fear et al. (2005) also measured lower organic matter at the more saline end of the salinity gradient making a determination of the relative significance of Cl and DOC on the denitrification rate difficult. One possible reason for the loss of DOC is coagulation and settling or filtration induced by the addition of NaCl, although Naþ is not known to be a particularly effective coagulant (Stumm and Morgan, 1981). Pearson correlation coefficients do not suggest a significant correlation between DOC concentrations and Cl1 dosages (Table S1) and even if the Naþ did have an effect on the reduction in DOC, it is uncertain why there was a significant difference in DOC at 5000 mg L1 when both sediments were exposed to the same Cl concentration. Although the average DOC concentration decreased by nearly half in both sets of wetlands after the addition of 500 mg L1 Cl, they remained largely unchanged as higher doses of Cl were added. Despite the decrease in DOC concentration with the addition of 500 mg L1 Cl, which could potentially produce a corresponding decrease in denitrification rates due to carbon limitation, denitrification rates did not significantly decrease. In contrast, denitrification rates only significantly declined at 1000 mg L1 and above in forested wetlands as the concentration of Cl increased suggesting that the decrease was a function of Cl concentration and not DOC.
Fig. 2. Dissolved organic carbon (DOC) concentrations from filtered 96 h incubations relative to measured Cl dosage. Error bars reflect the standard deviation across three wetlands. A * above the data indicates a significant difference between roadside and forested wetlands (p < 0.05).
The alkalinity and pH both decreased as Cl concentration increased. Similar to the pattern observed in DOC concentrations, the largest drop in pH occurred between 0 and 500 mg L1 Cl dosage and the pH continued to decrease in smaller increments as the Cl concentration increased. Initial pH values for both sets of wetlands were 5.3e5.6 and decreased to 4.3 at 5000 mg L1 Cl dosage (Table 3). Alkalinity was lower in the roadside wetlands with a base level of 0.80 meq/L decreasing to 0.19 meq/L with Cl addition. The change in alkalinity was greater in the forested wetlands with a base level of 1.17 meq/L and decreased to 0.17 meq/ L with Cl addition. The decrease in pH and alkalinity likely resulted from the increase in NHþ 4 in the mesocosms as Cl concentrations increased (Table 3). A drop in pH has been observed before in runoff from soils exposed to salt water due to cation exchange (Wright et al., 1988). Decreases in pH have been shown to decrease denitrification rates (Baeseman et al., 2006). The optimal pH range for denitrification to occur is between 6 and 8. In the current study, the pH was below 6 and the pH values decreased in a similar pattern for both sets of wetlands. Since the denitrification rates exhibit contrasting patterns, especially at the lower concentrations of Cl, it appears that the decrease in pH was not as influential as the increase in Cl, a fact supported by the continued decrease in denitrification rate above the 1000 mg L1 dosage without further decreases in pH. The most dramatic differences observed in N species concenþ trations were for NHþ 4 where NH4 concentrations increased significantly (p < 0.05) in both forested and roadside wetlands from 0 to 5000 mg L1 Cl, increasing NHþ 4 in solution from about 0.5 or less mg L1 to 2.5 and 1.3 mg L1 in forested and roadside wetlands, respectively (Table 3). The addition of Naþ likely exchanges NHþ 4 on the sediments, a result that is consistent with other studies (Compton and Church, 2011; Green and Cresser, 2008; Green et al., 2008; Laverman et al., 2007). This increase in NHþ 4 could stimulate nitrification in some sediment systems (Baldwin et al., 2006), but we did not observe a corresponding increase in NO 3 concentration, and actually saw a decrease in NO 3 which likely contributed to the decrease in observed denitrification rates. Others have also observed a short term decrease in NO3 after incubation with NaCl over one day (Compton and Church, 2011). While TN increased since the majority of it consisted of NH4, DON was reduced in all cases, though not statistically significant. This is consistent with the reduction in concentration of DOC. Although there was not a statistically significant decrease in NO 3 , the average concentration went from 0.23 mg L1 in the forested wetlands at 500 mg L1 Cl dosage to 0.01 mg L1 at 5000 mg L1 Cl, which is about the same percentage drop in denitrification rate that was observed. Only at the dosage of 2500 mg L1 Cl was the difference between forested and roadside sediments significant (p < 0.05) with an 85% reduction in denitrification rate in the forested wetlands, suggesting a possible acclimation of the roadside sediments to Cl exposure. Hale and Groffman (2006) showed a similar trend while measuring denitrifying enzyme activity (DEA) in sediments that were exposed to Cl concentrations of 2500 mg L1 prior to the laboratory incubations and sediments that had not been previously exposed to high Cl concentrations. However, their findings (Hale and Groffman, 2006) showed no change in DEA in sediments incubated with 2500 mg L1 of Cl that had been previously exposed to Cl (i.e. the urban/suburban stream) while we observed a significant drop in denitrification rate. Chloride concentrations at the stream site they used had been recorded as high as 4600 mg L1 in the winter but were closer to 100 mg L1 at the time the sediments were collected in the summer, at which time the microbial population may have been different. Marton et al. (2012) also reported no change in denitrification potential at most locations after incubating sediments with 2000 or 5000 mg L1 Cl. At the time of
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Table 3 Nitrogen species concentrations in filtered samples for roadside and forested wetlands after 96 h exposure to different Cl concentrations. NH4, NO3, TN, and DON represent total ammonia, total oxidized nitrogen, total nitrogen and dissolved organic N species, respectively. Cl dosage (mg L1) 0
500
Forest NH4 NO3 TN DON pH
0.57 0.26 1.82 0.99 5.57
± ± ± ± ±
Roadside 0.32a,c 0.20a,c 0.79a,c 0.50a,c 0.05a,c
0.42 0.17 0.97 0.37 5.32
± ± ± ± ±
0.34a,e 0.17a,c 0.83a,c 0.33a,c 0.11b,f
5000
Forest 1.48 0.23 2.18 0.46 4.90
± ± ± ± ±
Roadside 0.76a,cd 0.15a,c 1.21a,c 0.36a,c 0.16a,d
0.56 0.16 0.96 0.24 4.83
± ± ± ± ±
0.32a,e 0.08a,c 0.61a,c 0.23a,c 0.21a,g
Forest 2.53 0.01 2.84 0.30 4.3
± ± ± ± ±
Roadside 0.55a,d 0.01a,c 0.78a,c 0.24a,c 0.04a,e
1.27 0.02 1.22 0.12 4.27
± ± ± ± ±
0.14b,f 0.03a,c 0.56b,c 0.11a,c 0.51a,g
A difference in the first letter after the value indicates significance at p < 0.05 between forest and roadside pairs at the same Cl dosage, while a difference in the second letter after the value indicates significance at p < 0.05 between Cl dosages at the same location.
sample collection (Marton et al., 2012), river water and pore water were freshwater and there was no historic data to indicate if the sediments had been exposed to Cl previously. In our study, Cl concentrations were elevated for the year at the time of sample collection. Our denitrification rate data suggests there was an adaptation due to long-term Cl exposure in these study sites, and perhaps there was also a threshold to a significant decrease in denitrification activity between 1000 and 2500 mg L1 Cl dosage that was also exceeded in the current study. 3.3. Comparison of denitrifying communities PCR of nosZ genes was successful for most samples except for the CF and CR samples. Low quality and quantity of extracted DNA may inhibit amplification of nosZ genes. With the nosZ gene detected in the other soil communities, T-RFLP patterns demonstrated distinct differences in the nosZ gene carrying denitrifying communities in the forested wetlands relative to the roadside wetlands, and they tended to group together. The cluster analysis of the nosZ T-RFLP data (Fig. S2) showed that denitrification assemblages from forested sites (AF and BF) grouped together, with a greater than 90% similarity, as did the roadside wetlands (AR and BR) with about 90% similarity. In addition, the AR and BR sites amended with Cl were only about 75 and 70% similar, respectively, suggesting different nosZ gene pools were expressed upon Cl exposure. In order to examine further details of denitrifying communities, the relative percent abundance of each terminal restriction fragment (T-RF), each nominally representing a different genetic variant, was calculated based on the heights of T-RFs observed (Fig. 3). There were clear differences in populations in forested versus roadside wetlands, with some T-RFs showing up
only in forested sites, e.g. T-RFs with 53, 323 and 402 bp at the AF site (Fig. 3A). In the BR roadside sites, the denitrifying populations corresponding to the T-RFs with 51, 255 and 320 bp disappeared, while the T-RFs with 317, 341, 348 and 445 bp appeared or increased in the roadside wetlands treated with elevated Cl concentration (Fig. 3B). In the AR roadside wetland, T-RFs with 320, 445 and 473 bp decreased or disappeared while T-RFs with 317 and 336 increased with the Cl exposure (Fig. 3B). Thus, T-RFLP fingerprints clearly demonstrated that there were shifts in nosZ gene carrying denitrifying organisms, corresponding both to a difference in the forested and roadside wetlands, as well as in response to the elevated Cl exposure in the roadside wetland sediments. 3.4. Other factors affecting denitrification Correlation matrices (Table S1) for each set of wetlands combined with the chemical trend data suggest that Cl concentration is the primary driver of denitrification inhibition in these wetland sediments. Denitrification rates evaluated across individual timespecific wetland data were significantly (p < 0.05) negatively correlated with Cl concentrations for both the forested and roadside wetlands (R ¼ 0.64 and 0.69, respectively). Other water chemistry factors that could affect denitrification were positively correlated (p < 0.05) for the forested (pH, R ¼ 0.81) and roadside (DOC, R ¼ 0.60) wetlands. However, the relationship of pH to denitrification rates likely is strongly influenced by the cocorrelation of Cl addition on pH due to reduction in pH during exposure. Similarly, while DOC was correlated significantly (p > 0.05) with denitrification for the forested wetlands, the relationship was not significant when evaluated across all data. Additionally, the DOC concentrations and pH levels in the two sets of
Fig. 3. Comparison of denitrifying communities based on selected nosZ gene T-RFs in the A) sediments collected from the forested wetlands (AF and BF) and roadside wetlands (BR and AR), and B) in the wetland sediments with 5000 mg L1 Cl concentration treatments (BR þ Cl and AR þ Cl) or without treatments (BR and AR). Relative abundance of each T-RF was calculated based on the height of T-RFs in each fingerprint.
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microcosms contradict the trends that would be expected if these chemical shifts were impacting denitrification rates. Although we chose roadsides sites in less developed areas, it is likely that trace metal concentrations were higher in roadside wetlands than forested wetlands. Increasing concentrations of NaCl further in the mesocosms would likely have resulted in additional €ckstro € ma et al., 2004; trace metal release from the sediments (Ba Lofgren, 2001). However, if metals were responsible for the inhibition of denitrification, which has been observed in some cases (Sakadevan et al., 1999), there likely would have been a more significant decrease in denitrification with the addition of Cl to roadside wetlands compared to forested wetlands where the concentrations of metals were likely lower. The fact that denitrification rates in roadside wetlands did not decrease as much as those in the forested wetlands at higher Cl additions again supports Cl as the primary factor impacting denitrification rates. Molecular analysis suggests that the denitrifying community structure differs between the roadside and forested wetlands. The long-term stressors from roadside disturbance may lead to a shift in denitrifying community structures. Our T-RFLP analysis is limited based on the identification of only nosZ genes and the measurement of N2 as the final product. Future studies could examine N2O production as well as the complete denitrification gene pathway to get a more complete picture of changes in the denitrifying community. There is no data in the literature showing shifts in denitrifier communities in response to Cl exposure, but Cl has been shown to impact methane production and resulted in significant changes in T-RFLP of archeal genes in freshwater wetlands exposed to NaCl (Baldwin et al., 2006). Regardless, the higher denitrification rates in the roadside wetlands with Cl addition suggest that nosZ containing denitrifying communities in the roadside wetlands are able to accommodate the exposure to additional Cl better than the forested wetlands. Given that the roadside wetlands historically have been exposed to elevated concentrations of Cl, we hypothesize that the denitrifying community may have adapted to the presence of Cl. In comparison to other studies, the results generated here support the hypothesis that historic exposure results in microbial function adapted to Cl (Hale and Groffman, 2006; Marton et al., 2012). However, denitrification rates measured in the roadside wetlands only significantly decreased at the highest concentrations of Cl added (2500 and 5000 mg L1), suggesting that even with historic exposure to Cl, a threshold exists beyond which denitrification rates are severely impacted. Our study sites differed from those of prior studies in having a longer retention time (i.e. increased time for adaptation to occur) and in being hydrologically isolated (i.e. lower chance of recruitment). In combination with the input concentrations or nutrients as well as stressors, these differences are critical for determining the real impacts of road salt on N removal in distributed wetlands near developed landscapes. Similar reductions in N processing could be observed in manmade structures, such as dry ponds, wet ponds, or treatment wetlands meant to capture and reduce both volumes of flow and contaminant concentrations. While it is expected that wet ponds and wetlands, in particular, can provide a hot spot for N removal (EPA, 2012), their performance in practice over the long-term and the variation seasonally is unknown. Given the likelihood of Cl exposure to both natural and manmade wet ecosystems during winter storm events, as well as the lingering of Cl in the ecosystem through the spring, it is possible that denitrification rates are significantly reduced during winter and spring when significant levels of NO 3 come through the system from spring thaw and the start of the growing season. Cl concentrations have been measured as high as 25,000 mg L1 in snow melt (Makepeace et al., 1995) which is well above the concentration after which we saw a
greater than 90% reduction in denitrification rates. In these particular wetlands, elevated conductivity levels were observed in the roadside wetlands in the following spring and summer. Whereas the conductivity in forested wetlands remained less than 30 mS/cm, the roadside wetlands remained typically above 400 mS/ cm, and in the case of site CR, remained above 1000 mS/cm until June. While these are only conductivity values, and not Cl, it is reasonable to assume Cl is significantly contributing to the increased conductivity, especially because these are vernal pools with little or no flow. It is possible that in more developed areas where salt use may be higher or more frequent, or in wetlands with little or no outflow, Cl concentrations could persist at levels shown to dramatically decrease denitrification rates. Our microbial data suggests distinct differences in the denitrifying populations between both previously exposed roadside wetlands and unexposed forested wetlands, as well as upon Cl exposure. Future studies could focus on a more detailed analysis of denitrifying communities, encompassing the full denitrification pathway genes using more advanced methods to answer questions regarding the ability of denitrifying communities to adapt to increases in salinity. Determining the threshold at which denitrifying communities can no longer denitrify and the timing required for adaptation to occur could aid in either setting limitations on Cl exposure in certain productive wetland areas, or in the design of functional stormwater capture systems for N removal to protect coastal ecosystems from the detrimental impacts of excess N loading. Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.envpol.2016.01.068. References APHA, 2000. Standard Methods for the Examination of Water and Wastewater, 15th ed. American Public Health Association, Washington, DC. n, M., Morse, J.L., Colman, B.P., Bernhardt, E.S., 2013. Drought-induced saltArdo water incursion leads to increased wetland nitrogen export. Glob. Change Biol. 19, 2976e2985. ASTM, 1994a. Method 2320B: Carbonate, Biocarbonate and Total Alkalinity. American Society of Testing & Materials, Philadelphia, PA. ASTM, 1994b. Method D4129-88 Standard Test Method for Total and Organic Carbon in Water by High Temperature Oxidation and by Coulometric Detection. American Society of Testing & Materials, Philadelphia, PA. €ckstro €ma, M., Karlssona, S., B€ Ba ackmanb, L., Folkesonb, L., Lindc, B., 2004. Mobilisation of heavy metals by deicing salts in a roadside environment. Water Res. 38, 720e732. Baeseman, J.L., Smith, R.L., Silverstein, J., 2006. Denitrification potential in stream sediments impacted by acid mine drainage: effects of pH, various electron donors, and iron. Microb. Ecol. 51, 232e241. Baldwin, D.S., Rees, G.N., Mitchell, A.M., Watson, G., Williams, J., 2006. The shortterm effects of salinization on anaerobic nutrient cycling and microbial community structure in sediment from a freshwater wetland. Wetlands 26, 455e464. Bannerman, R.T., Owens, D.W., Dodds, R.B., Hornewer, N.J., 1993. Sources of pollutants in wisconsin stormwater. Water Sci. Technol. 28, 241e259. Bettez, N.D., Marino, R., Howarth, R.W., Davidson, E.A., 2013. Roads as nitrogen deposition hot spots. Biogeochemistry 114, 149e163. Brady, S., 2012. Road to evolution? Local adaptation to road adjacency in an amphibian (Ambystoma maculatum). Sci. Rep. 2, 235. Brady, S.P., 2013. Microgeographic maladaptive performance and deme depression in response to roads and runoff. PeerJ 1. Capps, K.A., Rancatti, R., Tomczyk, N., Parr, T.B., Calhoun, A.J.K., Hunter, M., 2014. Biogeochemical hotspots in forested landscapes: the role of vernal pools in denitrification and organic matter processing. Ecosystems 17, 1455e1468. Compton, J.E., Church, M.R., 2011. Salt additions alter short-term nitrogen and carbon mobilization in a coastal Oregon Andisol. J. Environ. Qual. 40, 1601e1606. Culman, S.W., Bukowski, R., Gauch, H.G., Cadillo-Quiroz, H., Buckley, D.H., 2009. TREX: software for the processing and analysis of T-RFLP data. BMC Bioinform. 10. David, M.B., Wall, L.G., Royer, T.V., Tank, J.L., 2006. Denitrification and the nitrogen budget of a reservoir in an agricultural landscape. Ecol. Appl. 16, 2177e2190.
N.A. Lancaster et al. / Environmental Pollution 212 (2016) 216e223 (EPA), Environmental Protection Agency, 2012. Water: Polluted Runoff. Fear, J.M., Thompson, S.P., Gallo, T.E., Paerl, H.W., 2005. Denitrification rates measured along a salinity gradient in the eutrophic Neuse River Estuary, North Carolina, USA. Estuaries 28, 608e619. Findlay, S.E.G., Kelly, V.R., 2011. Emerging indirect and long-term road salt effects on ecosystems. In: Ostfeld, R.S., Schlesinger, W.H. (Eds.), Year in Ecology and Conservation Biology. Blackwell Science Publ, Oxford, pp. 58e68. Green, S.M., Cresser, M.S., 2008. Nitrogen cycle disruption through the application of de-icing salts on upland highways. Water Air Soil Pollut. 188, 139e153. Green, S.M., Machin, R., Cresser, M.S., 2008. Effect of long-term changes in soil chemistry induced by road salt applications on N-transformations in roadside soils. Environ. Pollut. 152, 20e31. Groffman, P.M., Gold, A.J., Howard, G., 1995. Hydrologic tracer effects on soil microbial activities. Soil Sci. Soc. Am. J. 59, 478e481. Groffman, P.M., Boulware, N.J., Zipperer, W.C., Pouyat, R.V., Band, L.E., Colosimo, M.F., 2002. Soil nitrogen cycle processes in urban Riparian zones. Environ. Sci. Technol. 36, 4547e4552. Groffman, P.M., Law, N.L., Belt, K.T., Band, L.E., Fisher, G.T., 2004. Nitrogen fluxes and retention in urban watershed ecosystems. Ecosystems 7, 393e403. Hale, R.L., Groffman, P.M., 2006. Chloride effects on nitrogen dynamics in forested and suburban stream debris dams. J. Environ. Qual. 35, 2425e2432. Henry, S., Bru, D., Stres, B., Hallet, S., Philippot, L., 2006. Quantitative detection of the nosZ gene, encoding nitrous oxide reductase, and comparison of the abundances of 16S rRNA, narG, nirK, and nosZ genes in soils. Appl. Environ. Microbiol. 72, 5181e5189. Kaushal, S.S., Groffman, P.M., Likens, G.E., Belt, K.T., Stack, W.P., Kelly, V.R., Band, L.E., Fisher, G.T., 2005. Increased salinization of fresh water in the northeastern United States. Proc. Natl. Acad. Sci. U. S. A. 102, 13517e13520. Laverman, A.M., Canavan, R.W., Slomp, C.P., Van Cappellen, P., 2007. Potential nitrate removal in a coastal freshwater sediment (Haringvliet Lake, The Netherlands) and response to salinization. Water Res. 41, 3061e3068. Lofgren, S., 2001. The chemical effects of deicing salt on soil and stream water of five catchments in southeast Sweden. Water Air Soil Pollut. 130, 863e868. Long, A., Heltman, J., Tobias, C., Philips, R., Song, B., 2013. Co-occurring anammox, denitrification, and codenitrification in agricultural soils. Appl. Environ. Microbiol. 79, 168e176. Makepeace, D.K., Smith, D.W., Stanley, S.J., 1995. Urban stormwater quality - summary of contaminant data. Crit. Rev. Environ. Sci. Technol. 25, 93e139. Mann, C.J., Wetzel, R.G., 1995. Dissolved organic carbon and its utilization in a riverine wetland ecosystem. Biogeochemistry 31, 99e120. Marton, J.M., Herbert, E.R., Craft, C.B., 2012. Effects of salinity on denitrification and greenhouse gas production from laboratory-incubated tidal forest soils. Wetlands 32, 347e357. O'Brien, J.M., Dodds, W.K., Wilson, K.C., Murdock, J.N., Eichmiller, J., 2007. The
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Environmental Pollution journal homepage: www.elsevier.com/locate/envpol
Impact of chloride on denitrification potential in roadside wetlands* Nakita A. Lancaster a, Joseph T. Bushey a, Craig R. Tobias b, Bongkeun Song c, 1, Timothy M. Vadas a, d, * a
Department of Civil and Environmental Engineering, University of Connecticut, Storrs, CT 06269, USA Department of Marine Sciences, University of Connecticut Avery Point, Groton, CT 06340, USA c Department of Biology and Marine Biology, University of North Carolina Wilmington, Wilmington, NC 28403, USA d Center for Environmental Science and Engineering, University of Connecticut, Storrs, CT 06269, USA b
a r t i c l e i n f o
a b s t r a c t
Article history: Received 1 September 2015 Received in revised form 10 January 2016 Accepted 24 January 2016 Available online xxx
Developed landscapes are exposed to changes in hydrology and water chemistry that limit their ability to mitigate detrimental impacts to coastal water bodies, particularly those that result from stormwater runoff. The elevated level of impervious cover increases not only runoff but also contaminant loading of nutrients, metals, and road salt used for deicing to water bodies. Here we investigate the impact that road salt has on denitrification in roadside environments. Sediments were collected from a series of forested and roadside wetlands and acclimated with a range of Cl concentrations from 0 to 5000 mg L1 for 96 h. Denitrification rates were measured by the isotope pairing technique using 15NeNO 3 , while denitrifying community structures were compared using terminal restriction fragment length polymorphism (TRFLP) of nitrous oxide reductase genes (nosZ). Chloride significantly (p < 0.05) inhibited denitrification in forested wetlands at a Cl dosage of 2500 or 5000 mg L1, but the decrease in denitrification rates was less and not significant for the roadside wetlands historically exposed to elevated concentrations of Cl. The difference could not be attributed to other significant changes in conditions, such as DOC concentrations, N species concentrations, or pH levels. Denitrifying communities, as measured by T-RFs of the nosZ gene, in the roadside wetlands with elevated concentration of Cl were distinctly different and more diverse compared to forested wetlands, and also different in roadside wetlands after 96 h exposures to Cl. The shifts in denitrifying communities seem to minimize the decrease in denitrification rates in the wetlands previously exposed to Cl. As development results in more Cl use and exposure to a broad range of natural or manmade wetland structures, an understanding of the seasonal effect of Cl on denitrification processes in these systems would aid in design or mitigation of the effects on N removal rates. © 2016 Elsevier Ltd. All rights reserved.
Keywords: Denitrification Chloride Microbial communities nosZ T-RFLP
1. Introduction Anthropogenic activities have increased the nitrogen (N) loading in rural, suburban, urban, and mixed land use watersheds. One of the most important consequences of higher N loading is the increase in N exports from watersheds to coastal ecosystems. Excessive N exports to receiving water bodies can result in eutrophication, hypoxic conditions, and/or loss of biodiversity (Paerl,
*
This paper has been recommended for acceptance by Charles Wong. * Corresponding author. Glenbrook Rd., Unit 3037, Storrs, CT 06269, USA. E-mail address: [email protected] (T.M. Vadas). 1 Present address of B.K. Song: Department of Biological Sciences, Virginia Institute of Marine Science, Gloucester Point, VA 23062, USA. http://dx.doi.org/10.1016/j.envpol.2016.01.068 0269-7491/© 2016 Elsevier Ltd. All rights reserved.
1997). The input rate of N to coastal ecosystems has doubled since the 1940's, mostly due to agriculture and the burning of fossil fuels (Vitousek et al., 1997). As N inputs increase, the ability of freshwater systems to act as a buffer for N processing between the terrestrial landscape and coastal ecosystems is more important. In forested streams networks, for example, up to 95% of N inputs are retained in the watershed (Groffman et al., 2004). However, retention of N in suburban/urban watersheds has been observed to range from 25% to 95% (Groffman et al., 2004), and the reasons for the variability are not well understood. A number of hydrological, geomorphological, and biogeochemical changes in developed watersheds as well as elevated N loading may alter N removal. Changes in hydrology are driven by the increase in impervious cover in developed areas which generates high volumes of stormwater runoff flowing into streams where
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N removal and uptake is variable (Groffman et al., 2002; Paul and Meyer, 2001; Sweeney et al., 2004). At the interfaces, in wetlands and riparian zones which typically have higher organic matter and longer retention times, denitrification rates are often higher. There are many instances of distributed wetlands such as vernal pools in developed landscapes, often adjacent to paved areas that could be active sites for denitrification. Capps et al. (2014) found that vernal pools are hotspots of leaf litter turnover and denitrification, playing a significant role in nutrient dynamics across the forested landscape. However, the stormwater runoff that may enter these sites has elevated concentrations of nutrients and inorganic contaminants (Bannerman et al., 1993). High concentrations of N in particular can lead to biological saturation where the available N exceeds the processing ability of organisms (O'Brien et al., 2007). N processing is further impacted by increased pollutants associated with stormwater including metals (Sakadevan et al., 1999) or chloride from road salt (Green and Cresser, 2008; Green et al., 2008; Groffman et al., 1995; Hale and Groffman, 2006). Road salt usage is high in developed landscapes contributing to high concentrations of Cl in surrounding water bodies, the negative environmental impacts of which have been documented (Findlay and Kelly, 2011). As Cl concentrations increase in urban and rural environments, reaching levels as high as 4600 mg L1 (Kaushal et al., 2005), determining the impacts that the combined increase in N and Cl will have on N processing within watersheds is critical. Elevated Cl concentrations are known to have negative impacts on aquatic plants and animals (Kaushal et al., 2005) and cause changes to microbial community structure and respiration rates (Baldwin et al., 2006). The impact of Cl on N cycling has also been investigated; however, the results vary. In forest soils treated with 100 mg L1 Cl, nitrification and N mineralization rates decreased, but denitrification was not inhibited (Groffman et al., 1995). Sodium chloride exposure has also been observed to lead to a release of ammonia N, an increase in pH and an increase in n et al., rates of nitrification and mineralization of organic N (Ardo 2013; Compton and Church, 2011; Green and Cresser, 2008; Green et al., 2008). Denitrification potential decreased in freshwater stream sediments from forested watersheds exposed to 2500 mg L1, but was unchanged in stream sediments from suburban watersheds (Hale and Groffman, 2006). Upon exposure to much higher Cl levels, 2e10% salinity, sediment denitrification rates typically did not vary (Laverman et al., 2007; Marton et al., 2012). These studies independently used a wide variety of salinity levels with differing effects on denitrification, but the Hale and Groffman (2006) study in particular, suggests that microbial communities that are exposed to elevated concentrations of Cl for prolonged periods of time may adapt to the physiological stress associated with an increase in salinity, and thus observed no significant change in denitrification rates in suburban streams. We hypothesize that denitrification rates in distributed freshwater wetlands that are continuously exposed to Cl concentrations from road salt applications will not vary with increased exposure to Cl, while wetlands not previously exposed to Cl will have decreased rates. The purpose of this study was to determine how Cl addition impacted denitrification rates and denitrifying communities in sediments historically impacted by Cl relative to unimpacted sediments. The study compares roadside wetland areas historically exposed to elevated concentrations of Cl to unexposed forested wetland areas. In conjunction with varying Cl exposure, changes in water chemistry and denitrification rates were measured. In addition, denitrifying community structure in the soils were compared based on Terminal Restriction Fragment Length Polymorphism (T-RFLP) of the gene encoding nitrous oxide reductase (nosZ), which is an enzyme that reduces N2O to N2. Here, we
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evaluate the impact of Cl on denitrification rates, DOC, concentrations of N species, and denitrifying community structures to demonstrate the potential impact that road salt use will have on N removal and retention in developed landscapes. 2. Materials and methods 2.1. Sample collection Sediment slurries were collected from six wetland locations in Northeast Connecticut, within the Yale Myers Experimental Forest (YEF) (Fig. S1). The majority of YEF is undeveloped with some surrounding suburban areas, and with routinely salted local roads passing through the area. The soil in YEF and the surrounding area is classified as glacial till and the region is forested by a mix of hardwoods (YEF, 2009). This region was chosen to isolate the role of Cl used in road salt applications (a mixture of sand and NaCl) in the area and minimize the potential effect of other urban development impacts (e.g. other contaminants or heat). The specific sampling locations were selected based on observed Cl levels and proximity to roadways, with three of the locations within 10 m of a roadway and three located more than 200 m from a roadway, each with similar canopy cover and size. These particular wetlands are vernal pools and were a subset of paired wetlands from prior research in the area (Brady, 2012). Pools were paired based on leaf-off hemispherical photographs, had maximum depths at the center of typically about 70 cm, and areas of about 500e1500 m2 (Brady, 2013). The soils collected were from areas with no vegetation, though emergent vegetation was present in some parts of the pool. The roadside pools receive salt inputs through either surface runoff or salt spray deposition. Historically, the elevated conductivity in these pools persists through the spring and early summer. Sediment and water samples were collected during the snowfall season in November and January. All sampling events occurred after snowstorms during which road salt was applied. During the month of January, all of the wetlands were covered with a thick layer of ice necessitating the use of an axe to break through the ice. Sediments were collected using a 200 diameter sediment corer. Three 600 length cores were collected from each location, large leaf litter was removed and sediments were homogenized to form a composite sample. Grab samples of the surface waters were collected into acid-washed (3% HCl v/v) HDPE bottles. Sediments for Cl exposure experiments were stored at 4 C in the dark and used within 96 h. Water samples were filtered (0.45 mm) and frozen (20 C) in the dark prior to analysis of dissolved organic carbon (DOC), Cl, pH, alkalinity, and the concentration of different N species. 2.2. Chloride exposure microcosms Sediments samples were exposed to a range of Cl concentrations, diluted from NaCl stock, the most common road salt (Findlay and Kelly, 2011), for 96 h at a ratio of 20 g wet sediment to 40 mL salt solution. Salt exposure concentrations were 0, 500, 1000, 2500, or 5000 mg L1 Cl. After acclimating for 96 h, 2 g of wet sediment from each slurry was divided into a 12 mL Exetainer® tube (Labco, High Wycombe, United Kingdom), sealed with caps containing gastight septa and flushed with He gas according to published methods (Long et al., 2013). The vials were incubated overnight at 20 C to reduce the background concentrations of nitrate in the sample prior to laboratory assessment of the denitrification rates using isotope pairing. Sediment sub-samples for microbial community analysis were frozen at 80 C prior to extractions for denitrifying community analysis described below. Additional microcosms were established to measure changes in
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water chemistry parameters as a function of Cl exposure that might alter microbial processes. Microcosms were prepared in HDPE bottles by mixing 150 g of wet sediment with 300 mL of filtered site water. Sodium chloride was added to establish Cl concentrations of 0, 500, 1000, 2500, or 5000 mg L1, and samples were placed sideways on a shaker table rotating at 60 rpm for 96 h. The slurries were first centrifuged and the supernatant filtered (0.45 mm) prior to analysis for DOC, Cl, pH, alkalinity, and concentrations of different N species. 2.3. Water and soil chemistry analysis Filtered water samples from the wetlands and Cl microcosms were analyzed for DOC, Cl, pH, alkalinity, Total Nitrogen (TN), þ NO 3 þ NO2 , NH4 , and dissolved organic N concentrations via standard methods (APHA, 2000). The NO 3 þ NO2 data are referred to as NO3 in the results. DOC concentrations were analyzed via thermal oxidation (ASTM, 1994b) using an Apollo 9000 TOC analyzer (Teledyne Tekmar Company, Mason, OH). Chloride concentrations and pH were analyzed via ion-selective electrodes (Fisher Scientific Accumet probes, Pittsburgh, PA). Alkalinity was measured via titration to an endpoint of pH 4.3 (ASTM, 1994a). Nitrogen species concentrations were analyzed via a Lachat QuikChem 8500 Flow Injection Ion Auto Analyzer using Lachat Instruments (Loveland, CO) methods 10-107-04-1-A for nitrate þ nitrite, 12-107-06-1-A for ammonia, and 10-107-04-3-D for total N. Detection limits were 0.003, 0.003, and 0.008 mg L1 for ammonia, nitrate and total N, respectively. Percent organic matter was measured via loss on ignition at 450 C (Robertson et al., 1999). Total organic carbon and total nitrogen in soils were measured with a Perkin Elmer® 2400 CHN analyzer according to EPA Method 440. Soil pH was measured in a 1:1 (v/v) soil slurry according to EPA Method 9045D. Quality control was assessed via continuous calibration verification, lab duplicates and blanks, field duplicates and blanks, and matrix spikes. Recovery of spikes and quality control samples were within 10% of expected values. Differences in concentrations between forested and roadside sediments were evaluated using a student's t-test at the 95% confidence interval using Sigmaplot 13.0 software. 2.4. Denitrification rate analysis Denitrification and anammox rates were determined based on 30 the transformation of 15N-labeled NO N2 and 29N2, respec3 to tively, at 20 C (Thamdrup and Dalsgaard, 2002). This method yields a metabolic potential without N limitation that standardizes for NO 3 availability under anaerobic conditions (Long et al., 2013). After anoxic incubation under a He headspace to remove existing nitrate as stated above, the exetainers® were re-flushed with He for 14 5 min and spiked with 0.1 mL of 550 mM 15NeNO NeNHþ 3 and 4 using a He-flushed syringe. Following spiking, samples were vortexed and allowed to equilibrate for 30 min. A continuous flow isotope ratio mass spectrometer (IRMS; Thermo Delta V, Thermo Scientific, Waltham, MA) was programmed to collect gas samples from the vials four different times at 15 min intervals. Helium blanks were included at the beginning and end of each run to detect any potential leaks from the vials during the experiment. The peak area measurements for 30N2 and 29N2 were converted to mass using air calibration standards (Scotts Specialty Gases, Philadelphia, PA). Denitrification rates were determined from the slope of the linear regression of the time series data. In order to compare rates to other studies, rates on a per gram sediment basis were converted to a per area (m2) basis using a sediment bulk density of 2 g/cm3 and a conservative estimate of the denitrification zone thickness of 0.5 cm (David et al., 2006). Anammox rates were below detection in
these samples. Differences in rates between forested and roadside sediments were evaluated using a student's t-test at the 95% confidence interval using Sigmaplot 13.0 software. 2.5. Denitrifying community analysis DNA was extracted from 0.25 g of the Cl exposure microcosm sediments described above following a 96 h exposure. A PowerSoil DNA Isolation Kit (MO BIO Laboratories, Inc.) was used, following the manufacturer's protocol. The nosZ genes were amplified using the GoTaq Master mix (Promega, Fitchburg WI) with the primers of 6-FAM-labled nosZ1F and nosZ2R as described in Henry et al. (2006). The amplicons were purified using the Wizard SV Gel and PCR Clean-Up System (Promega, Fitchburg WI). DNA concentration of purified amplicons was measured using the Quant-iT ds DNA Assay Kit (Life Technologies, Carlsbad, CA). A total of 20 ng of PCR product was digested overnight at 37 C with 5 units of Cfo1 restriction endonuclease (Promega, Fitchburg WI). The digested products were precipitated with isopropenol and run on a 3130x/ Genetic Analyzer (Life Technologies, Carlsbad CA). Fragment analysis was conducted using the Gene Mapper 4.0 (Life Technologies, Carlsbad CA) and T-RFLP Analysis Expedited (T-REX) programs (Culman et al., 2009). Variations in nosZ gene fingerprints were assessed using a BrayeCurtis similarity matrix and cluster analysis in the Primer-5 software package (Primer-E Ltd, Lutton, UK). Sites CF and CR did not yield nosZ gene amplifications. 3. Results and discussion 3.1. Water and soil chemistry Dissolved organic carbon concentrations in samples collected during November were higher at all but one sampling site, but not significantly different than samples collected in January (Table 1). Forested wetlands had, on average, a DOC concentration of 11 mg L1 in the fall and 5.8 mg L1 in winter. The average DOC concentration in roadside wetlands were not significantly different between November and January, although site C had much higher concentrations of between 6.1 and 8.7 mg L1. Higher concentrations were expected during fall months since it is at the end of the increases in primary productivity generally observed during the summer and also includes fresh allocthonous inputs (Mann and Wetzel, 1995). With the exception of CR, the results presented here are consistent with the expected seasonal trend. CR was the only wetland that was not covered with ice during the January sampling event and the higher DOC concentration at that time might have been influenced by runoff, snowmelt, or the warmer wetland temperature (Westerhoff and Anning, 2000). The concentrations of Cl measured in the roadside wetlands were significantly higher (p < 0.05) than concentrations measured in forested wetlands (Table 1). On average the roadside wetlands were about 30-fold higher in Cl concentration compared to the forest wetlands, and concentrations were slightly lower in January. This was a unique year in that a heavy snowstorm hit New England at the end of October, just before sampling, resulting in the increase of road salt application causing high Cl concentrations prior to the typical winter season. The elevated concentrations of Cl in roadside wetlands are consistent with previous observed measurements which showed Cl concentrations ranging from 80e300 mg L1 in the roadside wetlands and 2e6 mg L1 in forested wetlands during late spring and summer months that helped identify potential sampling locations. Although observed concentrations were lower in the spring and summer due to an increase in rainfall and the termination of road salt application, roadside wetlands still held elevated concentrations throughout
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Table 1 Average water quality dissolved concentrations measured for forested and roadside wetlands in November and January. All concentrations are in mg L1 ± 1 SD. DOC e dissolved organic carbon; Cl e chloride; TN e total nitrogen (N); NO3 e total nitrate plus nitrite; NH4 e total ammonia species; DON e dissolved organic N (calculated by difference).
November January
mg L1
DOC
Forested Roadside Forested Roadside
11 3.8 5.8 3.7
± ± ± ±
Cl 4.2 2.0 4.0 4.3
NO3eN
TN
160 4800 110 3600
± ± ± ±
37a 550b 18a 2200b
0.26 0.10 0.27 0.25
± ± ± ±
0.052a 0.06b 0.20ac 0.27bc
0.008 0.006 0.027 0.036
± ± ± ±
NH4eN 0.004ac 0.003a 0.015bc 0.015b
0.019 0.016 0.11 0.14
± ± ± ±
DON 0.001 0.017 0.11 0.20
0.24 0.078 0.13 0.073
± ± ± ±
0.06a 0.045b 0.10ac 0.087bc
A difference in the letter after the value indicates significance at p < 0.05 between water quality parameters.
the year compared to forested wetlands. It is not uncommon for Cl concentrations to remain elevated during the year due to storage in the watershed and aquifers (Findlay and Kelly, 2011). Concentrations of different N species were all less than 0.3 mg L1 and DON was in the highest proportion (Table 1). Only a few comparisons showed significant differences (p < 0.05), including the roadside and forested wetland concentrations of TN and DON in November, and NO3 concentrations between November and January. The concentration of NO3 in both sets of wetlands was higher during the January sampling event. Typically roadside environments have higher concentrations of NH4 and NO3 compared to forested environments due to near source mobile emissions within 10 m of the roadway (Bettez et al., 2013), but they were not significantly different in this case. The significantly higher DON in the forested wetland compared to the roadside wetland during November likely reflects a combination of increased microbial activity and leaf litter input in the interior of the forest. By January, a larger portion of TN was in the form of NH4, possibly from the mineralization of biomass during a relatively mild winter, but there was no significant difference between forested and roadside wetlands. The soils used in this study had varying organic matter content, but a similar and relatively low C:N ratio (Table 2). There was not a clear trend among the sites, i.e. not all roadside ponds had the lowest organic matter, and TOC and TN were actually not significantly different between any sites. The C:N ratio was relatively low, less than 17 at all sites, and that is below the level that is considered N limiting in the literature (Palta et al., 2014). Soil pH was over 5 in both roadside and forested wetlands, but slightly lower on average in the roadside wetlands. Acidification has been observed in watersheds exposed to salt solutions and could be due to cation exchange at those sites (Wright et al., 1988). In the wetlands that were selected for this study, excess N inputs are limited to atmospheric deposition and stormwater runoff from impervious surfaces. Direct atmospheric deposition likely influenced the forested wetlands in the study area to the same extent that it impacted the roadside wetlands since wetlands were selected based on similar canopy cover. Runoff from impervious surfaces was limited to the adjacent roadway and did not appear to carry higher concentrations of NO3 into the roadside wetlands. While the measured concentrations reflect the winter season when
biological activity is greatly reduced, concentrations in roadside wetlands here or in similar developed landscapes in the spring or summer are expected to increase from higher primary productivity or fertilizer application. Overall, our initial assessment of aqueous water quality parameters indicates that the most distinct difference between the forested and roadside sites was Cl concentrations. Chloride exposure microcosms were established to assess the difference in denitrification rates in previously exposed (roadside) or unexposed (forested) sediments at a range of Cl concentrations. 3.2. Denitrification rates The rate of denitrification decreased sharply at a Cl dosage of 2500 mg L1 and above in both the forested and roadside wetlands, but was only significantly different (p < 0.05) from the control (zero Cl dose) in the forested wetlands (Fig. 1). The average denitrification rates were similar and not significantly different between the sites at 1000 mg L1 and below and were similar to those typically observed in freshwater sediments. Aerial rates in these wetlands ranged from 14 to 53 mmol N/m2/hr, which is near the low end of the rates observed in the sediments of freshwater lakes (Pina-Ochoa and Alvarez-Cobelas, 2006). Some studies have observed stimulatory effects of increasing salinity on denitrification rates in tidal forested wetlands (Marton et al., 2012), but that was not observed here. At a Cl dosage of 2500 mg L1, the rate of denitrification dropped significantly (p < 0.05) by 99% in the forested wetlands. Rates in roadside wetlands dropped by 90%, but the decrease was not statistically significant. At 5000 mg L1 the rates decreased slightly more, by 99% and 94% for the forested and roadside wetlands, respectively. Cl has been demonstrated to negatively impact denitrification rates in an estuary, with rates
Table 2 Soil characteristics measured for forested and roadside wetlands. Site
Organic matter (%)
AF AR BF BR CF CR
27.5 3.4 14.0 10.4 9.9 33.3
± ± ± ± ± ±
11.7 0.1a 2.7 1.5b 5.1 8.6b
TOC (%) 13.7 1.6 4.6 2.5 4.2 12.5
± ± ± ± ± ±
6.9 0.6 1.5 0.5 0.6 3.6
TN (%) 0.88 0.14 0.34 0.2 0.33 0.76
± ± ± ± ± ±
C:N ratio 0.42 0.04 0.14 0.04 0.06 0.25
15.5 11.6 13.9 12.9 12.7 16.7
± ± ± ± ± ±
0.4a 0.5b 1.4acd 0.2bc 0.5d 0.7e
Soil pH 5.57 5.29 5.53 5.23 5.62 5.45
A difference in the letter after the value indicates significance at p < 0.05 between parameters.
Fig. 1. Average denitrification rates for forested and roadside wetlands as a function of Cl dosage to microcosms. Error bars reflect the standard deviation of three replicate wetlands. A * above the data indicates a significant difference between roadside and forested wetlands (p < 0.05).
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measured at the more saline end of a salinity gradient being lower relative to those at the freshwater end (Fear et al., 2005; Tobias et al., 2001), but this could be a function of DOC availability (Wallenstein et al., 2006) or other water quality parameters as well. In order to assess any impact water quality parameters might have had on denitrification rates, we also assessed the changes in these same parameters during Cl dosage experiments. While the average DOC concentrations did decrease upon Cl exposure, the decreases were not significantly different (p < 0.05) from the control (Fig. 2). There was only one significant difference in DOC observed between forested and roadside wetlands at an exposure concentration of 5000 mg L1 Cl. There was a general trend of decreasing DOC concentrations as chloride dosage increased, from starting concentrations of 14 or 5 mg L1 in the forested and roadside wetlands, respectively, to around 6 and 2.3 mg L1, respectively, at 5000 mg L1 Cl, though the roadside wetlands always had lower average DOC concentrations compared to the forested wetlands. This was at least in part due to the fact that at least two of the roadside wetland soils had lower C content relative to forested wetland soils (Table 2). Fear et al. (2005) also measured lower organic matter at the more saline end of the salinity gradient making a determination of the relative significance of Cl and DOC on the denitrification rate difficult. One possible reason for the loss of DOC is coagulation and settling or filtration induced by the addition of NaCl, although Naþ is not known to be a particularly effective coagulant (Stumm and Morgan, 1981). Pearson correlation coefficients do not suggest a significant correlation between DOC concentrations and Cl1 dosages (Table S1) and even if the Naþ did have an effect on the reduction in DOC, it is uncertain why there was a significant difference in DOC at 5000 mg L1 when both sediments were exposed to the same Cl concentration. Although the average DOC concentration decreased by nearly half in both sets of wetlands after the addition of 500 mg L1 Cl, they remained largely unchanged as higher doses of Cl were added. Despite the decrease in DOC concentration with the addition of 500 mg L1 Cl, which could potentially produce a corresponding decrease in denitrification rates due to carbon limitation, denitrification rates did not significantly decrease. In contrast, denitrification rates only significantly declined at 1000 mg L1 and above in forested wetlands as the concentration of Cl increased suggesting that the decrease was a function of Cl concentration and not DOC.
Fig. 2. Dissolved organic carbon (DOC) concentrations from filtered 96 h incubations relative to measured Cl dosage. Error bars reflect the standard deviation across three wetlands. A * above the data indicates a significant difference between roadside and forested wetlands (p < 0.05).
The alkalinity and pH both decreased as Cl concentration increased. Similar to the pattern observed in DOC concentrations, the largest drop in pH occurred between 0 and 500 mg L1 Cl dosage and the pH continued to decrease in smaller increments as the Cl concentration increased. Initial pH values for both sets of wetlands were 5.3e5.6 and decreased to 4.3 at 5000 mg L1 Cl dosage (Table 3). Alkalinity was lower in the roadside wetlands with a base level of 0.80 meq/L decreasing to 0.19 meq/L with Cl addition. The change in alkalinity was greater in the forested wetlands with a base level of 1.17 meq/L and decreased to 0.17 meq/ L with Cl addition. The decrease in pH and alkalinity likely resulted from the increase in NHþ 4 in the mesocosms as Cl concentrations increased (Table 3). A drop in pH has been observed before in runoff from soils exposed to salt water due to cation exchange (Wright et al., 1988). Decreases in pH have been shown to decrease denitrification rates (Baeseman et al., 2006). The optimal pH range for denitrification to occur is between 6 and 8. In the current study, the pH was below 6 and the pH values decreased in a similar pattern for both sets of wetlands. Since the denitrification rates exhibit contrasting patterns, especially at the lower concentrations of Cl, it appears that the decrease in pH was not as influential as the increase in Cl, a fact supported by the continued decrease in denitrification rate above the 1000 mg L1 dosage without further decreases in pH. The most dramatic differences observed in N species concenþ trations were for NHþ 4 where NH4 concentrations increased significantly (p < 0.05) in both forested and roadside wetlands from 0 to 5000 mg L1 Cl, increasing NHþ 4 in solution from about 0.5 or less mg L1 to 2.5 and 1.3 mg L1 in forested and roadside wetlands, respectively (Table 3). The addition of Naþ likely exchanges NHþ 4 on the sediments, a result that is consistent with other studies (Compton and Church, 2011; Green and Cresser, 2008; Green et al., 2008; Laverman et al., 2007). This increase in NHþ 4 could stimulate nitrification in some sediment systems (Baldwin et al., 2006), but we did not observe a corresponding increase in NO 3 concentration, and actually saw a decrease in NO 3 which likely contributed to the decrease in observed denitrification rates. Others have also observed a short term decrease in NO3 after incubation with NaCl over one day (Compton and Church, 2011). While TN increased since the majority of it consisted of NH4, DON was reduced in all cases, though not statistically significant. This is consistent with the reduction in concentration of DOC. Although there was not a statistically significant decrease in NO 3 , the average concentration went from 0.23 mg L1 in the forested wetlands at 500 mg L1 Cl dosage to 0.01 mg L1 at 5000 mg L1 Cl, which is about the same percentage drop in denitrification rate that was observed. Only at the dosage of 2500 mg L1 Cl was the difference between forested and roadside sediments significant (p < 0.05) with an 85% reduction in denitrification rate in the forested wetlands, suggesting a possible acclimation of the roadside sediments to Cl exposure. Hale and Groffman (2006) showed a similar trend while measuring denitrifying enzyme activity (DEA) in sediments that were exposed to Cl concentrations of 2500 mg L1 prior to the laboratory incubations and sediments that had not been previously exposed to high Cl concentrations. However, their findings (Hale and Groffman, 2006) showed no change in DEA in sediments incubated with 2500 mg L1 of Cl that had been previously exposed to Cl (i.e. the urban/suburban stream) while we observed a significant drop in denitrification rate. Chloride concentrations at the stream site they used had been recorded as high as 4600 mg L1 in the winter but were closer to 100 mg L1 at the time the sediments were collected in the summer, at which time the microbial population may have been different. Marton et al. (2012) also reported no change in denitrification potential at most locations after incubating sediments with 2000 or 5000 mg L1 Cl. At the time of
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Table 3 Nitrogen species concentrations in filtered samples for roadside and forested wetlands after 96 h exposure to different Cl concentrations. NH4, NO3, TN, and DON represent total ammonia, total oxidized nitrogen, total nitrogen and dissolved organic N species, respectively. Cl dosage (mg L1) 0
500
Forest NH4 NO3 TN DON pH
0.57 0.26 1.82 0.99 5.57
± ± ± ± ±
Roadside 0.32a,c 0.20a,c 0.79a,c 0.50a,c 0.05a,c
0.42 0.17 0.97 0.37 5.32
± ± ± ± ±
0.34a,e 0.17a,c 0.83a,c 0.33a,c 0.11b,f
5000
Forest 1.48 0.23 2.18 0.46 4.90
± ± ± ± ±
Roadside 0.76a,cd 0.15a,c 1.21a,c 0.36a,c 0.16a,d
0.56 0.16 0.96 0.24 4.83
± ± ± ± ±
0.32a,e 0.08a,c 0.61a,c 0.23a,c 0.21a,g
Forest 2.53 0.01 2.84 0.30 4.3
± ± ± ± ±
Roadside 0.55a,d 0.01a,c 0.78a,c 0.24a,c 0.04a,e
1.27 0.02 1.22 0.12 4.27
± ± ± ± ±
0.14b,f 0.03a,c 0.56b,c 0.11a,c 0.51a,g
A difference in the first letter after the value indicates significance at p < 0.05 between forest and roadside pairs at the same Cl dosage, while a difference in the second letter after the value indicates significance at p < 0.05 between Cl dosages at the same location.
sample collection (Marton et al., 2012), river water and pore water were freshwater and there was no historic data to indicate if the sediments had been exposed to Cl previously. In our study, Cl concentrations were elevated for the year at the time of sample collection. Our denitrification rate data suggests there was an adaptation due to long-term Cl exposure in these study sites, and perhaps there was also a threshold to a significant decrease in denitrification activity between 1000 and 2500 mg L1 Cl dosage that was also exceeded in the current study. 3.3. Comparison of denitrifying communities PCR of nosZ genes was successful for most samples except for the CF and CR samples. Low quality and quantity of extracted DNA may inhibit amplification of nosZ genes. With the nosZ gene detected in the other soil communities, T-RFLP patterns demonstrated distinct differences in the nosZ gene carrying denitrifying communities in the forested wetlands relative to the roadside wetlands, and they tended to group together. The cluster analysis of the nosZ T-RFLP data (Fig. S2) showed that denitrification assemblages from forested sites (AF and BF) grouped together, with a greater than 90% similarity, as did the roadside wetlands (AR and BR) with about 90% similarity. In addition, the AR and BR sites amended with Cl were only about 75 and 70% similar, respectively, suggesting different nosZ gene pools were expressed upon Cl exposure. In order to examine further details of denitrifying communities, the relative percent abundance of each terminal restriction fragment (T-RF), each nominally representing a different genetic variant, was calculated based on the heights of T-RFs observed (Fig. 3). There were clear differences in populations in forested versus roadside wetlands, with some T-RFs showing up
only in forested sites, e.g. T-RFs with 53, 323 and 402 bp at the AF site (Fig. 3A). In the BR roadside sites, the denitrifying populations corresponding to the T-RFs with 51, 255 and 320 bp disappeared, while the T-RFs with 317, 341, 348 and 445 bp appeared or increased in the roadside wetlands treated with elevated Cl concentration (Fig. 3B). In the AR roadside wetland, T-RFs with 320, 445 and 473 bp decreased or disappeared while T-RFs with 317 and 336 increased with the Cl exposure (Fig. 3B). Thus, T-RFLP fingerprints clearly demonstrated that there were shifts in nosZ gene carrying denitrifying organisms, corresponding both to a difference in the forested and roadside wetlands, as well as in response to the elevated Cl exposure in the roadside wetland sediments. 3.4. Other factors affecting denitrification Correlation matrices (Table S1) for each set of wetlands combined with the chemical trend data suggest that Cl concentration is the primary driver of denitrification inhibition in these wetland sediments. Denitrification rates evaluated across individual timespecific wetland data were significantly (p < 0.05) negatively correlated with Cl concentrations for both the forested and roadside wetlands (R ¼ 0.64 and 0.69, respectively). Other water chemistry factors that could affect denitrification were positively correlated (p < 0.05) for the forested (pH, R ¼ 0.81) and roadside (DOC, R ¼ 0.60) wetlands. However, the relationship of pH to denitrification rates likely is strongly influenced by the cocorrelation of Cl addition on pH due to reduction in pH during exposure. Similarly, while DOC was correlated significantly (p > 0.05) with denitrification for the forested wetlands, the relationship was not significant when evaluated across all data. Additionally, the DOC concentrations and pH levels in the two sets of
Fig. 3. Comparison of denitrifying communities based on selected nosZ gene T-RFs in the A) sediments collected from the forested wetlands (AF and BF) and roadside wetlands (BR and AR), and B) in the wetland sediments with 5000 mg L1 Cl concentration treatments (BR þ Cl and AR þ Cl) or without treatments (BR and AR). Relative abundance of each T-RF was calculated based on the height of T-RFs in each fingerprint.
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microcosms contradict the trends that would be expected if these chemical shifts were impacting denitrification rates. Although we chose roadsides sites in less developed areas, it is likely that trace metal concentrations were higher in roadside wetlands than forested wetlands. Increasing concentrations of NaCl further in the mesocosms would likely have resulted in additional €ckstro € ma et al., 2004; trace metal release from the sediments (Ba Lofgren, 2001). However, if metals were responsible for the inhibition of denitrification, which has been observed in some cases (Sakadevan et al., 1999), there likely would have been a more significant decrease in denitrification with the addition of Cl to roadside wetlands compared to forested wetlands where the concentrations of metals were likely lower. The fact that denitrification rates in roadside wetlands did not decrease as much as those in the forested wetlands at higher Cl additions again supports Cl as the primary factor impacting denitrification rates. Molecular analysis suggests that the denitrifying community structure differs between the roadside and forested wetlands. The long-term stressors from roadside disturbance may lead to a shift in denitrifying community structures. Our T-RFLP analysis is limited based on the identification of only nosZ genes and the measurement of N2 as the final product. Future studies could examine N2O production as well as the complete denitrification gene pathway to get a more complete picture of changes in the denitrifying community. There is no data in the literature showing shifts in denitrifier communities in response to Cl exposure, but Cl has been shown to impact methane production and resulted in significant changes in T-RFLP of archeal genes in freshwater wetlands exposed to NaCl (Baldwin et al., 2006). Regardless, the higher denitrification rates in the roadside wetlands with Cl addition suggest that nosZ containing denitrifying communities in the roadside wetlands are able to accommodate the exposure to additional Cl better than the forested wetlands. Given that the roadside wetlands historically have been exposed to elevated concentrations of Cl, we hypothesize that the denitrifying community may have adapted to the presence of Cl. In comparison to other studies, the results generated here support the hypothesis that historic exposure results in microbial function adapted to Cl (Hale and Groffman, 2006; Marton et al., 2012). However, denitrification rates measured in the roadside wetlands only significantly decreased at the highest concentrations of Cl added (2500 and 5000 mg L1), suggesting that even with historic exposure to Cl, a threshold exists beyond which denitrification rates are severely impacted. Our study sites differed from those of prior studies in having a longer retention time (i.e. increased time for adaptation to occur) and in being hydrologically isolated (i.e. lower chance of recruitment). In combination with the input concentrations or nutrients as well as stressors, these differences are critical for determining the real impacts of road salt on N removal in distributed wetlands near developed landscapes. Similar reductions in N processing could be observed in manmade structures, such as dry ponds, wet ponds, or treatment wetlands meant to capture and reduce both volumes of flow and contaminant concentrations. While it is expected that wet ponds and wetlands, in particular, can provide a hot spot for N removal (EPA, 2012), their performance in practice over the long-term and the variation seasonally is unknown. Given the likelihood of Cl exposure to both natural and manmade wet ecosystems during winter storm events, as well as the lingering of Cl in the ecosystem through the spring, it is possible that denitrification rates are significantly reduced during winter and spring when significant levels of NO 3 come through the system from spring thaw and the start of the growing season. Cl concentrations have been measured as high as 25,000 mg L1 in snow melt (Makepeace et al., 1995) which is well above the concentration after which we saw a
greater than 90% reduction in denitrification rates. In these particular wetlands, elevated conductivity levels were observed in the roadside wetlands in the following spring and summer. Whereas the conductivity in forested wetlands remained less than 30 mS/cm, the roadside wetlands remained typically above 400 mS/ cm, and in the case of site CR, remained above 1000 mS/cm until June. While these are only conductivity values, and not Cl, it is reasonable to assume Cl is significantly contributing to the increased conductivity, especially because these are vernal pools with little or no flow. It is possible that in more developed areas where salt use may be higher or more frequent, or in wetlands with little or no outflow, Cl concentrations could persist at levels shown to dramatically decrease denitrification rates. Our microbial data suggests distinct differences in the denitrifying populations between both previously exposed roadside wetlands and unexposed forested wetlands, as well as upon Cl exposure. Future studies could focus on a more detailed analysis of denitrifying communities, encompassing the full denitrification pathway genes using more advanced methods to answer questions regarding the ability of denitrifying communities to adapt to increases in salinity. Determining the threshold at which denitrifying communities can no longer denitrify and the timing required for adaptation to occur could aid in either setting limitations on Cl exposure in certain productive wetland areas, or in the design of functional stormwater capture systems for N removal to protect coastal ecosystems from the detrimental impacts of excess N loading. Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.envpol.2016.01.068. References APHA, 2000. Standard Methods for the Examination of Water and Wastewater, 15th ed. American Public Health Association, Washington, DC. n, M., Morse, J.L., Colman, B.P., Bernhardt, E.S., 2013. Drought-induced saltArdo water incursion leads to increased wetland nitrogen export. Glob. Change Biol. 19, 2976e2985. ASTM, 1994a. Method 2320B: Carbonate, Biocarbonate and Total Alkalinity. American Society of Testing & Materials, Philadelphia, PA. ASTM, 1994b. Method D4129-88 Standard Test Method for Total and Organic Carbon in Water by High Temperature Oxidation and by Coulometric Detection. American Society of Testing & Materials, Philadelphia, PA. €ckstro €ma, M., Karlssona, S., B€ Ba ackmanb, L., Folkesonb, L., Lindc, B., 2004. Mobilisation of heavy metals by deicing salts in a roadside environment. Water Res. 38, 720e732. Baeseman, J.L., Smith, R.L., Silverstein, J., 2006. Denitrification potential in stream sediments impacted by acid mine drainage: effects of pH, various electron donors, and iron. Microb. Ecol. 51, 232e241. Baldwin, D.S., Rees, G.N., Mitchell, A.M., Watson, G., Williams, J., 2006. The shortterm effects of salinization on anaerobic nutrient cycling and microbial community structure in sediment from a freshwater wetland. Wetlands 26, 455e464. Bannerman, R.T., Owens, D.W., Dodds, R.B., Hornewer, N.J., 1993. Sources of pollutants in wisconsin stormwater. Water Sci. Technol. 28, 241e259. Bettez, N.D., Marino, R., Howarth, R.W., Davidson, E.A., 2013. Roads as nitrogen deposition hot spots. Biogeochemistry 114, 149e163. Brady, S., 2012. Road to evolution? Local adaptation to road adjacency in an amphibian (Ambystoma maculatum). Sci. Rep. 2, 235. Brady, S.P., 2013. Microgeographic maladaptive performance and deme depression in response to roads and runoff. PeerJ 1. Capps, K.A., Rancatti, R., Tomczyk, N., Parr, T.B., Calhoun, A.J.K., Hunter, M., 2014. Biogeochemical hotspots in forested landscapes: the role of vernal pools in denitrification and organic matter processing. Ecosystems 17, 1455e1468. Compton, J.E., Church, M.R., 2011. Salt additions alter short-term nitrogen and carbon mobilization in a coastal Oregon Andisol. J. Environ. Qual. 40, 1601e1606. Culman, S.W., Bukowski, R., Gauch, H.G., Cadillo-Quiroz, H., Buckley, D.H., 2009. TREX: software for the processing and analysis of T-RFLP data. BMC Bioinform. 10. David, M.B., Wall, L.G., Royer, T.V., Tank, J.L., 2006. Denitrification and the nitrogen budget of a reservoir in an agricultural landscape. Ecol. Appl. 16, 2177e2190.
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