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Keywords:

  • silica;
  • watershed;
  • diatoms;
  • gypsy moth;
  • nitrate

Abstract

  1. Top of page
  2. Abstract
  3. 1. Introduction
  4. 2. Site Description
  5. 3. Methods
  6. 4. Results
  7. 5. Discussion
  8. 6. Conclusions
  9. Acknowledgments
  10. References
  11. Supporting Information

[1] Dissolved silica concentrations in western Virginia streams showed a significant bias toward declines (p < 0.0001) over the time period from 1988 to 2003. Streams with the greatest declines were those that had the highest mean dissolved silica concentrations, specific to watersheds underlain by basaltic and granitic bedrock. We examined potential geochemical, hydrological, and biological factors that could account for the observed widespread declines, focusing on six core watersheds where weekly stream chemistry data were available. No relationships were evident between stream water dissolved silica concentrations and pH, a finding supported by the results from a geochemical model applied to the dominant bedrock mineralogy. Along with changes in watershed acidity, changes in precipitation and discharge were also discounted since no significant trends were observed over the study period. Analyses of two longer-term data sets that extend back to 1979 revealed that the initiation of the dissolved silica declines coincided with the timing of a gypsy moth (Lymantria dispar) defoliation event. We develop a conceptual model centered on benthic diatoms, which are found within each of the six core watersheds but in greater abundance in the more silica-rich streams. Gypsy moth defoliation led to greater sunlight penetration and enhanced nitrate concentrations in the streams, which could have spurred population growth and silica uptake. The model can explain why the observed declines are primarily driven by decreased concentrations during low-flow conditions. This study illustrates lasting effects of disturbance on watershed biogeochemistry, in this case causing decadal-scale variability in stream water dissolved silica concentrations.

1. Introduction

  1. Top of page
  2. Abstract
  3. 1. Introduction
  4. 2. Site Description
  5. 3. Methods
  6. 4. Results
  7. 5. Discussion
  8. 6. Conclusions
  9. Acknowledgments
  10. References
  11. Supporting Information

[2] Stream water dissolved silica concentrations are thought to vary over long timescales (i.e., centuries to millennia) owing to factors such as temperature [Walker et al., 1981] and atmospheric CO2 levels [Berner et al., 1983; Brady, 1991; Brady and Carroll, 1994], and over short timescales (i.e., days to months) due to differences in hydrological flow paths [Wels, 1991; Scanlon et al., 2001] and mineral contact time [Hornberger et al., 2001; Subagyono et al., 2005]. For intermediate timescales on the order of years to decades, silica is generally assumed to be a stable constituent of stream water [Davis, 1964; Likens et al., 1977]. Most watershed studies until now have focused on event-based or seasonal variations in dissolved silica concentrations, a fact that can be attributed to the above assumption as well as to the scarcity of longer-term data sets. Given that chemical weathering of rock material is the only significant source of dissolved silica in stream water [Likens et al., 1977], this particular solute is considered to be a useful indicator of the biogeochemical and hydrological processes that occur within watersheds. In this study, we examine temporal trends of dissolved silica concentrations over intermediate timescales and infer the dominant factors that influence dissolved silica dynamics.

[3] This analysis of stream chemistry trends is aided by a spatially and temporally extensive data set. Hydrochemical data have been collected in Shenandoah National Park (SNP) and surrounding areas of western Virginia as part of the Shenandoah Watershed Study (SWAS) since 1979. Weekly stream chemistry data are collected at six core sites, with records for two of these sites extending back to the initiation of the program. The monitoring effort at the core sites is augmented by quarterly sampling at an additional 60 sites, as part of the Virginia Trout Stream Sensitivity Study (VTSSS), which began in 1988. Locations of the sampling sites are shown in Figure 1, and the durations of the stream chemistry records analyzed in this study are provided in Table 1.

image

Figure 1. Shenandoah Watershed Study (SWAS) and Virginia Trout Steam Sensitivity Study (VTSSS) watersheds sampling locations. Circle diameters indicate the magnitude of the dissolved silica concentration trends and solid circles represent the direction of observed trends. Areas of the six core watersheds are shown on the map of Shenandoah National Park.

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Table 1. Characteristics and Sampling History for the Six Core Watersheds in Shenandoah National Park That Are Sampled on a Weekly Basisa
 Watershed Area, km2Length of RecordBedrock ClassMain Geological Formations [Gathright, 1976]Mean Stream pH (Range)Mean [SiO2], μmol L−1 (SiO2 Trend, μmol L−1 yr−1)
  • a

    Stream chemistry statistics apply to the entire lengths of record, as indicated for each watershed.

  • b

    Hampton Formation consists of metasiltstone, metasandstone, and phyllite.

  • c

    Weaverton Formation is mainly quartzite.

  • d

    Erwin Formation is a resistant quartzite unit.

  • e

    Pedlar Formation consists of granodiorite.

  • f

    Catoctin Formation is a thick, metamorphosed basalt.

White Oak Run (WOR1)5.11979–2003siliciclasticmainly Hampton Formation,b some Weaverton Formationc6.1 (5.2–6.7)80.1 (−0.01)
Deep Run (DR01)3.11979–2003siliciclasticHampton Formation and Erwin Formationd5.5 (4.8–6.7)94.17 (−0.05)
Paine Run (PAIN)12.41987–2003siliciclasticmainly Hampton Formation, some Erwin Formation5.8 (4.6–6.4)88.8 (−0.80)
North Fork Dry Run (NFDR)2.31987–2003graniticPedlar Formatione6.5 (5.9–7.2)169.9 (−0.48)
Staunton River (STAN)10.51992–2003graniticPedlar Formation6.7 (6.0–7.2)139.4 (−0.82)
Piney River (PINE)12.61992–2003basalticmainly Catoctin Formation,f some Pedlar Formation7.1 (6.4–7.5)221.0 (−1.38)

[4] Dissolved silica concentration time series for each of the 66 watersheds were analyzed with linear regression over the time period 1988–2003 (the exception to this is the time period for Staunton and Piney Rivers, where sampling commenced in 1992). Owing to the unavailability of the discharge data in all but five of the watersheds, flow-weighed concentrations were not used in the analysis. Linear regression analysis was chosen over nonparametric alternatives (e.g., Seasonal Kendall Tau [Hirsch and Slack, 1984]) as the residuals of the regressions met normality conditions and the timescales of serial correlation were much less than the overall timeframe of the analysis. A majority of the sites (53 out of 66), exhibited decreasing trends in dissolved silica concentration (Figure 2). The null hypothesis of a median slope of zero was tested by applying a two-sided Wilcoxon signed rank test to the population of slopes. The bias toward a negative median was determined to be statistically significant (p < 0.0001). Given this finding, we seek to determine the geochemical, hydrological, or biological factors that could explain the observed decline in stream water dissolved silica.

image

Figure 2. Trends in dissolved silica concentrations from 1988–2003. Declines are observed in 53 out of the 66 study sites. Note that Piney River (PINE) and Staunton River (STAN) span a more limited timeframe of available data from 1992–2003. Sites with asterisks indicate the six core watersheds in Shenandoah National Park sampled on a weekly basis.

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[5] Previous studies on the temporal dynamics of silica in stream water have examined several processes that control silica concentrations. Each of the following factors could potentially account for the decreasing dissolved silica concentrations.

1.1. Geochemical Factors

[6] Unlike other regions of the northern and eastern U.S. where stream acidity is recovering in response to reduced atmospheric emissions of sulfur dioxide, streams in western Virginia have continued to become more acidic. This has been attributed to sulfur retention in the watershed soils, which cause a delaying effect on stream water chemistry [Webb et al., 2004]. Could changes in soil water acidity have affected rates of chemical weathering in these forested watersheds? Previous studies have reported contradictory results as to whether this could be the case. While April et al. [1986] suggested that weathering rates in Adirondack watersheds increased three-fold since pre-Industrial times due to enhanced acidity, Driscoll et al. [1989] reported remarkably stable concentrations of dissolved silica in the Hubbard Brook watershed over two decades despite increasing pH. Other studies have since supported the assessment that acidification has a minor impact on chemical weathering rates of silicate minerals [Sverdrup and Warfvinge, 1995; Oliva et al., 2003].

1.2. Hydrological Factors

[7] Dissolved silica has been employed as a geochemical tracer to examine hydrological flow paths in headwater catchments, but variations in hydrological conditions make it difficult to discern if silica availability or water-mineral contact time govern concentrations [Scanlon et al., 2001]. Reliance of dissolved silica concentration on contact time is contested, with some research concluding it is nearly independent [Kennedy, 1971; Wels, 1991] and others maintaining that there is some dependence [Buttle and Peters, 1997]. Field studies have shown that infiltrated precipitation quickly attains dissolved silica levels that are equivalent with antecedent soil water [Davis, 1964; McKeague and Cline, 1963; Asano et al., 2003]. Regardless of the influence of contact time, dissolved silica concentrations associated with hydrological flow paths are typically distinct. Groundwater usually has higher dissolved silica concentrations relative to soil water, with this difference attributed to increased availability of silicate minerals for dissolution in bedrock [Asano et al., 2003]. Water that reaches the stream channel through overland flow generally contains the lowest silica concentration relative to both ground and soil water, due to minimal silicate mineral contact [Wels, 1991; Scanlon et al., 2001]. The observed decreases in silica concentrations could possibly be explained by a trend of increasing precipitation, which would lead to greater portions of the stream water deriving from the relatively silica-poor shallow subsurface or overland flow.

1.3. Biological Factors

[8] Still other studies have examined biological processes controlling silica, such as the often underestimated influence of vegetation [Conley, 2002; Derry et al., 2005]. Alexandre et al. [1997] showed that vegetation uptake, storage, and release are significant components of chemical weathering dynamics, while Fulweiler and Nixon [2005] surmised that uptake of silica by terrestrial vegetation was responsible for substantial declines in stream water concentrations in a forested watershed during the onset of the growing season. Shorter-term studies examining in-stream algal growth have cited diatom populations as a cause for seasonal silica concentration declines [Kilham, 1971; House et al., 2001], and in some cases found diatom populations to be Si-limited [Davis et al., 1978; Perkins and Underwood, 2000]. On the basis of a three-year study, Wall et al. [1998] concluded that in-stream biological processes were controlling seasonal variations in nitrate and dissolved silica, and attributed this to uptake by diatoms.

[9] A final biological factor that needs to be considered is disturbance, which has been shown to temporarily increase the concentration of dissolved ions in stream water. Likens et al. [1970] observed an increase in dissolved silica directly following watershed deforestation. An example of disturbance relevant to the current study is the impact by invasive gypsy moth species (Lymantria dispar) which invaded the study area for several years and was responsible for changes in the biogeochemical state of the afflicted watersheds [Webb et al., 1995; Eshleman et al., 1998]. We consider disturbance effects on watershed biological factors as a possible cause for the changing levels of dissolved silica in the western Virginia streams.

[10] This study utilizes an extensive and continuous data set to investigate the observed temporal trends in dissolved silica. The specific objectives of this paper are (1) to relate the magnitude of the dissolved silica trends to specific characteristics of the study watersheds, and (2) to identify the dominant processes that have caused observed decreases in the majority of the study streams. This information will help to better define the biogeochemical processes and rates of chemical weathering in forested Appalachian watersheds.

2. Site Description

  1. Top of page
  2. Abstract
  3. 1. Introduction
  4. 2. Site Description
  5. 3. Methods
  6. 4. Results
  7. 5. Discussion
  8. 6. Conclusions
  9. Acknowledgments
  10. References
  11. Supporting Information

[11] All sites analyzed in this study are located in western Virginia, an area comprised of both the Blue Ridge and the Valley and Ridge physiographic provinces. The bedrock underlying the sites is spatially variable, consisting of Precambrian granites and basalts, and Paleozoic sedimentary units [Webb et al., 1989]. The bedrock dictates the type of soils that form in the region, predominantly ultisols and inceptisols. The study sites are headwater catchments, forested by second-growth oak-hickory cover with stands of hardwoods and pine [Ryan et al., 1989]. The 60 VTSSS sites examined in this study were sampled quarterly over a 16 year period (1988–2003). Their geographic distribution is shown in Figure 1.

[12] The 6 forested watersheds sampled on a weekly basis are all located within the boundary of Shenandoah National Park. SNP encompasses over 780 km2 of the Blue Ridge Mountains, with basaltic and granitic bedrock underlying the northeast and siliciclastic rocks underlying the southwest of the Park [Gathright, 1976]. Average annual precipitation in the Park is 1350 millimeters and the average temperature is 8.4°C [Furman et al., 1998]. Precipitation is relatively evenly distributed on a seasonal basis, and increases with altitude in the Park [Shaffer, 1984]. Precipitation amount and composition are measured by the National Atmospheric Deposition Program at Big Meadows within SNP. Hourly discharge is recorded at the outlets of the core watersheds except for DR01 over the time period corresponding to the stream chemistry observations.

[13] Characteristics of the 6 forested watersheds comprising the core sites in this study are detailed in Table 1 and their geographic distribution is shown in Figure 1. Three of these watersheds, White Oak Run (WOR1), Deep Run (DR01), and Paine Run (PAIN) are located in the southern portion of SNP overlying siliciclastic bedrock. North Fork Dry Run (NFDR) and Staunton River (STAN) are centrally located within SNP and are both predominantly on granitic bedrock and are centrally located within SNP. Piney River (PINE), the northernmost site in the Park and the largest watershed in the study, is the only core site overlying basaltic bedrock.

[14] At the time of Shenandoah National Park's establishment in 1935, the land had been excessively grazed, logged, and farmed [Lambert, 1989]. The Park was permitted to recover naturally, with minimal human influence, and reach 95% reforestation [Conners, 1988]. Disturbance to SNP has occurred in many forms, including atmospheric acid deposition and gypsy moth defoliation. The SWAS program, supported mainly by the National Park Service, has been monitoring the effects of these disturbances on streams in SNP for over 25 years.

3. Methods

  1. Top of page
  2. Abstract
  3. 1. Introduction
  4. 2. Site Description
  5. 3. Methods
  6. 4. Results
  7. 5. Discussion
  8. 6. Conclusions
  9. Acknowledgments
  10. References
  11. Supporting Information

[15] Stream water samples collected in this study are analyzed for temperature, conductivity, alkalinity, and pH, as well as dissolved silica and a suite of major ions. Field and laboratory methods employed by SWAS and VTSSS to measure these variables are described in detail in the Analytical Procedure Manual on the SWAS website (http://swas.evsc.virginia.edu). The six core SWAS sites are the focus of this investigation to determine the cause of decreasing dissolved silica concentrations that have been observed on a more widespread basis throughout western Virginia over the recent sixteen-year period.

3.1. Geochemical Methods

[16] Trends in dissolved silica concentrations were examined for geographical clustering to establish if the overall decline was driven by a particular subset of the western Virginia study region. Underlying bedrock geology for the watersheds was considered as a possible determinant of the dissolved silica behavior. Trends were also compared with mean dissolved silica concentrations in the streams to test for significance in this relationship.

[17] The PHREEQC model [Parkhurst and Appelo, 1999] was implemented to calculate equilibrium SiO2 concentrations in stream water for the six core sites, and the sensitivities of these concentrations were evaluated with respect to pH. Model inputs were mean temperature and primary bedrock mineralogy of the watersheds, while pH was allowed to vary at 0.5 increments over a range of 4–8. The model calculates the expected mineral proportions that are chemically weathered from the watershed bedrock by the stream water and determines equilibrium concentrations of dissolved constituents. The equilibrium SiO2 concentrations determined by PHREEQC were compared with observed concentrations of SiO2 as a function of pH. This information was used to evaluate if changes in soil water or rain water acidity could be responsible for the observed declines in dissolved silica.

3.2. Hydrological Methods

[18] Average annual precipitation and discharge totals from 1988 to 2003 were examined for temporal trends to evaluate if these factors could be related to the observed decline in dissolved silica concentrations. Also, general relationships between dissolved silica concentration and stream discharge were established, and temporal trends in this relationship were accounted for by separating the data into early (1992–1997) and late (1998–2003) time periods. Although earlier data exists for three of the intensive sites, only data from 1992–2003 were used to provide a consistent timeframe and to capture the stream water dynamics over the post–gypsy moth defoliation period.

3.3. Biological Methods

[19] Efforts to relate the trends in stream chemistry to biological factors focused on the impact of the gypsy moth defoliation event that occurred in the early portion of the 1988–2003 timeframe. Stream chemistry data for the White Oak Run (WOR1) watershed, which extend back to 1979, were selected for more detailed analysis and were used to place the current study within a wider temporal context, particularly with regard to the effects of the gypsy moth disturbance on the stream water biogeochemistry. Nitrate concentrations from WOR1 were used to identify the timing of the watershed defoliation. Deposition of insect frass, leaching from damaged foliage, and enhanced litterfall associated with this event caused a dramatic pulse in stream water nitrate levels; this occurrence was taken as the dividing point for separate linear trend analyses performed on the pre- and post- gypsy moth portions of the long-term dissolved silica time series.

[20] From 1987 to 1992, the gypsy moth outbreak caused wide-spread tree deaths that opened up the forest to pioneering plant species. Geographic Information Systems (GIS) analysis using ESRI ArcINFO software and vegetation data obtained from the US Geological Survey produced maps of vegetation and areas of gypsy moth defoliation [Young et al., 2005]. These maps were used to determine if the dissolved silica trends could be correlated with the vegetation composition of the forested watersheds. Various plant species uptake and store silica at different rates and amounts [Fulweiler and Nixon, 2005], and the data from this analysis allows us to examine the possibility that the terrestrial vegetation has influenced the long-term trends in stream-water dissolved silica.

[21] Finally, we evaluated benthic diatoms as a potential factor that could have affected the long-term silica dynamics in these headwater streams. Each of the six core sites was sampled for benthic diatoms in an effort to estimate relative abundance. Nine samples were collected from each of the six streams. Within each stream, cobbles were selected as representative substrates and an equal area (20.3 cm2) of periphyton was scraped from each rock. Three cobbles were selected from three separate riffles along each stream channel in an effort to accurately represent the diatom assemblages that were present. To ensure each sample area on the cobbles was equivalent, the scraped portion was defined with a piece of PVC pipe. The individual samples were preserved in Lugol's Iodine and examined directly in a wet mount with a compound light microscope four weeks after collection. The number of diatom individuals in a 10–50 mL subsample was enumerated and the dominant genera identified. A count of at least 300 individual cells was reached for each subsample, and both live and dead diatoms were recorded. The diatoms were then measured to determine an average area for each genera using phase contrast microscopy.

4. Results

  1. Top of page
  2. Abstract
  3. 1. Introduction
  4. 2. Site Description
  5. 3. Methods
  6. 4. Results
  7. 5. Discussion
  8. 6. Conclusions
  9. Acknowledgments
  10. References
  11. Supporting Information

[22] Sites exhibiting the largest declines in dissolved silica were geographically clustered in the northern region of SNP (Figure 1). This region is primarily underlain by basaltic bedrock, the geological classification associated with the highest mean stream water dissolved silica concentrations. There exists a significant (p < 0.0001) relationship between the means and temporal trends in dissolved silica concentrations, with those watersheds having the highest mean values also exhibiting the greatest declines in dissolved silica through time.

4.1. Geochemistry

[23] Silica concentrations for each of the six core sites plotted as a function of pH indicate that silica concentrations and pH values are clustered with respect to the three main bedrock classifications underlying the sites (Figure 3). The siliciclastic sites, WOR1, DR01, and PAIN, have the lowest concentrations of silica and the lowest pH values in an approximate range of 5 to 6.5. NFDR and STAN, underlain by granitic bedrock, have higher silica concentrations and higher pH values from approximately 6 to 7. PINE, the only site with a basaltic bedrock classification, was observed to have the highest silica concentrations and highest pH values, near neutral.

image

Figure 3. Dissolved silica concentrations as a function of pH for the six core sites. Also shown are the PHREEQC model results for equilibrium concentrations with respect to amorphous silica, quartz, and gibbsite and kaolinite.

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[24] Observed dissolved silica–pH relationships for the six core sites show no dependence of silica concentrations on the stream water acidity. As a confirmation, the results from the PHREEQC model show that modeled equilibrium values of dissolved silica for water in contact with primary bedrock mineralogy (amorphous silica, quartz, gibbsite and kaolinite) exhibit minimal variation with respect to pH from 4 to 8. Observed declines in dissolved silica, therefore, do not seem to be directly attributable to any changes in watershed acidity.

4.2. Hydrology

[25] Linear regression analysis of annual precipitation within SNP shows a slight increase over the study period, although it is not statistically significant (p = 0.45). Also, the time series of annual discharge for the five core sites where discharge was measured showed no significant trends (0.35 ≤ p ≤ 0.99), and the direction of the trends was found to be inconsistent between sites. From 1992 to 2004, discharge decreased for 3 of the 5 core sites (WOR1, STAN, PINE), and increased for the two others (NFDR and PAIN). Simple analyses of the trends in watershed hydrological inputs and outputs were thus unable to explain the decreases in dissolved silica concentrations.

[26] A more detailed examination of the stream discharge and stream chemistry demonstrates that a consistent negative power relationship exists between dissolved silica and discharge for each of the five core sites (Figure 4). Temporal information was included in the analysis by separating the data into early (1992–1997) and late (1998–2003) periods. The general negative power-relationship holds for both the early and late periods, but noticeable differences are observed for concentrations primarily during periods of low discharge (note logarithmic axes). This analysis suggests that the observed declines in dissolved silica concentrations through time can be attributed to dynamics during low-flow conditions. These conditions prevail during the summertime when the relatively silica-rich groundwater is the main contributor to stream discharge. Any explanation for the observed declining dissolved silica levels must be able to account for the fact that concentrations during high flow are similar, while concentrations during low-flow conditions changed during the study period.

image

Figure 4. Log-log plots of concentration-discharge relationships for the five core sites where discharge was measured. Main differences between the early (1992–1997) and late (1998–2003) periods occur during low-flow conditions.

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4.3. Biology

[27] The time series of nitrate levels in WOR1 indicate that 1991 was the year of peak gypsy moth defoliation (Figure 5a). Time series of dissolved silica concentrations over this 25-year period reveal a change in the trend direction coincident with this widespread disturbance event (Figure 5b). DR01, the only other weekly-sampled sites where the stream chemistry record extends prior to the defoliation, also had a positive trend in dissolved silica concentration before the defoliation and a negative trend after the arrival of the gypsy moth (data not shown).

image

Figure 5. Time series of (a) nitrate concentrations and (b) dissolved silica concentrations for White Oak Run (WOR1) from 1979–2003. Gypsy moth defoliation occurred in 1991, causing a pulse in stream water nitrate concentration. The timing of this pulse corresponds to a reversal in the dissolved silica trend.

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[28] GIS analysis of the vegetation coverage within SNP indicated that relative abundances of vulnerable trees such as oak (Quercus) and pine (Pinus), or resistant varieties such as maple (Acer), white ash (Fraxinus americana) and tulip poplar (Liriodendron tulipifera), within the study watersheds were not significantly correlated with their corresponding dissolved silica trends. In terms of geographical distribution of the disturbance, defoliation data from Young et al. [2005] revealed that the gypsy moth progressed from north to south within the region over the time period from 1987 to 1992.

[29] Sampling of benthic diatoms at each of the six core sites revealed a large range of diatom diversity and population sizes. Projected diatom surface density (cm2 cm−2) at each site was calculated by multiplying diatom counts of each genera by their respective mean areas (one-sided) and dividing by the sampling area (Table 2). The two sites with the highest mean stream water silica concentrations, PINE and NFDR, had the highest projected surface densities (0.29 and 0.51, respectively). These two sites also had the greatest diatom diversity, with seven genera identified. Their populations, however, were dominated by synedra, a needle-shaped diatom that has the largest cell size of the diatoms measured in the SNP streams. The other four core sites had significantly lower diatom densities, all below 0.1. Diatom populations for these streams were dominated by smaller genera, specifically fragliara, which were observed in ribbon-like colonies [Round et al., 1990].

Table 2. Diatom Density (One-Sided Diatom Area per Unit Rock Surface Area) Measured Within Each Watersheda
WatershedMean [SiO2], μmol L−1 (SiO2 Trend, μmol L−1 yr−1)Diatom Density (cm2 cm−2)Diatom Genera (% Diatom Surface Area)
  • a

    Also shown are the diatom genera, with their relative abundance noted in terms of percentage of overall surface area.

White Oak Run (WOR1)80.1 (−0.01)0.0738Synedra (36.8%), Fragliara (23.2%), Reimeria (22.9%), Gomphocymbella (16.4%), Cocconeis (0.8%), Cymbella (0.1%)
Deep Run (DR01)94.17 (−0.05)0.0149Fragliara (71.2%), Synedra (16.9%), Gomphocymbella (5.0%), Reimeria (4.9%), Cocconeis (2.1%)
Paine Run (PAIN)88.8 (−0.80)0.0631Fragliara (43.6%), Reimeria (35.0%), Gomphocymbella (10.4%), Synedra (9.7%), Cocconeis (1.3%)
North Fork Dry Run (NFDR)169.9 (−0.48)0.509Synedra (97.5%), Tabellaria (1.2%), Fragliara (0.6%), Gomphocymbella (0.4%), Diatomella (0.2%), Cymbella (0.1%), Reimeria (<0.1%)
Staunton River (STAN)139.4 (−0.82)0.0117Synedra (64.8%), Fragliara (27.5%), Gomphocymbella (7.0%), Cymbella (0.6%)
Piney River (PINE)221.0 (−1.38)0.286Synedra (97.2%), Fragliara (1.2%), Cocconeis (0.6%), Pseudostauosira (0.4%), Gomphocymbella (0.3%), Reimeria (0.2%), Cymbella (0.1%)

5. Discussion

  1. Top of page
  2. Abstract
  3. 1. Introduction
  4. 2. Site Description
  5. 3. Methods
  6. 4. Results
  7. 5. Discussion
  8. 6. Conclusions
  9. Acknowledgments
  10. References
  11. Supporting Information

5.1. Evaluation of Possible Factors Influencing Declines in Dissolved Silica

[30] Bedrock lithology is known to influence the rate of physical and chemical weathering [Bluth and Kump, 1994], and therefore it is not surprising that high silica concentrations are associated with watersheds underlain by the more reactive basaltic and granitic bedrock, while lower concentrations are found in siliciclastic watersheds. What deserves further scrutiny is the overall decline in dissolved silica concentrations in western Virginia streams and the fact that the magnitudes of the trends are related to bedrock geology and mean silica concentration. The geographical analysis provides an empirical context for the dissolved silica trends, however a process-based explanation is sought.

[31] Trends in atmospheric deposition were considered as a possible cause for changes in the chemical weathering rates of the study watersheds. Decreases in acidic deposition have been reported in the eastern U.S. over the past 12–20 years [Stoddard et al., 2001; Skjelkvåle et al., 2005] due to state and federal emission control programs. These decreases, however, do not necessarily yield commensurate improvements in stream water quality. In the Blue Ridge region of Virginia, reduced sulfate deposition has not lead to significant changes in stream water sulfate concentrations or acid-neutralizing capacity, making this region distinct from areas in the northern U.S. that have more recently formed soils [Webb et al., 2004]. The availability of sulfate in watershed soils, which forms sulfuric acid and weathers silica ions from the regolith and underlying bedrock [Rice and Bricker, 1995], has not changed considerably in the western Virginia region due to the large sulfate absorption capacity of the watershed soils.

[32] Even if watershed acidity within the study region had become less acidic, the effects on dissolved silica concentrations would have been minor. Stream water concentrations exhibit little correlation with pH for the six core sites (Figure 3), a finding that was supported by the equilibrium dissolved silica concentrations estimated by the PHREEQC model applied to the dominant bedrock mineralogy. These observational and model results are consistent with Brady and Walther [1989], who reported that dissolution of most silicate forms is independent of pH within naturally-occurring ranges. Trends in atmospheric deposition and watershed acidity are thus not likely to be responsible for the declines in dissolved silica concentrations.

[33] Precipitation has an influence on stream water dissolved silica concentrations through its effect on hydrological flow paths [Scanlon et al., 2001]. In general, water associated with overland or shallow subsurface flow is depleted in dissolved silica relative to groundwater due to reduced contact time and less availability of weatherable material. If precipitation had increased over the study period, groundwater would account for a smaller fraction of stream flow, leading to a decrease in dissolved silica. This was not the case, however, as both precipitation and stream discharge showed no significant trends. Furthermore, dissolved silica concentrations differed between the early and late portions of the time series at like levels of discharge (Figure 4), providing clear evidence that a factor other than hydrology caused the declining trends in dissolved silica.

[34] Analyses of stream chemistry data for the two core watersheds with records extending back to 1979 reveal that dissolved silica concentrations actually increased until a time coincident with gypsy moth disturbance, after which the trend reversed (WOR1 data are shown in Figure 5). The declining portion of the WOR1 record overlaps the majority of the timeframe for the more widespread trend analysis (1988–2003). Rather than geochemical or hydrological factors being responsible for the declining trends in dissolved silica concentration, the extended analysis shown in Figure 5 points to the disturbance event and its legacy as being central to the observed declines.

[35] The gypsy moth, an invasive species native to Europe and Asia, has had a devastating impact on hardwood forests throughout eastern North America [Elkinton and Liebhold, 1990; Davidson et al., 1999]. It infested the northern region of the Virginia survey area beginning in 1986 and reached the southern end by 1991. The gypsy moth attacked several key tree species, including many oaks and hardwoods, causing intense defoliation as well as some tree deaths. Gaps in the forest canopy led to increased growth of understory vegetation, in some cases leading to species replacement [Jedlicka et al., 2004]. Species-specific uptake rates of silica, a key nutrient for plant health, are unknown for the majority of trees present in SNP, as the literature has focused mainly on agricultural plants. In a review of leaf carbon to silica ratios in tree species, Fulweiler and Nixon [2005] report that this ratio for oak is relatively low compared to other sampled species, suggesting that if oak had been replaced on a widespread basis following gypsy moth disturbance, the species-specific uptake rate would not necessarily be enhanced.

[36] Gypsy moth disturbance reduces the overall net primary productivity of the forests during the period of defoliation, followed by a rebound and thus greater silica uptake by the terrestrial vegetation during forest recovery. This does not appear to be responsible for the long-term declines in dissolved silica concentrations following gypsy moth defoliation, however, since tree productivity is impaired for several years at most [e.g., Muzika and Liebhold, 1999] before returning to pre–gypsy moth conditions. An alternative model is needed to describe the trends that are observed over the more long-term, decadal timescales.

5.2. A Conceptual Model: Benthic Diatoms

[37] In keeping with the findings from the above analysis, an adequate conceptual model to account for the observed trends in dissolved silica concentrations must: (1) result in long-term decreased concentration through time, (2) be associated with the gypsy moth defoliation, (3) primarily impact concentrations during low-flow conditions, and (4) explain why streams with highest mean dissolved silica concentrations have experienced the greatest declines through time. Benthic diatoms, which remove silica from the water column in the formation of their shell during growth and are fundamental to the food webs of stream ecosystems [e.g., Mayer and Likens, 1987; Lamberti, 1996], were present in the streambeds at each of the core sites. Here we develop a conceptual model to examine the role of benthic diatoms in affecting long-term trends in stream water dissolved silica.

[38] Gypsy moth defoliation resulted in increased sunlight penetration through the forest canopy, and this increased exposure in streams could have caused photosynthetic diatom populations to expand [Round et al., 1990]. In addition to increased radiation, nitrate levels in streams dramatically increased following gypsy moth defoliation and remained high over a period of many years (Figure 5b). Since freshwater diatoms can be nitrate- or silica-limited [Kilham, 1971; Wall et al.,1998; House et al., 2001], the enhanced nitrate concentration in the streams could have exceeded a threshold such that this limitation on population growth was removed. A steady increase in diatom population would cause a reduction in dissolved silica concentrations in the years following the defoliation event.

[39] For this conceptual model to hold, the statistically significant relationship between mean dissolved silica concentrations and slopes of the dissolved silica trends should be explained. If nitrate was no longer limiting the diatom populations during the post-gypsy moth time period, stream water silica instead could have been limiting. Both silica uptake rate and diatom cell division rate are commonly modeled according to a Michaelis-Menton equation with respect to dissolved silica concentration [Martin-Jezequel et al., 2000]. This implies that streams with higher mean dissolved silica concentrations should also have the largest diatom populations and uptake rates. In fact we see that PINE and NFDR, the two core sites with the highest mean dissolved silica concentrations, also have the highest diatom densities. Beyond this, the relationship between diatom density and mean stream water dissolved silica concentration is not well-defined. A more spatially extensive diatom sampling program would be needed to more rigorously test this hypothesis.

[40] Another possible explanation for the statistically significant mean versus trend relationship is the fact that the gypsy moth defoliation progressed from the north (dominated by basaltic and granitic bedrock) to the south (dominated by siliciclastic bedrock) of the study area during the early portion of the 1988–2003 timeframe. The later arrival of the gypsy moth in the southern watersheds could have reduced the slope of the linear best fits to the dissolved silica time series.

[41] Uptake of dissolved silica by diatoms would more prominently influence low-flow concentrations. This is due to the fact that during these periods there is less overall mass of silica in the water column, and therefore uptake would have a greater affect on the relative amount of mass. For high-flow conditions, although approximately the same amount of mass would be removed, the change in mass would be small compared with the large initial mass of dissolved silica within the water column, and therefore the concentrations would not be greatly impacted. This can be illustrated by a simple conservation of mass model, assuming steady state conditions (and therefore assuming constant residence time with the stream network):

  • equation image

where C is dissolved silica concentration (μmol L−1), Q is stream discharge (m3 s−1), r is diatom dissolved silica uptake rate (μmol s−1), and the 1000 is a conversion factor (L m−3). This assumes a power-law relationship between concentration and discharge for flow into the stream network, described by the coefficient a and the exponent b. Results of this simple model are shown for Staunton River in Figure 6, depicting the influence of in-stream uptake rate on the concentration-discharge relationships. Values of uptake rate, r, were chosen on the basis of fitting the model to the data for the early and late portions of the time series. Literature values, specific to the diatom genera and environmental conditions of the stream, are not available as an independent confirmation of these fitted estimates. This simple model suggests that increases in uptake rate by the diatom community could lead to decreases in stream concentrations, most notably during low-flow conditions.

image

Figure 6. Concentration-discharge plot for Staunton River, with the time series divided into early (1992–1997) and late (1998–2003) periods. Also shown are the modeled effects of in-stream uptake of dissolved silica by diatoms. Note that diatom uptake more strongly influences concentrations during low-flow conditions.

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6. Conclusions

  1. Top of page
  2. Abstract
  3. 1. Introduction
  4. 2. Site Description
  5. 3. Methods
  6. 4. Results
  7. 5. Discussion
  8. 6. Conclusions
  9. Acknowledgments
  10. References
  11. Supporting Information

[42] Dissolved silica in western Virginia streams from 1988–2003 showed a significant bias toward decreasing concentrations through time. Stream chemistry analyses on two watersheds with records dating back to 1979 indicate that this trend is not representative of a more long-term phenomenon. Rather, the observed declines in dissolved silica concentrations appear to be associated with a gypsy moth defoliation that occurred in the study watersheds during the early portion of the 1988–2003 time series. In-stream uptake of silica by benthic diatoms is the most likely explanation for the decreasing concentrations. The findings reported here illustrate long-lasting effects of disturbance on watershed biogeochemistry, evident for dissolved silica which is typically considered to be stable stream water constituent over decadal timescales. Over longer timescales, periodic disturbances affecting the forested watersheds would hypothetically lead to alternating increasing and decreasing trends in dissolved silica concentrations (see Figure 5b) as diatom populations grow, eventually die off, and the diatom shells dissolve in the stream water. Mobilization of diatom frustules is not accounted for in the present study, as the monitoring efforts focused solely on the dissolved portion of the stream water silica. It should be noted, however, that this biologically derived particulate matter could be significant in terms of defining the overall watershed silica budget.

[43] Although the presence of diatoms has the greatest impact on stream water dissolved silica concentrations during periods of low discharge, this is significant enough to affect overall dissolved silica fluxes. Like earlier studies that have underscored the importance of considering internal biological cycling in estimating chemical weathering rates in watersheds [e.g., Conley, 2002; Derry et al., 2005], this research confirms that biological factors do indeed play a significant role in governing silica flux, but also shows that biological uptake can be a nonstationary process. Long-term monitoring of stream chemistry is needed to capture the transient effects of biological processes and to place observed trends into a wider temporal context.

Acknowledgments

  1. Top of page
  2. Abstract
  3. 1. Introduction
  4. 2. Site Description
  5. 3. Methods
  6. 4. Results
  7. 5. Discussion
  8. 6. Conclusions
  9. Acknowledgments
  10. References
  11. Supporting Information

[44] This research is a contribution to the Shenandoah Watershed Study and the Virginia Trout Stream Sensitivity Study. Funding and support for these programs has been provided by the National Park Service, the U. S. Environmental Protection Agency, the U.S.D.A. Forest Service, the Virginia Department of Game and Inland Fisheries, and Trout Unlimited. Susie Maben and Frank Deviney were responsible for the stream water sampling and the laboratory analyses. Janet Herman provided help with the geochemical modeling and Linda Blum provided instruction and lab facilities for the diatom survey. Jack Cosby and Rick Webb provided many helpful comments in the preparation of this manuscript. Comments from the Associate Editor and two anonymous reviews were extremely valuable and are appreciated.

References

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  2. Abstract
  3. 1. Introduction
  4. 2. Site Description
  5. 3. Methods
  6. 4. Results
  7. 5. Discussion
  8. 6. Conclusions
  9. Acknowledgments
  10. References
  11. Supporting Information
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Supporting Information

  1. Top of page
  2. Abstract
  3. 1. Introduction
  4. 2. Site Description
  5. 3. Methods
  6. 4. Results
  7. 5. Discussion
  8. 6. Conclusions
  9. Acknowledgments
  10. References
  11. Supporting Information
FilenameFormatSizeDescription
jgrg136-sup-0001-t01.txtplain text document1KTab-delimited Table 1.
jgrg136-sup-0002-t02.txtplain text document1KTab-delimited Table 2.

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