There has been increased salinization of fresh water over decades due to the
use of road salt deicers, wastewater discharges, saltwater intrusion,
human-accelerated weathering, and groundwater irrigation. Salinization can
mobilize bioreactive elements (carbon, nitrogen, phosphorus, sulfur)
chemically via ion exchange and/or biologically via influencing of microbial
activity. However, the effects of salinization on coupled biogeochemical
cycles are still not well understood. We investigated potential impacts of
increased salinization on fluxes of bioreactive elements from stream
ecosystems (sediments and riparian soils) to overlying stream water and
evaluated the implications of percent urban land use on salinization
effects. Two-day incubations of sediments and soils with stream and
deionized water across three salt levels were conducted at eight routine monitoring
stations across a land-use gradient at the Baltimore Ecosystem Study
Long-Term Ecological Research (LTER) site in the Chesapeake Bay watershed.
Results indicated (1) salinization typically increased sediment releases of
labile dissolved organic carbon (DOC), dissolved inorganic carbon (DIC),
total dissolved Kjeldahl nitrogen (TKN) (ammonium
Salt concentrations in freshwaters are rapidly increasing at a regional scale in the USA and worldwide (e.g., Nielsen et al., 2003; Kaushal et al., 2005, 2014a; Rengasamy, 2006; Findlay and Kelly, 2011; Steele and Aitkenhead-Peterson, 2011; Corsi et al., 2015). Most of the increased salinization can typically be attributed to road salt deicers and other industrial uses, wastewater discharges, groundwater irrigation, saltwater inundation caused by sea-level rise, and human-accelerated weathering (e.g., Findlay and Kelly, 2011; Aitkenhead-Peterson et al., 2009; Ardón et al., 2013; Kaushal et al., 2013). Increased salinization can have important environmental consequences for drinking water supplies, freshwater biodiversity, degradation of soils and groundwater, degradation of vehicles and infrastructure, and mobilization of inorganic and organic contaminants (Nielson et al., 2003; Kaushal et al., 2005; Findlay and Kelly, 2011; Corsi et al., 2015). Moreover, salinization is difficult if not impossible to reverse and, thus, remediation is unlikely. Recent studies have further shown that increased salinization can influence biogeochemical cycles of bioreactive elements such as carbon and nitrogen (Green et al., 2008, 2009a, b; Green and Cresser, 2008; Compton and Church, 2011; Lancaster, 2012; Steele and Aitkenhead-Peterson, 2013) as well as phosphorus and sulfur (Nielson et al., 2003; Kulp et al., 2007; Compton and Church, 2011; Kim and Koretsky, 2011, 2013). Chemically, salinization affects mobilization of these bioreactive elements through its direct influences on ion exchange and sorption capacity of sediments/soils (e.g., for ammonium and soluble reactive phosphorus (SRP)), as well as via indirect effects due to changes in pH and sodium-induced dispersion (e.g., for dissolved organic carbon (DOC)) (Nielsen et al., 2003; Green et al., 2008; Compton and Church, 2011; Ardón et al., 2013). Biologically, salinization can be a stressor to some microorganisms in fresh water but may also enhance the activities of other microorganisms due to nutrient releases (Kulp et al., 2007; Srividya et al., 2009; Kim and Koretsky, 2011, 2013). Evidence is accumulating that increased salinization is an important process during the urban evolution of geochemical cycles in watersheds globally from decades to centuries (Kaushal et al., 2014a, 2015). Furthermore, salinization has significant ecosystem effects over broader spatial and temporal scales (e.g., Findlay and Kelly, 2011; Kaushal and Belt, 2012; Corsi et al., 2015).
Although there has been increasing research, more work needs to be done regarding the effects of increased salinization on coupled biogeochemical cycles. Prior studies have commonly investigated the effects of salinization on fluxes and transformations of individual bioreactive elements (e.g., Green et al., 2008, 2009a, b; Green and Cresser, 2008; Compton and Church, 2011; Kim and Koretsky, 2011; Lancaster, 2012; Steele and Aitkenhead-Peterson, 2011). However, biogeochemical cycles of bioreactive elements are generally linked in sedimentary diagenesis (Middelburg and Levin, 2009), and transformations of nitrogen, phosphorus, and sulfur are highly dependent on availability of organic carbon (Duan and Kaushal, 2013). For example, organic carbon provides an energy source for microbes responsible for biogeochemical transformations (e.g., Newcomer et al., 2012), and decomposition of organic carbon can facilitate certain redox reactions of bioreactive elements including denitrification, iron reduction and release of soluble reactive phosphorus, and sulfate reduction (Sobczak et al., 2003; Middelburg and Levin, 2009). Regarding organic carbon, prior studies generally investigated the effects of salinization on bulk concentrations of dissolved organic carbon (DOC). However, DOC within aquatic systems consists of not a single compound but a broad suite of organic molecules of varied origin and composition, which may respond differently to salinization. Until recently, relatively little work has been done to improve our conceptual understanding of the effects of salinization on coupled biogeochemical cycles.
Previous studies have shown freshwater salt concentrations vary across land use, with the highest concentrations of salt occurring in urban watersheds (e.g., Kaushal et al., 2005). Green et al. (2008, 2009) reported that soils in urban watersheds that have already experienced exposure to road salting were less responsive to salinization in DOC release than unexposed soils in rural areas. However, urbanization may increase stream sediment organic matter (Duan and Kaushal, 2013) via algal and wastewater inputs (Daniel et al., 2002; Kaushal et al., 2014a) and influence other physical, chemical, and biological characteristics of stream ecosystems (Paul and Meyer, 2001). Despite these two competing impacts of urbanization, biogeochemical impacts of salinization across watershed land use are still less recognized. Moreover, most current studies regarding the effects of salinization focus on soils or anaerobic lake sediments. Very little work has been done to examine stream sediments that may be exposed to high salt concentrations under more aerobic conditions. It is known that stream sediments and soils differ in particle size, structure, and organic matter composition and sources (e.g., Hedges and Oades, 1997). Thus, insights learned from studying the biogeochemical effects of salinization in soils may not always directly apply to stream sediments.
Our primary objective was to investigate the effects of increased salinization on potential fluxes (release or retention) of bioreactive elements (carbon, nitrogen, phosphorus, sulfur) from stream ecosystems and how the effects of salinization change with watershed land use and/or stream substrates (sediments and riparian soils). Sediments and riparian soils collected from sites across a rural–urban land-use gradient were incubated in salt solutions to mimic the effects of runoff with high levels of road salt deicers. Changes in water chemistry were monitored as a function of salt concentrations and land use. Three hypotheses were tested: (1) the effect of salinization on soil leaching and sediment retention/release of bioreactive elements change with watershed urbanization; (2) retention/release of nitrogen, phosphorus, and sulfur in response to salinization can be abiotically and/or biologically coupled with carbon biogeochemistry; and (3) salinization effects on release/transformation of bioreactive elements vary between stream sediments and riparian soils. We expected significant release of organic carbon and coupled transformations with nitrogen, phosphorus, and sulfur as salinity increased. Additionally, we expected experimental salinization increases biogeochemical fluxes with increasing watershed urbanization due to more carbon availability. An improved understanding of the effects of increased salinization on release/retention of bioreactive elements can contribute to our understanding of urban drivers of changes in water quality, microbial communities, and ecosystem functions (Kaushal and Belt, 2012; Kaushal et al., 2014a) and improve water quality by benefitting our assessment and management of salt use.
Characteristics of study subwatersheds.
Watershed land cover and impervious surface cover (ISC %) data are from Shields et al. (2008) and the National Land Cover Database (NLCD) of 2006. Both land cover and impervious statistics were based on 30 m resolution land cover data.
Surface sediments from stream channels and top soil in riparian zones were
collected from eight long-term monitoring sites across a rural–urban land-use
gradient. All eight sites are routinely sampled as part of the Baltimore Ecosystem Study (BES) Long-Term
Ecological Research (LTER) site supported by the US National
Science Foundation. Land-use varies from forest to low-density
residential, agricultural, to suburban and urban (Table 1). The main focal
watershed of the BES LTER site is the Gwynns Falls (GFGR), a 17 150 ha watershed in
the Piedmont physiographic province that drains into the northwest branch of
the Patapsco River that flows into the Chesapeake Bay (Fig. 1). The Gwynns
Falls sites traverse a rural–suburban to urban gradient from Glyndon (GFGL),
Gwynnbrook (GFGB), Villa Nova (GFVN) to Carroll Park (GFCP) (Table 1). An
agricultural stream (MCDN) is a small tributary to the Gwynns Falls draining
a watershed dominated by row crop agriculture (corn, soybeans), while Dead
Run (DRKR) is an urbanized tributary of the Gwynns Falls between GFVN and
GFCP. Samples were also taken from a small urban tributary to the Gwynns
Falls, approximately 700 m above GFCP, which is highly contaminated
with sewage (Kaushal et al., 2011). Baisman Run (BARN) is a low-density
residential watershed located in the nearby Gunpowder Falls watershed that
drains primarily forest land cover (Table 1; Fig. 1). The BES LTER site
provides access to extensive background information and long-term monitoring
of major anions, nutrients, and carbon concentrations and fluxes in streams
(
Land use of the Gwynns Falls and Baisman Run watersheds,
showing sites from which sediment, soil, and stream water were collected for
salinization experiments. Baisman Run is a watershed with forest as the
dominant land use, and it is located in the nearby Gunpowder River. Solid
and open circles represent sites of the main stem and tributaries,
respectively. The resolution of the land-use data is 30 m, and land-use and
stream channel location data are from the US Department of Agriculture
(
Stream water, sediments, and soils for laboratory salinization experiments
were collected on 8 March 2013, 1 day before a snow storm in the
Baltimore–Washington D.C. metropolitan region. Three liters of stream water
were collected at each of the eight sites for the experiments and water quality
analyses. The surface sediments and top soils (approximately 15 cm) were
collected at these same sites with a shovel. Sediment samples were taken
simultaneously along four cross-sections perpendicular to stream flow within 50 m
of the primary sampling site (Duan and Kaushal, 2013). Along each stream
cross section, surface sediments at three sites (left, middle, and right)
were collected. All sediments collected at these sites were well mixed to
make a composite sample. Soil samples from the riparian zone were also
collected similar to sediment samples. Because the sites GFCP and GFGR were
located very close to each other, only one composite soil sample was
collected to represent these two sites. So, laboratory salinization
experiments with soils were conducted at seven rather than eight sites. The sediment
and soil samples were transferred to glass jars and placed immediately into
a cooler and brought back to lab. In the lab, sediments were sieved through
a 2 mm sieve, and the < 2 mm fractions of sediments and soils were
homogenized for incubation experiments (e.g., plant roots were picked from
soils and discarded). The homogenized sediments and soils were sampled for
determination of ash-free dry weight (AFDW). In addition, approximately 100 mL
aliquots of steam water were filtered through pre-combusted GF/F Whatman
filters, and the filtrates were used for water quality analyses. The
filtrates were stored in a refrigerator for analyses of optical properties
and dissolved inorganic carbon (DIC) measurements. Another aliquot was
similarly filtered but frozen prior to analyses of DOC, nutrients, and major anions. The remainder of the stream water,
sediments, and soils were temporarily stored at 2–4
For each laboratory salinization experiment, 60 g sub-samples of homogenized
sieved sediments (< 2 mm) were inserted into a series of 125 ml glass
flasks to cover the bottom of the flasks, and 100 mL of unfiltered stream
waters were carefully added with a pipette in order to not disturb the
sediments. In order to evaluate the potential effects of salinization, pure
NaCl salt (J. T. Baker) was amended to unfiltered stream water to obtain
three
concentration levels (0, 2, and 4 g Cl L
All filtrates were analyzed for major forms of bioreactive elements –
nitrate, total dissolved nitrogen (TDN), SRP,
sulfate, DOC, DIC, and optical properties of DOC
(absorbance and fluorescence). DOC, TDN, and DIC concentrations were measured
on a Shimadzu total organic carbon analyzer (TOC-V CPH/CPN) (Duan and Kaushal
2013). Nitrate and sulfate concentrations were measured with a Dionex ion
chromatograph (ICS-1500, Dionex INC., USA), with an eluent of 3.5 mM of
Na
Water chemistry, sediment, and soil ash-free weight prior to salinization incubation experiments.
– is used when samples were not taken. DOC, P
AFDW of the sediment and soil samples was analyzed as
an index of organic matter content. Sediment and soil ash weights were
calculated as the difference in weights before and after combustion at
550
Sediment fluxes were calculated as the net changes in the concentrations of DOC, nitrate, SRP, or sulfate during the 2-day incubations. The values for nitrate and sulfate are presented as nitrate-N and sulfate-S. The changes in the control flasks (with water only), occurring in water without sediments or soils, were subtracted to obtain the fluxes that were released from sediments or soils. Positive or negative values represent net release from sediments or retention by sediments, respectively.
Effects of salinization on sediment/soil biogeochemical fluxes were examined
by performing linear regressions of these fluxes with salinity, using data
from six salinization experimental manipulations (three salinity levels with
duplicates). If the
In stream waters that were used for laboratory salinization experiments,
water chemistry varied considerably (Table 2). In general, concentrations of
chloride ion (Cl
Sediment AFDW also displayed an increasing trend with
ISC (from 0.61 to 1.90 %) except one surprisingly high value
(3.98 %) observed at GFCP (
Changes in DOC, DIC, protein-like fluorophore, humic-like
fluorophore, protein to humic (P
Releases of DOC and DIC and changes in specific UV absorption
(SUVA) with salinization (Cl
Changes in salinity effects on DOC, DIC, protein-like fluorophore,
P
Changes in TKN (DON
Sediments were consistently a net source of both DOC and DIC. Net DOC
releases from sediments consistently increased with increasing salinization
(all positive slopes and statistically significant in seven out of eight cases;
Moreover, the effects of salinization on sediment net releases differed among
DOC fractions. Salinization consistently and considerably increased net
releases of the protein-like fluorophore (all positive slopes;
Effects of laboratory salinization on net DOC and DIC releases from soils
were relatively more complex and not as consistent (both positive and
negative slopes). As mentioned earlier, laboratory salinization experiments
with soils were conducted only at seven sites, because GFCP and GFGR are very
close were considered as one site for soil experiments (same below). In four out
of seven cases, net DOC releases from soils decreased as experimental
salinization increased from 0 to 2 g Cl L
Effects of laboratory salinization experiments on biogeochemical carbon
fluxes from sediments (indicated by changes in their standardized fluxes per
g of Cl
Sediments were generally a net source of TKN (ammonium
Release of TKN (DON
Similar to sediments, salinization consistently increased net TKN releases
from soils and the increases were significant at six out of seven sites
(
Effects of laboratory salinization on sediment biogeochemical fluxes of TKN
(indicated by changes in their standardized fluxes per g of Cl
Correlations between
DIC, DOC, TKN, SRP, and SUVA stand for dissolved inorganic carbon, dissolved organic carbon, total Kjeldahl nitrogen, soluble reactive phosphorus, and specific ultraviolet absorption.
Correlation analyses suggested that there were links of the fluxes of the
measured chemical species of bioreactive elements. Here, the term flux was
used to mean net retention or net release of a chemical species based on
site. For example, there was a correlation between net releases of DIC flux
and net releases of DOC. Across soil laboratory salinization experiments, DIC
net releases linearly increased with DOC releases, and the correlations were
significant at four out of seven sites (
There was an inverse relationship between nitrate fluxes (release in soils or
retention in sediments) and DOC fluxes from sediments and soils.
Specifically, nitrate fluxes linearly decreased with increasing DOC fluxes
from sediments and soils, and the increases were statistically significant at
six of eight sites for sediment incubation experiments and at four out of seven sites for
soil leaching experiments (
Our study shows that the effects of salinization on retention and release of
bioreactive elements in sediments changes with watershed urbanization. Thus,
hypothesis 1 regarding changes in salinization effects with watershed
urbanization was partially supported by the data from sediment incubation
experiments. Overall, our results suggest that the effects of increased
salinization on sediment releases of DOC, protein-like fluorophore, TKN, and
DIC increased with ISC (Fig. 4; linear regressions, all
Instead, the interactive effects of watershed urbanization and salinization on sediment releases of DOC, protein-like fluorophore, TKN, and DIC fluxes may be explained by coinciding changes in stream sediment organic matter content (indicated by ash-free dry weight), which also showed an increase with increasing watershed ISC (Table 2 and Fig. 4f). The outlier site GFCP, which had unexpected larger salinization effects, was also highest in sediment ash-free dry weight. The reason for the outlier GFCP is not clear, but a much better correlation between sediment ash-free dry weight and watershed ISC was reported in our previous study at the same Baltimore LTER sites (Duan and Kaushal et al., 2013). In any case, organic matter content in urban stream sediments was generally higher than in rural streams (also reported in Paul and Meyer, 2001), probably due to additional organic matter inputs from algal (Kaushal et al., 2014b) and anthropogenic sources (e.g., wastewater; Daniel et al., 2002). Our recent work showed that gross primary production and organic matter lability increased with watershed urbanization (Kaushal et al., 2014b). Wastewater inputs from sewer leaks are common in the urban tributaries in the lower Gwynns Falls (DEPRM and Baltimore City Department of Public Works, 2004; Kaushal et al., 2011). As quantity and quality of sediment organic matter increase across the rural–urban land-use gradient, we hypothesize that the releases of labile DOC (indicated by protein-like fluorophore), total DOC, TKN, and DIC increase in response to salinization.
This study suggests that mechanisms responsible for salinization effects on
DOC mobilization differ between soils and sediments, as we proposed in
hypothesis 3. Previous studies have shown contrasting effects of salinization
(e.g., suppression or inconsistent effects) on DOC mobilization in soils
(Amrhein et al., 1992; Evans et al., 1998; Green et al., 2008, 2009a; Compton
and Church, 2011; Ondrašek et al., 2012). These variations were
attributed to soil types (Amrhein et al., 1992; Evans et al., 1998), water to
soil ratios (Amrhein et al., 1992), water chemistry (Evans et al., 1998),
leaching time (Compton and Church, 2011), and historical exposure to road
salt deicers (Green et al., 2008, 2009a). Two competing effects of salts have
been suggested upon which solubilization of organic matter is dependent:
sodium dispersion and pH suppression (Amrhein et al., 1992; Bäckström
et al., 2004; Green et al., 2008, 2009a). That is, upon salt additions, the
replacements of Ca
Our results from DOC characterization can provide further information for
interpreting the differences in salinization effects on DOC releases between
sediment and soils. Our results showed that only protein-like fluorophores
were consistently and considerably remobilized from sediments with
salinization (Fig. 2), which suggested that the increased DOC releases from
sediments were mainly attributed to the releases of protein-like (or labile)
fractions. Similar findings were also reported by Li et al. (2013), which
showed that KCl can considerably increase the mobility of microbially derived
labile organic matter (indicated by the fluorescence index). Meanwhile,
chemical analyses suggest that the protein-like fluorophores consist
primarily of proteinaceous materials (e.g., proteins and peptides; Yamashita
and Tanoue, 2003; Maie et al., 2007), and this DOC fraction is generally
hydrophilic and low molecular weight (LMW) compounds (e.g., Sommerville and
Preston, 2001). Results of Chen et al. (1989) and Fuchs et al. (2006) showed
that solubility of the proteinaceous materials in LMW is generally
affected by neither pH within the normal range of 6–9 nor colloid coagulation. Therefore,
increasing ionic strength (or salinization) can enhance the solubility of the
proteinaceous materials via sodium dispersion (Green et al., 2008, 2009a)
or through nonspecific electrostatic interactions at low salinities (called a
“salting-in” effect) (Tanford, 1961; Chen et al., 1989). Furthermore,
because stream sediments are generally enriched in these labile,
proteinaceous materials derived from biofilms (algae and microbes) and
wastewater organics in urban watersheds (Daniel et al., 2002; Kaushal et al.,
2011, 2014b; Kaushal and Belt, 2012; Newcomer et al., 2012; Duan et al.,
2014b), it is reasonable that salinization can mobilize a
large amount of protein-like labile dissolved organic matter from stream
sediments. Relative to the proteinaceous materials, humic substances are
larger hydrophobic molecules occurring in the colloidal size range (e.g.,
Aiken et al., 1985). This DOC fraction is readily subjected to flocculation
(e.g., Sholkovit, 1976), sorption to mineral surfaces (Fox, 1991; Hedges and
Keil, 1999), and pH suppression (Kipton et al., 1992; Li et al., 2007, 2013)
with increasing ionic strength or salinization. The potential
instability of the humic-like DOC fraction upon salinization was further
supported by our present results and previous studies (e.g., Li et al.,
2013), which showed that salinization consistently decreased SUVA of DOC
released from soils and sediments (Figs. 2 and 3) – a parameter indicating DOC
aromaticity (Weishaar et al., 2003). Relative to stream sediments, soil
organic matter consists primarily of humic substances (up to 60–70 %;
Griffith and Schnitzer, 1975). In our study, although humic substances were
not measured, much higher SUVA values were observed in DOC leached from soils
(around 10 Lmg
Our laboratory experiments suggests that simultaneous net releases of DIC and
DOC were examples of coupled biogeochemical cycles in response to
salinization, as predicted in hypothesis 2. The effects of salinization on
DIC fluxes from sediments and soils may involve shifts in carbonate chemistry
(e.g., dissolution of carbonate minerals) or organic carbon mineralization
and CO
We observed consistent mobilization of TKN (e.g., NH
Our results further suggest that nitrate transformation and DOC
remobilization were also coupled during salinization experiments with both
sediments and soils, which supports hypothesis 2. In contrast to DOC, DON, or
NH
Our results suggest that SRP release from sediments or soils during laboratory salinization experiments was associated with changes in DOC aromaticity (indicated by SUVA), supporting hypothesis 2. Different from N, P is primarily a particle reactive element, and a large fraction of dissolved P (e.g., up to 88 %; Cai and Guo, 2009) is in the form of colloids or humic Fe (Al-) phosphate complexes (Hens and Merckx, 2001; Turner et al., 2004; Regelink et al., 2013). This is because the sorption capacity for SRP per unit mass is about 5000 times larger for colloids than for the immobile soil matrix (McGechan and Lewis, 2002). However, the stability of colloids decreases with increasing ionic strength and decreasing pH (e.g., Bunn et al., 2002; Saiers et al., 2003), both of which can be induced by salinization (e.g., Green et al., 2008). An example of this salinization effect is rapid flocculation of freshwater SRP and colloids in estuaries in response to mixing of fresh water with seawater (e.g., Sholkovit, 1976). Thus, increased salinization may decrease stability of the colloidal humic Fe (Al-) phosphate complexes, leading to reduced releases of SRP from sediments and soils (Figs. 5 and 6) and a coupling between SRP and SUVA (Table 3). However, biological controls such as temporary inhibition of microbial activity at higher salinities could provide an alternative explanation (Srividya et al., 2009). An explanation for the increased releases of SRP from sediments at higher salinities at two urban sites (GFGR and GFCP) is not clear. Probably, these SRP increases was related to the release of large amounts of labile DOC and resulting changes in redox condition favorable for SRP release. In summary, decreases in SRP fluxes from sediments and soils in response to salinization were likely a result of colloid coagulation and microbial inhibition at higher salinities, while the increases in SRP fluxes from sediments at urban sites warrant further evaluation within the context of redox changes.
A conceptual diagram summarizing potential effects of salinization on DOC quality, DOC and TKN releases from sediments and soils, as well as linkage to release/retention of DIC, nitrate, and SRP during sediment and soil salinization.
Relative to C, N, and P, effects of salinization on sulfur transformations are relatively less known. Kim and Koretsky (2011, 2013) reported salinization inhibited porewater sulfate reduction in lake sediments. However, our results show large variability in the effects of salinization on net sulfate release from either sediments or soils (Figs. 5 and 6), and sulfate reduction seems to not be dominant in free-flowing streams. Effects of increased salinization on sulfate releases warrant further investigation in future studies, however.
The potential effects of salinization on biogeochemical fluxes from soils and
stream sediments are summarized in Fig. 7. As shown in this figure, releases
of labile DOC (thus total DOC) and TKN (primarily DON and ammonium) from
sediments can potentially increase during episodic stream salinization, due
to “salting-in” effects (or Na dispersion) of proteinaceous organic matter
and NH
This research was supported by NSF Awards EAR 1521224, DBI 0640300, CBET 1058502, EAR-1426844 and DEB-0423476, NASA grant NNX11AM28G, and Maryland Sea grant awards SA7528085-U and R/WS-2.Edited by: T. J. Battin