www.biogeosciences.net/9/2697/2012/ doi:10.5194/bg-9-2697-2012 © Author(s) 2012. CC Attribution 3.0 License.

Abstract. Summer cyanobacterial blooms caused an elevation in pH (9 to ~10.5) that lasted for weeks in the shallow and tidal-fresh region of the Sassafras River, a tributary of Chesapeake Bay (USA). Elevated pH promoted desorption of sedimentary inorganic phosphorus and facilitated conversion of ammonium (NH4+) to ammonia (NH3). In this study, we investigated pH effects on exchangeable NH4+ desorption, pore water diffusion and the flux rates of NH4+, soluble reactive phosphorus (SRP) and nitrate (NO3−), nitrification, denitrification, and oxygen consumption. Elevated pH enhanced desorption of exchangeable NH4+ through NH3 formation from both pore water and adsorbed NH4+ pools. Progressive penetration of high pH from the overlying water into sediment promoted the mobility of SRP and the release of total ammonium (NH4+ and NH3) into the pore water. At elevated pH levels, high sediment-water effluxes of SRP and total ammonium were associated with reduction of nitrification, denitrification and oxygen consumption rates. Alkaline pH and the toxicity of NH3 may inhibit nitrification in the thin aerobic zone, simultaneously constraining coupled nitrification–denitrification with limited NO3− supply and high pH penetration into the anaerobic zone. Geochemical feedbacks to pH elevation, such as enhancement of dissolved nutrient effluxes and reduction in N2 loss via denitrification, may enhance the persistence of cyanobacterial blooms in shallow water ecosystems.


Introduction
Nutrient releases from sediment into the water column can support a substantial fraction of the primary production in shallow coastal and estuarine ecosystems (e.g.North Carolina estuaries, Fisher et al., 1982;Potomac River Estuary, Kemp and Boynton, 1984;Baltic Sea, Koop et al., 1990; Chesapeake Bay, Cowan and Boynton, 1996).Enhanced nitrogen and phosphorus fluxes may promote high levels of phytoplankton biomass (Kemp et al., 2005).Such phytoplankton blooms lead to the sustained accumulation of phytodetritus in sediment, fueling nutrient recycling through organic matter remineralization (Cowan and Boynton, 1996;Nixon et al., 1996).The consequences, such as decreased water clarity, depletion of bottom-water oxygen and the decomposition of phytodetritus, may enhance sediment respiration, decrease redox potential, limit nutrient uptake by benthic microalgae, and generally increase nutrient fluxes (Kemp et al., 2005).
In the deep, anoxic region of the Chesapeake Bay and other estuaries, phosphorus flux is usually promoted by the dissolution of Fe-oxides and their conversion to iron sulfides (Cornwell and Sampou, 1995;Roden and Tuttle, 1992).The increase in ammonium release from sediments tends to coincide with inhibition of nitrification by oxygen depletion and generation of reductants (HS − /H 2 S) in sediment, which consequently constrain denitrification (Kemp et al., 2005;Cornwell et al., 1999).Nevertheless, in oxic shallow water ecosystems, benthic nutrient releases are generally less redox influenced.
Driven by rapid utilization rates of inorganic carbon for photosynthesis during dense algal blooms (Hansen, 2002; Y. Gao et al.: Effects of cyanobacterial-driven pH increases Hinga, 2002), persistent high pH in shallow water can influence benthic dynamics with progressive pH penetration from the overlying water into sediments (Bailey et al., 2006).When pH is above a critical threshold (9-9.2),inorganic phosphorus desorbs from iron oxides at mineral surfaces (Andersen, 1975;Eckert et al., 1997).Elevation of pore water pH (∼9.8 in tidal-fresh regions, Eckert et al., 1997) can promote sediment release of soluble reactive phosphorus (SRP), and simultaneously support P demand during cyanobacterial blooms in lakes (Xie et al., 2003) and tidal fresh and oligohaline estuaries (Seitzinger, 1991;Andersen, 1975).
In contrast to pH-driven P cycling, the effects of pH on sediment N transformations and release are not well understood (Soetaert et al., 2007).During the decomposition of sediment organic matter, remineralized ammonium (NH + 4 ) may both adsorb onto sediment mineral surfaces and accumulate in pore water.Exchangeable NH + 4 is weakly bonded to negatively charged particle surfaces, buffering pore water NH + 4 concentrations (Rosenfeld, 1979).In estuaries, fine grained sediment generally has a large pool of adsorbed NH + 4 (Wang and Alva, 2000;Weston et al., 2010), with freshwater sediments having considerably more adsorbed ammonium relative to saline sediments (Seitzinger, 1991).Once alkaline pH results in the conversion of NH + 4 to dissolved ammonia (NH 3 ), formation of non-ionized NH 3 may decrease NH + 4 cation adsorption on sediments and potentially alter the balance between pore water and exchangeable NH + 4 .Besides assimilation by plants and bacteria, remineralized N can diffuse/advect from sediment into the overlying water.Alternatively, part of the NH + 4 can be oxidized sequentially to NO − 2 /NO − 3 and then reduced to N 2 in the suboxic/anoxic layer (Cornwell et al., 1999).However, shifts in the NH + 4 -NH 3 equilibrium, associated with high pH penetration passing through the redox boundary, may change the rates of pore water diffusion and nitrification-denitrification.In the tidalfresh and oligohaline parts of the Potomac River (Chesapeake Bay), Seitzinger (1987) observed both increased SRP and NH + 4 fluxes with pH elevation.Experimental NH + 4 flux rates increased from <10 to over 100 µmol m −2 h −1 when pH was raised from 8 to 10 in laboratory incubations (Seitzinger, 1987).During an algal bloom in the Potomac estuary, Bailey et al. (2006) observed a three-fold increase of NH + 4 efflux when the bottom water pH rose from neutral to above 9.In soil studies, the combined influence of alkaline pH (>8) and toxic NH 3 production can reduce the NH + 4 soil inventory due to NH 3 volatilization, decrease the efficiency of nitrification and denitrification, and inhibit enzyme activity as well (Simek et al., 2002;Cuhel et al., 2010).
We hypothesize that increased sediment pH facilitates not only P desorption but also the conversion of NH + 4 to NH 3 with consequent changes in sediment N cycling.In this study we examined the influence of pH on exchangeable NH + 4 desorption in near-surface sediments.Impacts of high pH condition on sediment-water nutrient exchange were de-termined via flux rate measurements of dissolved nutrients (SRP, NH + 4 , NO − 3 ), respiration rates (O 2 ), and denitrification rates (N 2 ) using intact sediment cores.We also calculated the diffusive flux rates of pore water NH + 4 , NH 3 and SRP to confirm direct flux measurements.
Since nitrification products may be released as NO − 3 and denitrified, we independently measured potential nitrification rates using slurries (Henriksen et al., 1981) and nitrification rates using an inhibitor (Caffrey and Miller, 1995).The potential nitrification method makes testing pH effects over a large pH gradient relatively simple, although it homogenizes the vertical gradients in sediments (e.g.redox potential and NH + 4 ) and disrupts the aggregation of aerobic/anaerobic microbiota (Killham, 1994;Garcia-Ruiz et al., 1998).A CH 3 Finhibition method can make up the shortcomings of slurry experiments.The shortcoming of nitrifying inhibitor likely leads to increased accumulation of pore water ammonium and non-specific inhibition of other N transformations such as ammonification (Capone et al., 2009).Therefore, both methods were taken to compare the nitrification responses to pH elevation.
We experimentally addressed these questions using sediment cores incubated with a continuous water flow-through system.In the Chesapeake Bay, cyanobacterial blooms occur frequently in the shallow and tidal freshwater tributaries, such as the Sassafras River and Potomac River (Tango and Butler, 2008).Relative to sea water, tidal fresh and oligohaline water have low pH buffering (Price et al., 2008), facilitating high pH levels from cyanobacterial photosynthetic carbon uptake.During dense cyanobacterial blooms at the Sassafras River, high pH persisted in the range of 9 to 10.5 for several weeks (eyesonthebay.net).When such high pH is in contact with bottom sediment, pH penetration into sediment can impact nutrient biogeochemical processes (Bailey et al., 2006).

Study site and collection of cores
In the upper Sassafras River, we collected sediments at the Powerline site (75°49.712, 39°22.646 ) on 18 June 2008 and Budds Landing (75°50.380, 39°22.310 ) on 14 July 2009 using 7 cm-inner diameter, 30 cm-long acrylic cores (Table 1).Water depths were 0.8 m at Powerline and 1.3 m at Budds Landing.Dissolved oxygen (DO), salinity, pH and temperature were measured with a YSI 600XLM multiparameter sensor.Vertical profiles of irradiance were recorded using a 2π Li-Cor underwater PAR light sensor.Bottom water was pumped through an inline filter (nominally 0.8 µM) to minimize autotrophic and microbial respiration, and nutrient recycling as well.We transported samples to Horn Point Laboratory within 4 h.Sediment cores were gently aerated overnight with aquarium pumps in order to reach oxygen saturation before experimentation and to equilibrate temperature, O 2 , and N 2 -N in the overlying water and near surface pore waters (Kana et al., 2006).Although photosynthetic carbon removal may result in high in situ pH (Table 1), the experimental aeration increased the dissolved CO 2 concentration and pH neutralization in the water column (Table 2).Surface sediments (top 2 cm) were homogenized for potential nitrification measurements, while the remaining sediment core was used for grain size measurement.

Experimental design
We incubated experimental cores (at least triplicates) at several pH levels to investigate pH effects on nutrient exchange at the sediment-water interface.Sediment-free blank cores were incubated identically at each pH level to correct for water column metabolic activity and nutrient cycling.Consistent with an absence of photosynthetically active radiation in bottom water at the time of collection, a dark temperature-controlled environmental chamber was used to maintain the sediments and replacement-water reservoir at in situ water temperatures of 25 °C for Powerline and 27 °C for Budds Landing, respectively.Sediment cores were sealed with acrylic o-ring tops that suspended a magnetic stir-bar beneath it, and a rotating magnetic turntable was set in the center that drove all magnetic bars to mix the overlying water at same speed.A flow-through system was set up for each core, which faciliated maintaining constant conditions (e.g.pH, nutrient levels, oxygen, flow speed) inside of sediment cores.In each water reservoir, the filtered water was bubbled with air to maintain saturated oxygen and adjusted to the desired pH with 0.1 mol l −1 NaOH.We continuously pumped water from reservoirs into the overlying water (∼500 ml) of the cores at 10 ml min −1 using a Rainin Rabbit peristaltic pump.
For the Powerline experiments, the overlying water pH of 4 replicate sediment cores was increased stepwise from 7.8 ± 0.5 to 9.2 ± 0.05 and 9.6 ± 0.03, with a five-day equilibration at each elevated pH.An alternative approach was used with sediments from Budds Landing.Nine cores were incubated at ambient pH for the initial fluxes, and then triplicate cores were subjected to pH manipulation for each treatment.After 7 days of exposure to higher pH levels, the overlying water pHs were 7.4 ± 0.3, 9.2 ± 0.05, and 9.5 ± 0.2 (Table 2).
For both sites, nutrient fluxes (SRP, NH + 4 and NO − 3 ), oxygen consumption (O 2 ) and coupled nitrificationdenitrification (net N 2 flux) were measured on the 1st day of incubation of sediments and after the pH reached the target values.After the termination of flux incubations, the sediments from each pH treatment were sectioned for pore water profiles of nutrients and determination of pH.Sediment cores from Budds Landing were used to evaluate the pH impacts on nitrification rates and ammonium desorption.Based on Br − penetration (Martin and Banta, 1992), we estimated the diffusion/advection coefficients of NH + 4 , NH + 3 and SRP.The remaining sediment cores were used to estimate the percent water content.Dissolved O 2 and N 2 subsamples were preserved in 7 ml glass tubes by adding 10 µl of 50 % saturated HgCl 2 solution (Kana et al., 2006), and stored under water at near-ambient temperature until analysis.To preserve total dissolved ammonium ( NH x = NH + 4 + NH 3 ) at higher pH levels, 2.5 µl of 0.1 mol l −1 sulfuric acid was added into the sample vials.Flux rates were calculated from the regression coefficients of the time-concentration data in sediment.Blank incubations were used to correct sediment core fluxes for any changes in concentration caused by water column activity.

Sediment pore-water chemistry
Over the top ∼10 cm sediments, samples were sectioned for pH and pore water analysis in a nitrogen-filled glove bag to minimize oxidation artifacts (Bray et al., 1973).Vertical changes of pH were measured immediately with a flat surface pH electrode.Sediments were sectioned into 50 ml centrifuge tubes and centrifuged at 2000 G for 10 min.Supernatant solutions were filtered through a 0.45 µm 25 mm diameter cellulose acetate syringe filter and appropriately diluted for analysis of Br − , Fe, SRP, NH + 4 , and NH x .The total iron, mostly Fe 2+ , was acidified for preservation (Gibbs, 1979).

Nitrification potential and nitrification rates
The effect of pH on nitrification was estimated using sediments from Budds Landing.Measurements included slurry incubation for potential nitrification (Henriksen et al., 1981) and CH 3 F inhibition of nitrification in intact sediment cores (Caffrey and Miller, 1995).Potential nitrification was measured with the surface sediments (0-2 cm depth) from Budds Landing.In O 2 -saturated Sassafras River water, pH was preadjusted with NaOH to values from 7 to 11.For 3 centrifuge tubes at each pH level, we added NH 4 Cl to final concentration of 1.0 mM and then added 1ml homogenized sediment.We gently mixed the suspensions in darkness at 27 °C using a orbital shaker, and took subsamples for NO − 3 at 0, 12, and 24 h to calculate rates of potential nitrification.Changes of NO − 3 in a sediment-free control were used as a background correction.
Alternatively, nitrification rates were measured based on the assumption that addition of CH 3 F can cut off nitrification and enhance direct NH + 4 fluxes (Caffrey and Miller, 1995).The CH 3 F method was carried out immediately after the end of flux measurements.Briefly, saturated solutions of CH 3 F were injected into the overlying water of intact cores to a final concentration of ∼100 mg l −1 .After 24 h of aerobic dark pre-incubation, ammonium flux rates were measured using our standard flux procedure.Due to the inhibition of am-monium oxidation, increased flux rates of ammonium after CH 3 F treatment were interpreted as the nitrification rate.

Molecular diffusive flux rates
Diffusion coefficients in sediment were estimated from Br − penetration profiles (Martin and Banta, 1992).Bromide (NaBr) was added as a tracer into the overlying water to a final concentration of ∼6 mM.After 24 h, vertical profiles of pore water Br − were measured to calculate the diffusion coefficient (D Br ), which was corrected for temperature and sediment porosity.The measured D Br was compared to the theoretical coefficient (D * Br ) to aid in correction of diffusion coefficients for other species (Martin and Banta, 1992;Schulz et al., 2006;Rao and Jahnke, 2004).
Using the pH-dependent equilibrium (Eq.1), we calculated pore water NH 3 and NH + 4 concentrations: where the equilibrium constant (pK a ) is 9.25 at 25 °C; constants were corrected for ionic strength and temperature (Mulholland, 2008).
The dissolved NH 3 concentration, [NH 3 ], can be calculated (Van Nester and Duce, 1987): where [ NH x ] is the sum of dissolved NH 3 and NH + 4 .The diffusion coefficients (D i ) of NH 3 , NH + 4 and SRP were corrected using the Br − diffusion estimates in pore water and the theoretical coefficient (D * Br ).Applying Fick's first law, the NH 3 , NH + 4 and SRP fluxes were calculated by: where F i is the flux of different species (µmol cm −2 s −1 ).The diffusion coefficient (D i ) is influenced by tortuosity (θ ), temperature and sediment properties; ∂C i ∂x is the gradient of nutrient concentration (C i ) and depth (x) (Table 3).The diffusion coefficients of NH 3 , NH + 4 and SRP in sediments were corrected using the D Br estimates and the diffusion coefficients in a particle free solution at in situ temperature (Martin and Banta, 1992;Rao and Jahnke, 2004;Schulz et al., 2006).Percent water and the dry sediment density (ρ ∼2.5 g cm −3 ) were used to calculate porosity (∅) (Boudreau, 1997): ) (4)

-N)
In order to estimate pH effects on ammonium desorption from sediment, surface sediments were collected from Budds Landing in November 2008.Adsorbed NH + 4 was measured using KCl extraction (Morin and Morse, 1999).Triplicate 1 ml wet samples of the top 2 cm sediment were extracted twice with 39 ml of 2 mol l −1 KCl; samples were shaken for 2 h at the field temperature (10 °C).After centrifugation and filtration, the increase in NH + 4 concentration relative to the blank was used to quantify adsorbed ammonium.Adsorption coefficients (K) were used to describe this ion exchange behavior, following Rosenfeld (1979) and Mackin and Aller (1984): where ĈN is exchangeable NH + 4 on a dry mass basis (µmol g −1 ) and C N is the pore water ammonium concentration (µmol l −1 ).Porosity was measured for the top 2 cm of sediment (Table 3).
To simulate response of adsorbed ammonium to pH elevation, the homogenized sediment (0-2 cm) was suspended in pH adjusted water from the sampling site.We added 1 ml of wet sediment to 39 ml of pH-adjusted water.To inhibit dissimilatory NO − 3 reduction to NH + 4 , we used oxygensaturated water and left 5 ml headspace in the centrifuge tube.
NH x was measured after shaking, centrifugation and filtration.Assuming a NH 3 equilibrium between the aquatic and atmospheric phase, the total release of ammonium is estimated as the sum of total dissolved ammonium in the sample water ( NH x l ) and NH 3 gas within the head space (NH 3−g ): The headspace NH 3 was estimated from (1) the pH at the end of incubation, (2) ionic strength corrections for NH 3 (γ NH 3 ) and NH + 4 (γ NH +

Concentrations of NH +
4 , SRP and Fe were analyzed using colorimetric methods (Gibbs, 1979;Parsons et al., 1984).Concentrations of NO − 3 , NO − 2 and pore water Br − were determined using ion chromatography (Kopp and McKee, 1983).Dissolved N 2 and O 2 were measured by the ratios of N 2 :Ar and O 2 :Ar using membrane inlet mass spectrometry (Kana and Weiss, 2004;Kana et al., 1994).Percent water was determined as the weight loss of wet sediment after drying at 65 °C.After pre-treatment with sodium hypochlorite overnight to remove carbonates and organic matter, grain size was analyzed by wet sieving and followed by pipet analysis of the remaining silt and clay (Folk, 1974).

Physical conditions
The Powerline and Budds Landing sites have similar physical conditions, including shallow and aerobic water columns, low salinity (<0.2), and fine grain-sized sediments (Table 1).Light attenuation coefficients were 4.8 m −1 at Powerline and 4.2 m −1 at Budds Landing, resulting in dim to dark conditions at the sediment surface.Both sites often have experienced cyanobacterial blooms associated with high pH in summer (Maryland Department of Natural Resources, eyesonthebay.net).At the time of collection, in situ pH was 9.4 and 7.3 in bottom water at Powerline and Budds Landing, respectively (Table 1).After the air-water DIC equilibrium overnight, the initial pH levels were neutral and similar in our experimental control groups (Table 2).

Penetration of pH and pore water iron
Vertical profiles of pH and Fe rapidly responded to the diffusion of overlying water pH (= 9.6) into the pore water (Fig. 1a and b).Sediment pH under ambient condition indicated a weak acid condition, being nearly constant with depth; the elevated pH water column treatments resulted in pH > 9.0 in the top 1-2 cm of sediment, decreasing downward until values were similar to the control.Although pH may be buffered by cation exchange (e.g.Ca 2+ , Mg 2+ ), sulfate reduction, and anaerobic generation of acid (Cai et al., 2010), such high pore water pH levels (pH > 9.5) have been observed during algal blooms in tidal-freshwater estuaries (Magalhaes et al., 2002).Our elevated pH profiles in sediments were similar to a time-series study of pH penetration by Bailey et al. (2006) in the Potomac River.Sediment incubations at high pH (∼10) showed a downward movement of high pH over time and achieved pH > 9 at 4 to 8 cm depth in a week incubation (Bailey et al., 2006).
In our aerobic incubations, pore water Fe 2+ was undetectable at the surface and peaked in the upper anoxic sediment horizon.Increased pH lead to a reduction in Fe 2+ through hydroxide precipitation (Hutchins et al., 2007).As pH increased to 9.6 in the overlying water, the peak concentration of Fe 2+ simultaneously decreased from 118 µmol l −1 to 64 µmol l −1 , and its peak position shifted from 1.75 cm downward to 2.5 cm.

Effect of pH on the pore water SRP profile
Elevation of pH below the sediment-water interface to P release into pore water, with the peak SRP concentrations increasing from <40 µmol l −1 to 102 µmol l −1 (Fig. 1d and  f).This change was consistent with pH-related P releases from surface metal hydroxide complexes (Seitzinger, 1991;Boers, 1991).Under aerobic pH-neutral conditions, iron oxyhydroxides usually adsorb or co-precipitate P, hindering the flux of SRP across the sediment-water interface (Slomp et al., 1998).In contrast to neutral pH conditions, highly alkaline pH levels enhanced P mobility by breaking surficial Fe-P bonds, which increased pore water SRP gradients and the diffusion rate (Figs.neutral pH to 39 µmol m −2 h −1 under alkaline pH treatments (Fig. 1 and Tables 3, 4).

Effect of pH on SRP flux
Flux rates of SRP significantly increased as pH increased at both stations (Fig. 2).SRP efflux rates increased from <5 µmol m −2 h −1 in the control, to 15-25 µmol m −2 h −1 at pH 9.2, and to 35-55 µmol m −2 h −1 at pH 9.6 (Fig. 2).SRP release at the Powerline site were consistent with its molecular diffusive rates estimated from pore water profiles (Table 4).In the oligohaline region of the Potomac River, SRP release from sediment increased from <10 µmol m −2 h −1 at neutral controls, to ∼40 µmol m −2 h −1 at pH 9.5, and to ∼110 µmol m −2 h −1 at pH 10 (Seitzinger, 1991).Similar large increases in SRP flux rates have been observed at pH levels of 9.5 in freshwater sediments (Boers, 1991).

4
Calculated from NH x and pH changes in water, the peak of dissolved NH + 4 was 132 µmol l −1 at pH 8.9 while NH 3 gradually increased with pH from the upper 8s (Fig. 3a).Desorbed NH + 4 increased from 646 to 2647 nmol g −1 as pH rose from 6.5 to 12 (Fig. 3b).Increased pH stimulated the release of the absorbed ammonium into pore water via the conversion of NH + 4 to NH 3 .Although mineral surface charges become more and more negative as pH increased, un-ionized NH 3 did not substantially adsorb to the solid phase.At pH < 8.9, both increases in NH + 4 and NH 3 (Fig. 3a) likely resulted from the desorption of exchangeable NH + 4 from mineral surface.When pH approached pK a (i.e.pH = 9.25), NH + 4 conversion rate (− NH + 4 / pH) peaked with 53 % of ammonium ( NH x ) converted into NH 3 (Eq.2).Moreover, a positive relationship existed between the absorbed NH + 4 on mineral surfaces and NH + 4 concentrations in the pore water (Eq.5).Elevated pH (∼9 to 12) lead to a sharp decrease in NH + 4 in the pore water (Fig. 3a) and more than 90 % of ammonium was transformed to NH 3 .The loss of NH + 4 , along with un-ionized NH 3 formation, may further mobilize absorbed ammonium until approximately 80 % of exchangeable ammonium was desorbed (Fig. 3b).

Effect of pH on the pore water ammonium profile
Under normal pH conditions, NH + 4 linearly increased downcore to 720 µmol l −1 with negligible NH 3 present (Fig. 1c).The diffusive flux rate, primarily as NH + 4 , was 149 µmol m −2 h −1 (Table 4).In contrast, the NH x concentration in the pH 9.6 treatment increased to 975 µmol l −1 at ∼3 cm depth (Fig. 1e), reflecting pH-driven ammonium desorption from solid phase to pore water.Relative to NH x profile, conversion of NH + 4 to NH 3 in surface horizons resulted in a steeper concentration gradient of NH + 4 , increasing NH + 4 diffusive fluxes (Table 4).Similar to observations of salinity-enhanced ammonium desorption (Gardner et al., 1991), reduction in NH + 4 concentrations at surface may further promote ammonium desorption (Eq.5).Dissolved NH 3 exhibited a very sharp peak at 0-3 cm depth, yielding a rapid upward flux.Diffusive flux rates were the sum of 243 µmol m −2 h −1 for NH + 4 and 234 µmol m −2 h −1 for NH 3 (Table 4).Without consideration of NH x speciation in high pH cores, the diffusive rate calculated from the concentration gradient and diffusion coefficient of NH + Table 4. Efflux rates of SRP and NH x in control and in high pH treatments in sediment cores from Powerline.Net flux rates (±SE) were compared to rates estimated from molecular diffusion of pore water.
Treatment The overlying water pH SRP flux rates (µmol-P m 2 h −1 ) Ammonium flux rates (µmol-N m 2 h was only 271 µmol m −2 h −1 , less than half of the observed NH x diffusive rate (Fig. 1e).

Effect of pH on DIN flux
For both Powerline and Budds Landing experimental cores (Fig. 4a), flux rates of NH x increased significantly in the high pH treatments relative to the control (p < 0.05, two-way ANOVA), but differences between pH 9.2 and 9.5 were not always significant.Compared to the control group, high pH (9.5-9.6)promoted NH x flux rates by about 6-fold at Powerline and by 2-fold at Budds Landing.These increases were consistent with the pH-induced ammonium desorption at surface sediments and the observed changes in the pore water profile.The conversion of NH + 4 to NH 3 and the steeper concentration gradients of these two components all resulted in elevated NH x fluxes.At the Powerline site, the ammonium release in the control was similar to the upward diffusion rate of ammonium, primarily as NH + 4 .The measured efflux rates of NH x at pH 9.6 were equivalent to the sum of the diffusive flux rate of NH + 4 and NH 3 (Table 4).Lack of consideration of NH 3 production would result in underestimation of ammonium flux rates by 25-35 % for the flux measurement and by 50 % for the diffusive flux estimation.
Ammonium remineralization, calculated either by the stoichiometric oxygen-based N remineralization or measured total inorganic nitrogen flux (i.e.NH + 4 + NH 3 + N 2 -N + NO − 3 ), suggests elevated pH dramatically promoted N efflux.If we assume that aerobic N remineralization stoichiometry from phytoplanktonic organic matter is 138O 2 :16N and denitrification is partly fuelled by the diffusion of water column NO − 3 into sediment (Cornwell et al., 1999), NH x flux accounted for 20-40 % of oxygen-based N remineralization in the control and 68-153 % of remineralized N in the high pH treatment.Alternatively, if nitrogen remineralization rates were evaluated from the sum of DIN flux, pH elevation increased ammonium flux as a proportion of total N remineralization from 22 % to 105 % for sediment at Powerline and 44 % to 87 % at Budds Landing.Both estimates reveal that high pH enhanced the proportion of ammonium release relative to the total remineralized N.However, the difference of NH + 4 remineralization between the two estimates may result from the use of O 2 consumption rates instead of CO 2 + and NH 3 concentrations were estimated fro 5 concentrations and pH in the aquatic phase (Eq.2).Desorbed NH 4 + was the sum 6 water and the volatilized NH 3 in the headspace of the sealed centrifuge tubes (Eq.7 Table 3).The dashed line represents 'total' absorbed NH 4 + , estimated by KCl extr 8 neutral sediment (Eq.5).The vertical and horizontal error bars are the stand 9 ammonium and the pHs, respectively.10 11 12 Fig. 3. Experimental pH effects on NH + 4 concentration in solution (A) and desorption of exchangeable NH + 4 (B), using the 0-2 cm homogenized sediments from Budds Landing collected in November 2008.Dissolved NH + 4 and NH 3 concentrations were estimated from the NH x concentrations and pH in the aquatic phase (Eq.2).Desorbed NH + 4 was the sum of NH x in water and the volatilized NH 3 in the headspace of the sealed centrifuge tubes (Eqs.( 6), ( 7), and Table 3).The dashed line represents "total" absorbed NH + 4 , estimated by KCl extraction of pH-neutral sediment (Eq.5).The vertical and horizontal error bars are the standard errors of ammonium and the pHs, respectively.
fluxes.The calculation of oxygen-based ammonium remineralization is affected by the production/reoxidation of reduced inorganic compounds (e.g.Fe 2+ , S 2− and Mn 2+ ), potential methanogenesis in organic-matter rich estuaries (Martens and Klump, 1980;Carini and Joye, 2008), and variable C:N ratios of organic matter.
No significant difference was found for NO − 3 flux rates among pH treatments (p > 0.05, ANOVA).Fluxes of NO − 34 perimental pH effects on total ammonium flux rates (A) and nitrate flux rates (B).res were taken from Powerline (PL) and Budds Landing (BL).Data are presented as rates ± standard error.At each site, different letters are used to show significant ukey test, P ≤ 0.05).directed into the sediment (Fig. 4b).In the oxygen saturated conditions, NO − 2 concentrations were generally too low to calculate flux rates via concentration changes over time.
Elevated pH may have two consequences for bacterially mediated nitrification.Nitrification is considered as first order or zero-order kinetics with respect to substrates NH + 4 /NH 3 availability (Park et al., 2010).However, increases in pH can enhance NH x desorption and the total inventories of exchangeable and pore water ammonium may be equal to or less than controls because of NH 3 volatilization.Moreover, high pH combined with abrupt changes in NH 3 (from zero to >550 µmol l −1 ) may result in the physiological inhibition of nitrification.In laboratory observations and modeling, both high pH and NH 3 have negative effects on nitrifying bacteria, ammonium-oxidizing bacteria (AOB, Nitrosomonas) and nitrite-oxidizing bacteria (NOB, Nitrobac- ter) (Van Hulle et al., 2007).Elevation of pH above 9 could inhibit enzyme activity of AOB and NOB since the optimal pH range is 6-8.5 for AOB and 5.5-8 for NOB (Van Hulle et al., 2007;Park et al., 2010).Even though nitrifying bacteria might survive out of the optimal pH range, they would pay an energy cost to maintain their cytoplasmic pH (Wood, 1988).
The accumulation of NH 3 can be toxic or inhibit the growth and enzyme efficiency of nitrifying bacteria (Anthonisen et al., 1976;Kim et al., 2006).
Although few field studies have been conducted on the nitrification response to high pH in sediments relative to water column and soil environments (Simek et al., 2002;Carini and Joye, 2008), sediment potential nitrification rates appear to be constrained by high pH (>8) in freshwater and were positively related to exchangeable NH + 4 in 36 stream surveys (Strauss et al., 2002).The inhibition of nitrification with elevated pH, with decreases of 80 % at pH > 9 relative to peak nitrification, has been observed in fine-grained sediment in the Arika Sea (Isnansetyo et al., 2011).

Effect of pH on nitrification rates
Elevated pH negatively impacted intact-core nitrification as measured by the changes in NH + 4 or NH + 4 flux rates after adding CH 3 F, an inhibitor of ammonium oxidation (Fig. 5b).Under neutral conditions, no significant difference existed between the evaluation of nitrification rates from NH + 4 flux (182 ± 49 µmol m −2 h −1 ) and from NH + 4 flux (210 ± 35 µmol m −2 h −1 ).Sediments in the upper Sassafras River show considerably higher nitrification rates than the <40 µmol m −2 h −1 typical observations from the mesohaline region of the Chesapeake Bay in summer (Kemp et al., 1990), reflecting the aerobic overlying water conditions.
Similar to nitrification potentials (Fig. 5a), increasing pH from neutral to 9.5 exerted a remarkable depression of nitrification rates, as evidenced by the >50 % reduction in nitrification under alkaline pH levels (Fig. 5b).If both dissolved and adsorbed NH + 4 /NH 3 in sediments are assumed to be the main substrates for nitrification (Seitzinger et al., 1991), high pH increases the diffusion of NH x through the oxic layer which may be lost before oxidation, lead to decreases in N availability, and functionally suppress nitrification.High pH penetration into the aerobic sediment surface (typically 1-2 mm), along with toxic NH 3 product, could suppress nitrification (Isnansetyo et al., 2011).In addition, nitrifying bacteria are obligate chemoautotrophs and grow with inorganic carbon in the form of CO 2 as their sole carbon source (Stanier et al., 1970); a reduction in CO 2 with pH elevation may therefore potentially inhibit nitrifying metabolism.

Effect of pH on denitrification
In aerobic Chesapeake Bay sediments, coupled nitrificationdenitrification is the key pathway to transform the rematerialized nitrogen to N 2 -N (Cornwell et al., 1999), while alternative N 2 production via annamox is inconsequential (Rich et al., 2008).Coupled nitrification-denitrification decreased from 180-280 µmol-N m −2 h −1 to less than 85 µmol-N m −2 h −1 as the overlying water pH increased to 9.6 (Fig. 6).Denitrification efficiency, the percentage of inorganic nitrogen released from the sediment as N 2 -N (Heggie et al., 2008), decreased from 84 % to 35 % at Powerline and from 64 % to 17 % at Budds Landing.
As pH increases, reduction of denitrification may be a consequence of limited NO − 3 supply and alkaline pH inhibition of denitrifying bacterial activity.The NO − 3 supply for denitrification may come from ammonium oxidation and diffusion from the overlying water.In this study, the contribution of NO − 3 (<7.5 µmol l −1 ) from the overlying water may be low relative to denitrification, evidenced by sediment NO − 3 uptake, accounting for <16 % of denitrification, and by the undetectable NO − 2 and NO − 3 concentrations in pore water as well.As pH rises, denitrification is likely limited by the NO − 3 supply, which mostly comes from the pH-suppressed nitrification (Fig. 6).Moreover, the optimal pH range for denitrification is 7-8 in soil and anaerobic sediments (Simek et al., 2002); higher pH may directly inhibit the activity of denitrifying bacteria.Nitrate reducing bacteria, such as Thioalkalivibrio nitratireducen, can survive in alkaline sediment and cultivation media at pH 10.However, the nitrite reductase activity of T. nitratireducens was maximal when pH ranged from 6.7-7.5, and 80 % of the activity was inhibited at high pH 9-10 (Filimonenkov et al., 2010).
Although dissimilatory nitrate reduction to ammonium (DNRA) in freshwater sediments appears to be minor relative to denitrification (Scott et al., 2008), DNRA usually occurs when NO − 3 inputs exceed the availability of carbon substrate for denitrification (Tiedje et al., 1989).As a consequence of pH elevation, limited NO − 3 consumption through anaerobic denitrification may provide the potential for DNRA and thus enhance ammonium production.Nevertheless, DNRA may play a minor role in explaining the enhanced ammonium fluxes.We did not expect high DNRA to occur in sediment with undetectable free sulfide concentrations.

Effect of pH on oxygen consumption
Oxygen consumption rates in the controls were higher in July at Budds Landing than in June at Powerline, partly a result of increased efficiency of bacterial organic matter decomposition with rising temperatures.However, oxygen consumption decreased as pH increased at both sites (Fig. 6).This is likely related to the alkaline pH effects on bacteria production, respiration and other oxidation metabolism (Tank et al., 2009).Assuming pH has no effect on organic matter remineralization to ammonium at each sampling site, we postulate that inhibition of nitrification by increased pH resulted in the reduction of oxygen consumption.
The molar ratio of O 2 to NH + 4 is 2 for nitrification.We calculated the changes of sediment oxygen consumption and NH + 4 flux before and after pH treatment, respectively.The measured slopes of NH + 4 and − O 2 fluxes were consistent with nitrifying stoichiometry (Fig. 7), which suggests high pH increased sediment NH + 4 diffusion into overlying water rather than enhancing coupled nitrificationdenitrification.Deviation of the − O 2 : NH + 4 flux rates from the theoretical 2:1 ratio may result from variation in sediment cores, such as oxidation of Mn (II) and Fe (II), and sediment buffering effects on OH − penetration in depth and magnitude.

Conclusion and ecological implications
Although cyanobacterial blooms are increasing in frequency and magnitude over time, determining the cause of such blooms remains a challenge (Glibert et al., 2011).Cyanobacterial blooms can be locally persistent and extensive, which may cause a dramatic increase in water column pH in poorly buffered water in lakes and tidal freshwater estuaries.Nutrients, especially N, limit primary production during the extensive summer blooms in Chesapeake Bay (Kemp et al., 2005).In our study region at the Sassafras River, diazotrophic cyanobacteria are dominant bloom-forming species during N-limited summer (O'Neil and Maryland DNR, unpublished data).
Enhanced nutrient release from bottom sediments can potentially satisfy the nutrient demand by algal growth, thus enhancing eutrophic conditions.Our study suggests pH elevation can increase inorganic N supply from sediment, making it available for organismal uptake.As pH increased above 9, the DIN efflux was more than doubled by promoting NH + 4 and NH 3 fluxes and inhibiting N 2 loss.Even though N 2fixing cyanobacteria can survive in N deficiency, they prefer to take up dissolved inorganic N rather than consuming energy for N 2 fixation (Paerl, 2008).The pH-induced release of ammonium from sediments may thereby be an important N source for primary productivity during the blooms.
With cyanobacterial-induced pH elevation, the different modes of N and P desorption result in discrepancies in the ratio of N:P supply.The release of P may be constrained by iron oxide adsorption at the oxic surface, and can be dramatically rates in the same core after pH was elevated from 7.8 to 9.5.Data from Budds Lan 5 changes of flux rates between control cores and cores at the pH of 9.2 and 9.5 a 6 incubation.The slope of the solid line is 2: 1, which is equal to the molar ratio of am 7 oxygen for nitrification (Eq.8).8 9 Fig. 7.The relationship between the increased NH x fluxes and the reduced oxygen consumption rates after pH elevation.Data from Powerline site are the changes of NH + 4 and O 2 flux rates in the same core after pH was elevated from 7.8 to 9.5.Data from Budds Landing site are the changes of flux rates between control cores and cores at the pH of 9.2 and 9.5 after 7 days incubation.The slope of the solid line is 2:1, which is equal to the molar ratio of ammonium to oxygen for nitrification (Eq.8).
increased above the threshold of pH 9-9.2 (Boers, 1991).In contrast, the corresponding desorption and release of NH + 4 from sediment may increase more gradually in response to pH increase.The interconversion of NH + 4 -NH 3 appears to be a key to changing N efflux rates.In this study, our experimental pH levels were at or above pK a , with only modest differences in ammonium efflux between middle pH (∼9.2) and high pH (9.5-9.6).We hypothesize that changes in sediment N cycling at pH in the upper 8s lead to NH 3 rapidly become an important N species.The molar ratios of DIN:SRP sediment efflux decreased from >70 to 9-12 when experimental pH rose from neutral to above 9.At pH levels in the high 8s, enhanced ammonium effluxes might result in elevated N:P ratios; further investigation of pH-related changes in sediment N cycling is warranted over a broader pH range.
At our field-validated experimental pH levels, which are consistent with photosynthetic-driven pH elevation by cyanobacterial blooms, the switch of sedimentary nutrient effluxes from high N to high P may reinforce N limitation and selectively support N 2 -fixing cyanobacterial blooms.Given higher P demand for diazotrophs, the augmentation of P flux with pH may boost the growth and persistence of algal blooms (Xie et al., 2003;Paerl, 2008).This, together with the increased DIN flux and diminished dentirification (N 2 loss), will lead to greater primary production and even faster element cycling in shallow waters.Cyanobacterial blooms appear to create a troublesome positive feedback that fosters

Flux
rates were measured on the first day of the incubation and after each equilibration period.The pumping of wwwet al.: Effects of cyanobacterial-driven pH increases treatment water was interrupted during flux measurements and briefly restarted to collect samples every 1.5 h, with a total of 4 time-points.Solute samples were filtered through a 0.45 µm cellulose acetate syringe filter and frozen at −4 °C.

K
a γ NH + Y. Gao et al.: Effects of cyanobacterial-driven pH increases 2.8 Chemical analysis

Figure 1 .Fig. 1 .
Figure 1.Powerline porewater profiles in the upper 10 cm of sediment under high pH (9.6) and 2 normal pH (7.4) treatments, including vertical changes of pH (A), porewater Fe (B), SRP (D,F), 3 and ∑NHx (C,E).The dashed line is the interface of sediment-water.Changes in ammonium 4 speciation, resulted from surface pH elevation, were calculated by the equilibrium of NH3 and 5 NH4 + .6 Fig. 1.Powerline pore water profiles in the upper 10 cm of sediment under high pH (9.6) and normal pH (7.4) treatments, including vertical changes of pH (A), pore water Fe (B), SRP (D, F), and NH x (C, E).The dashed line is the interface of sediment-water.Changes in ammonium speciation, resulted from surface pH elevation, were calculated by the equilibrium of NH 3 and NH + 4 .
1 and 2).As expected, elevated pH increased upward SRP diffusion from 5 µmol m −2 h −1 under 32 xperimental pH effects on SRP flux rates from sediments at Powerline (PL) and ing (BL).Error bars are the standard errors.Two-way ANOVA was used to test the n SRP release at both sites.With elevation in the experimental pH, SRP fluxes were different at each site (P < 0.01), but non-significantly different between stations (P erent letters are used to show significant difference.

Fig. 2 .
Fig. 2. Experimental pH effects on SRP flux rates from sediments at Powerline (PL) and Budds Landing (BL).Error bars are the standard errors.Two-way ANOVA was used to test the pH effects on SRP release at both sites.With elevation in the experimental pH, SRP fluxes were significantly different at each site (P < 0.01), but nonsignificantly different between stations (P > 0.05).Different letters are used to show significant difference.

Figure 3 .
Figure 3. Experimental pH effects on NH 4 + concentration in solution (A) and d 3 exchangeable NH 4 + (B), using the 0-2 cm homogenized sediments from Budds Land 4 in November 2008.Dissolved NH 4 + and NH 3 concentrations were estimated fro 5

Fig. 4 .
Fig. 4. Experimental pH effects on total ammonium flux rates (A) and nitrate flux rates (B).Sediment cores were taken from Powerline (PL) and Budds Landing (BL) sites.Data are presented as mean flux rates ± standard error.At each site, different letters are used to show significant difference (Tukey's test, P ≤ 0.05).

Fig. 6 .
Fig. 6.Experimental pH effects on denitrification rates (A) and oxygen consumption rates (B) of sediments from the Powerline (PL) and Budds Landing (BL) sites."*" indicates where measurements were not taken.Bars show the mean of triplicate cores, error bars are the standard error of the mean.Different letters are used to show significant difference (Tukey's test, P ≤ 0.05) due to pH changes at PL and BL, respectively.

Figure 7 .
Figure 7.The relationship between the increased ∑NH x fluxes and the reduc 3 et al.: Effects of cyanobacterial-driven pH increases their persistence by enhancing nutrient availability from sediments.

Table 1 .
Sediment grain size, ambient dissolved nutrients in water column and the flux rates before pH modification.Samples were collected at the Powerline site on 18 June 2008 and Budds Landing site on 14 July 2009.Grain size measurements were made after flux experiments; other measurements for dissolved nutrient concentrations (average ± SE, n = 3) in the bottom water and flux rates (mean ± SE, n = 4-9) were carried out before pH treatments.Negative flux rates indicate uptake by the sediment.

Table 2 .
Experimental overlying water pHs for experimental incubation of cores from the Powerline and Budds Landing sites.The Powerline incubations had a sequential change of pH (n = 4).The Budds Landing incubation had all cores (n = 9) at neutral pH on the first day, and then set 3 replicate cores at 3 different pHs, respectively.pH data are the mean values ± SE.

Table 3 .
The kinetic parameters used in calculation of diffusion rates and in calculation of ammonium adsorption-desorption in sediments.