Introduction
It is widely recognized that accelerated eutrophication of freshwater aquatic
environments is caused primarily by anthropogenic increases to dissolved
phosphorus (P) concentrations in surface water (Smith and Schindler, 2009).
Rapid cultural eutrophication of oligo or mesotrophic lacustrine and
palustrine systems is often attributed to increased external P loadings
originating in agricultural run-off or waste water treatment plant (WWTP)
effluent. The resultant excessive algal growth negatively impacts aquatic
ecosystems and economic activity (Pretty et al., 2003) and increases
the risk of infectious diseases (Chun et al., 2013). Strategies to mitigate
eutrophication have aimed to reduce point source and diffuse external
phosphorus loadings by instituting agricultural best-management practices in
the surrounding watershed (McLaughlin and Pike, 2014; Sharpley et al., 1994),
limiting P inputs to domestic waste water (Corazza and Tironi, 2011) and
upgrading WWTPs (Mallin et al., 2005). However, internal loading of P, from
sediments to surface water, remains poorly quantified in many systems and is
often the largest source of error in hydrodynamic and ecological phosphorus
models (Kim et al., 2013). Early diagenesis and mineralogical removal of
labile autochthonous organic phosphorus (Po) from solution is a
complex process and is poorly understood in highly dynamic systems despite
exerting a strong influence on the magnitude and timing of internal P
loading. This is particularly true in shallow lakes and wetlands due to the
high ratio of sediment surface area to water column depth (Søndergaard et
al., 2003).
As policy and infrastructure improvements continue in order to mitigate
external P inputs to aquatic systems, the relative importance of internal P
loads from legacy P in sediments to overall P budgets in eutrophic systems
is likely to increase (Reddy et al., 2011).
It has been widely demonstrated through laboratory and field investigations,
particularly in seasonally anoxic lakes, that sustained anoxic conditions
induced by water column stratification typically result in greater P
mobility and correspondingly higher water column P concentrations (Krom and
Berner, 1981; Einsele, 1936; Hongve, 1997; Katsev et al., 2006; Mortimer,
1941, 1971; Penn et al., 2000). The microbially mediated reductive
dissolution of iron(III) oxyhydroxides or iron(III) phosphate during
sustained periods of anoxia at the sediment–water interface (SWI) has long
been considered the main mechanism responsible for P release under anoxic
conditions (Bonneville et al., 2004; Hyacinthe and Van Cappellen,
2004).
More recently, considerable microbial polyphosphate accumulation and release
in response to alternating oxic–anoxic conditions at the SWI in lacustrine
environments has also been shown to occur (Gächter et al., 1988;
Gächter and Meyer, 1993; Hupfer et al., 2007; Sannigrahi and Ingall,
2005). In some cases this accumulation of polyphosphate by the microbial
community may account for 10 % of total phosphorus (Hupfer et al., 2007).
However, redox conditions in shallow, heavily bioturbated sediments are more
spatially and temporally variable than in seasonally anoxic lakes
(Aller, 1994; Gorham and Boyce, 1989), resulting in short-term redox oscillations even with continuous oxia at the SWI.
Additionally, the coupled biogeochemical cycles of other redox sensitive
elements such as sulfur and carbon have been shown to play important and
complex roles in phosphorus mobility (Gächter and Müller, 2003; Joshi
et al., 2015; O'Connell et al., 2015). For example, high bottom-water sulfate
concentrations have been shown to increase aqueous P in sediments by
decreasing the permanent mineralogical removal of P within vivianite
[Fe3(PO4)2(s)] and by decreasing the abundance of iron(III)
oxyhydroxides near to the SWI (Caraco et al., 1989). This is due to the
scavenging of iron during the formation of iron sulfide minerals within
sediment during diagenesis (Gächter and Müller, 2003). Carbon cycling
also exerts considerable control over phosphorus mobility within sediment.
The stoichiometry of freshly deposited organic matter (OM) in eutrophic water
bodies approaches that of primary production, i.e.
∼ C106 : N16 : P (Berner, 1977). Appreciable P may
therefore be released to the aqueous phase when organic carbon is mineralized
during microbial respiration of oxygen, nitrate, ferric iron or sulfate. In
addition to driving N, Fe and S cycling, mineralization of organic carbon and
concomitant P release has, in some places, been shown to be the primary
mechanism controlling phosphorus mobility at the SWI (Joshi et al., 2015).
Core incubations and in situ flux chambers frequently examine the effects of
anoxia on P mobility from sediments but the effects of repetitive redox
oscillations are rarely investigated in a controlled setting (Frohne et al.,
2011; Matisoff et al., 2016; Nürnberg, 1988; Thompson et al., 2006).
Consequently, the cumulative and reversible effects of oxic–anoxic cycling
on P distribution, speciation and mobility within sediments are poorly
understood.
The aim of this study is to elucidate the microbial and geochemical
mechanisms of in-sediment phosphorus cycling and release associated with
commonly occurring short redox fluctuations (days) in surficial sediments in
shallow eutrophic environments. Particularly, we aimed to (1) quantify the
redistribution of P between aqueous and mineral sediment pools during
fluctuating redox conditions; (2) determine whether the activity of hydrolytic
phosphatase enzymes acting on Po were influenced by redox
conditions; (3) assess whether the proportions of orthophosphate, Po,
and polyphosphate varied systematically with redox conditions; and
(4) ascertain whether P mobilization/immobilization mechanisms were reversible or
cumulative.
(a) Photograph of the sampling location taken on the day of
sampling illustrating the abundance of green filamentous algae.
(b) Map of Cootes Paradise and West Pond showing the sampling
location, local hydrological network and the King Street Waste Water
Treatment plant in Dundas, Hamilton. Colour represents area covered by surface water at
different water levels (m.s.l. is metres above sea level).
(c) Overview map showing the hydrological connection between Cootes
Paradise, Hamilton Harbour and Lake Ontario.
We conducted controlled bioreactor experiments using sediment suspensions,
designed to reproduce cyclic redox conditions analogous to those occurring in
nature (Aller, 1994, 2004). A combination of aqueous chemistry, sediment
sequential extractions (Ruttenberg, 1992), 31P NMR spectroscopy
(Cade-Menun, 2005) and extra-cellular enzyme assays (Deng et al., 2013) were
used to produce a comprehensive dataset describing bulk chemistry, microbial
and mineralogical controls on P mobility and speciation during redox
oscillations.
Methods
Field site and sampling
Surface sediment (0–12 cm), sediment cores (34 cm long, 10 cm diameter),
overlying water and green filamentous algae (GFA) were collected on
5 September 2013 from West Pond in Cootes Paradise Marsh
(43.26979∘ N, 79.92899∘ W) following established guidelines
(US EPA, September 2009). Cootes Paradise is a hypereutrophic, coastal
freshwater marsh, which drains into Lake Ontario via Hamilton Harbour (see
Fig. 1a–c). The marsh system suffered severe degradation due to rapid
urbanization, population growth and nutrient loadings in the 20th century
(Chow-Fraser et al., 1998). West Pond in particular received extremely high
external P loads from the King Street (Dundas) WWTP for several decades prior
to the installation of sand filters in 1987 (Painter et al., 1991). The
addition of sand filters, and other improvements, decreased P loadings from
the WWTP from 45 kg P day-1 in the early 1970s (Semkin et al., 1976)
to 4.5 kg P day-1 in the 1980s (Chow-Fraser et al., 1998) and
2.59 kg P day-1 in 2011 (Routledge, 2012). However, high external P
loads resulted in accumulation of legacy P in West Pond sediments with total
phosphorus concentrations reaching 200 µmol g-1 by the 1980s
(Theysmeyer et al., 1999). Consequently, dredging was conducted in 1999 in an
attempt to remediate the areas most affected by growth of green filamentous
algae (Bowman and Theysmeyer, 2014).
Despite these restoration efforts and decreases to the external P load,
pervasive growth of GFA during the summer persists in parts of Cootes
Paradise (Fig. 1a). Cyanobacteria are not commonly observed at this location,
potentially due to the high N : P ratios often associated with WWTP which
utilize tertiary P removal treatment (Conley et al., 2009; Stumm and Morgan,
1996).
Sediment characterization
Sediment cores were sliced every 3 cm, homogenized and characterized with
bulk sediment samples prior to bioreactor experiments. Organic carbon and
carbonate depth profiles were determined by thermogravimetric analysis
(TGA-Q500, TA Instruments Q500) (Pallasser et al., 2013). Water content and
bulk density (ρb) of the sliced sediment core were determined
gravimetrically after oven drying (Gardner, 1986). Identification and
quantification of crystalline mineralogy was determined by powder X-ray
diffraction (XRD) (Empyrean Diffractometer and Highscore Plus software Ver.
3.0e PANalytical). The density of benthic macro-invertebrates was also
quantified after sieving two additional cores (7.5 cm diameter, 18 cm
deep)
through 500 µm mesh.
Bioreactor experiment and redox oscillation procedure
An initial concentrated sediment suspension of approximately
500 g L-1 (dry weight equivalent) was prepared from freshly sampled
sediment (0–12 cm) and filtered overlying water (< 0.45 µm).
Surface water was used, rather than distilled water, to provide background
ionic strength and avoid osmotic shock to the microbial community. The
concentrated suspension was stirred vigorously for 5 min then passed through
a < 500 µm stainless steel sieve to remove larger solid organic
material and macro-invertebrates. This procedure was repeated until a
homogeneous suspension was achieved. The dry weight was then re-determined,
and the sieved solution was diluted with filtered surface water to a final
concentration of 247 ± 2 g L-1. The resulting suspension was
transferred to a bioreactor system (Applikon Biotechnology) after Thompson et
al. (2006) and Parsons et al. (2013). In addition to affording precise
temperature control and continuous logging of temperature, redox potential
(Eh) and pH, the system offers significant advancements over previous
designs (Thompson 2006; Guo, 2007; Parsons, 2013). The Eh, pH and
dissolved oxygen (DO) were measured using a combined autoclavable Mettler
Toledo InPro 3253i/SG open-junction electrode and an AppliSens low drift
polarographic sensor. The InPro electrode system, using a common reference
electrode, was chosen to help avoid potential interference between two
electrodes in close proximity. A multiparameter transmitter was used to
display current pH, Eh and temperature to automatically temperature-correct pH values and to adjust measured Eh to the standard hydrogen
electrode (SHE). DO was calibrated using 100 % saturation in air
(approximately 0.2905 atm) and 0 % saturation in N2 at constant
sparging of 30 mL min-1
The suspension was stirred continuously and sparged with 30 mL min-1
air for 11 days to equilibrate prior to the redox oscillation procedure.
During the 11-day oxic equilibration period, CO2 emissions were
monitored in the reactor exhaust gas using an IR sensor (Applikon
Biotechnology).
Subsequently, redox potential (Eh) variation was induced by the
modulation of sparging gases (30 mL min-1) between
N2 : CO2 and O2 : N2 : CO2. The suspension was
subjected to five cycles of anoxia (7 days) and oxia (7 days) at constant
temperature (25 ∘C) in the dark, while recording Eh, pH, DO and
temperature data. The suspension was sampled on days 1, 3, 5 and 7 of each
half-cycle. To separate solid and aqueous components from the sediment
suspension, syringe-extracted samples (15 mL) were centrifuged at 5000 rpm
for 20 min and the supernatant filtered through 0.45 µm
polypropylene membrane filters prior to all aqueous analysis. For samples
taken during anoxic half-cycles, centrifugation, filtering and subsampling
were performed in an anoxic glove box (N2 : H2 97:3 %,
O2 < 1 ppmv). Time periods were chosen to be representative of
short temporal fluctuations to redox conditions experienced by surficial
sediments (Aller, 1994; Nikolausz et al., 2008; Parsons et al., 2013).
Similarly, the temperature and dark conditions were chosen to reflect those
measured in surficial sediment during summer months at the field site. Summer
conditions were chosen as this is when maximum bioturbation, microbial
activity and OM input are expected within the sediment. Similar, long-running
batch reactor experiments using soil or sediment have previously experienced
a slowdown of metabolic processes due to depletion of labile organic carbon
(Parsons et al., 2013). Therefore, gaseous carbon and nitrogen losses from
the reactor were balanced by the addition of 3 g of freeze-dried and ground
GFA to the suspension at the onset of each anoxic cycle. The amount of algae
added was determined based on CO2 production from the reactor during the
initial 11-day oxic period.
Aqueous-phase methods
All reagents used were of analytical grade from Fluka, Sigma-Aldrich or Merck
unless stated otherwise and were prepared with 18.2 MΩ cm-1
water (Millipore). Total dissolved Na (70), K(100), Ca(20), Mg(0.5), Mn(1),
Fe(3), Al(100), P (2), Si (15) and S (15) concentrations (MDL in
µg L-1, in parentheses) were determined by ICP-OES (Thermo
Scientific iCAP 6300) after filtration (< 0.45 µm) and
acidification with HNO3 to < pH 2. Matrix-matched standards were used
for all calibrations and NIST validated multi-elemental solutions were used
as controls. SRP concentrations were determined by the molybdenum
blue/ascorbic acid method on a LaChat QuickChem 8500 flow injection analyzer
system (4500-P E: Phosphorus by Ascorbic Acid, National Water Quality
Monitoring Council, 1992; Murphy and Riley, 1962) (MDL
1.2 µg P L-1). DOC was determined using a Shimadzu
TOC-LCPH/CPN analyzer (Shimadzu) following HCl addition (< pH 2) to degas
dissolved inorganic carbon (MDL 71 µg C L-1).
Chloride, nitrate, nitrite and sulfate concentrations were measured by ion
chromatography using a Dionex ICS 5000 equipped with a capillary
IonPac® AS18 column. Aqueous sulfide was
stabilized with 20 µL 1 % zinc acetate per mL (Pomeroy, 1954)
after filtering and determined by the Cline method (Cline, 1969) (MDL
0.5 µM). Fe(aq)2+ was determined by the ferrozine
method immediately after filtering (Stookey, 1970; Viollier et al., 2000)
(MDL 3.8 µM). All aqueous analyses were conducted in triplicate.
The precision and accuracy for all techniques was < 5 RSD % and
±10 % with respect to certified reference materials (where
commercially available).
Summary of the modified SEDEX sequential extraction protocol used on
solid samples taken over a time series during the reactor experiment. Results
of this extraction are shown in Fig. 4.
Step
Extractant
Conditions
Target phase
1a
1 M MgCl2
pH 8 for 2 h at 25 ∘C
Exchangeable or loosely
1b
1 M MgCl2
pH 8 for 2 h at 25 ∘C
sorbed P (PEx)
1c
18.2 MΩ cm-1 H2O
2 h at 25 ∘C
2a
1 M NaHCO3
pH 7.6 for 16 h at 25 ∘C
Organic associated P
2b
1 M NaHCO3
pH 7.6 for 2 h at 25 ∘C
(PHum)
2c
1 M NaHCO3
pH 7.6 for 2 h at 25 ∘C
2d
1 M NaHCO3
pH 7.6 for 2 h at 25 ∘C
2e
1 M MgCl2
pH 8 for 2 h at 25 ∘C
3a
0.3 M Na3 citrate with 1 M NaHCO3
pH 7.6 for 8 h at 25 ∘C
Fe-bound P (PFe)
with 1.125 g of Na dithionite (CDB)
3b
CDB
pH 7.6 for 2 h at 25 ∘C
3c
1 M MgCl2
pH 8 for 2 h at 25 ∘C
4a
1 M Na acetate with acetic acid
pH 4 for 6 h at 25 ∘C
Authigenic carbonate fluorapatite
4b
1 M Na acetate with acetic acid
pH 4 for 2 h at 25 ∘C
plus biogenic apatite plus
4c
1 M MgCl2
pH 8 for 2 h at 25 ∘C
CaCO3-bound P (PCFA)
5
1 M HCl
16 h at 25 ∘C
Detrital apatite plus other
inorganic P (PDetri)
6
1M HCl
16 h after ashing at 550 ∘C at 25 ∘C
Residual/organic P (PResi)
Solid-phase methods: phosphorus and iron speciation
Phosphorus partitioning within the solid phase in the reactor experiment was
evaluated over a time series by both sequential extractions, using a
modification (Baldwin, 1996) of the SEDEX extraction scheme
(Ruttenberg, 1992) and solution 31P NMR spectroscopy
(Cade-Menun, 2005). The two approaches are complementary;
31P NMR spectroscopy provides information on the molecular speciation
of phosphorus, while sequential extraction provides information on the
association of the P species with operationally defined solid-phase
fractions. Therefore the combination of these two methods reveals
redistribution of P within the solid phase over time during oxic–anoxic
transitions.
The original SEDEX extraction scheme quantifies five different P reservoirs
within sediment by consecutively solubilizing progressively more recalcitrant
phases by using extracts of increasing severity. The reaction mechanisms
associated with each extraction step are discussed in detail within
Ruttenberg (1992). A modification of the SEDEX extraction scheme proposed by
Baldwin (1996), used here, incorporates an additional 16 h, 1 M NaHCO3
step (PHum) after the PEx step, to differentiate OM-associated P which may otherwise be co-extracted during the PFe
step. The pH of the NaHCO3 extraction step was adjusted to 7.6 to
minimize dissolution of carbonates prior to the PCFA extraction
step. A total of 15 samples between days 11 and 74 of the reactor experiment
were analysed in duplicate by sequential extraction. A summary of the full
sequential extraction method used, including target phases, reactants, pH,
temperature and reaction times is provided in Table 1.
Changes to iron speciation were also evaluated through a time series during
the experiment. To account for surface-sorbed or freshly precipitated Fe,
total Fe2+ production during anoxic half-cycles was estimated by a
partial extraction (1 h, 0.5 N HCl) on sampled suspensions.
Fe2+ / Fe3+ ratios were determined in extracts using a
modification of the ferrozine method (Stookey, 1970; Viollier et al., 2000).
Additionally, a thermodynamic model was implemented in PHREEQC (Parkhurst et
al., 1999) to assess the saturation index (SI) of various minerals over time
during the reactor experiment using measured pH, temperature, Eh and
concentration data.
(a) Proportions of bioturbating macro-invertebrates identified in the top 18 cm.
(b) Depth profiles of sampled sediment, water content weight %
(inverted black triangles), bulk density (white diamonds) OM % (white
squares) and carbonate % (black circles). (c) Mineralogical
composition of sediments from the zone of bioturbation determined by XRD (top
12 cm).
NaOH–EDTA extraction and solution 31P NMR spectroscopy
Molecular changes to P speciation were evaluated over a time series by
solution 31P NMR. Phosphorus was extracted directly from suspension
samples (∼ 2 g dry weight equivalent) prior to 31P NMR analysis.
The method used has been shown to allow quantitative analyses of Po
(monoesters and diesters), polyphosphates and orthophosphates (Amirbahman et
al., 2013; Cade-Menun et al., 2006; Cade-Menun and Preston, 1996; Reitzel et
al., 2007; Turner et al., 2003b). Briefly, samples were extracted in 25 mL
of 0.25 M (NaOH) and Na2EDTA (0.05 M) at ambient laboratory
temperature (∼ 22 ∘C) for 4 h. Subsequently, the tubes were
centrifuged (2300g for 20 min), and the supernatant extracted via
syringe and then neutralized with 2 M HCl to a pH of 7 to avoid the
breakdown of polyphosphates during freeze-drying (Cade-Menun et al., 2006).
This solution was then filtered to < 0.45 µm. Prior to
freeze-drying 1 mL aliquots of each sample were diluted and analysed by
ICP-OES spectroscopy for Al, Ca, Fe, Mg, Mn and P. The remaining extracts
were frozen at -80 ∘C and freeze-dried for 48 h. The freeze-dried extracts were re-dissolved in 1.0 mL D2O, 0.6 mL 10 M NaOH
and 0.6 mL of the NaOH–EDTA extractant solution and were allowed to stand
for 10 min with occasional vortexing. Samples were centrifuged for 20 min
at 2300g, transferred to 10 mm NMR tubes and stored at 4 ∘C
before analysis within 12 h.
Solution 31P NMR spectra were obtained using a 600 MHz spectrometer
equipped with a 10 mm broadband probe. The NMR parameters were 90∘
pulse, 0.68 s acquisition time, 4.32 s pulse delay, 12 Hz spinning,
20 ∘C and 2200 to 2900 scans (3–4 h) (Cade-Menun et al., 2010).
Phosphorus compounds were identified by their chemical shifts
relative to an external
orthophosphoric acid standard, with the orthophosphate peak in all spectra
standardized to 6 ppm. Peak areas were calculated by integration on spectra
processed with 10 and 7 Hz line broadening, using NUTS software (Acorn NMR,
Livermore CA, 2000 edition). Peak assignments were grouped into compounds or
groups of specific compound classes when direct identifications could not be
made (Cade-Menun, 2005).
Extracellular enzyme assays
Rates of enzymatic hydrolysis of Po were estimated through
extracellular enzyme activities. Degradation rates for
phosphomonoesters, phosphodiesters and pyrophosphate were determined
fluorometrically through use of the MUF-tagged substrates: MUF phosphate
(MUP), bis(MUF)phosphate (DiMUP, Chem-Impex International) and MUF
pyrophosphate (PYRO-P, Chem-Impex International) respectively. Additionally
MUF β-D-glucopyranoside (MUGb) was used in order to compare
phosphatase enzyme activity to the activity of β-glucosidase
(cellulase) (Dunn et al., 2013). Enzyme activities were determined using a
microplate reader (Flexstation3, Molecular Devices) using a modification of
Deng et al. (2013). Briefly, 1 g dry weight equivalent of suspension from
the reactor was stirred with 100 mL of 100 mM HEPES buffer at pH 7.5 in a
pyrex dish for 10 min at 280 rpm to allow for complete homogenization.
Subsamples (100 µL) of the buffered soil suspension were removed
during continuous mixing using a multi-channel pipette and placed into
microplate wells, which were loaded into the microplate reader. Four
replicate wells were filled per substrate. Plates were left to equilibrate at
30 ∘C for 5 min inside the reader before the automatic addition of
100 µL of substrate, resulting in a final substrate concentration
of 667 µM. Each well was triturated thoroughly during addition of
the substrate. Excitation fluorescence was set at 365 nm. Emission intensity
at 450 nm was recorded at 5 min intervals over a 6 h period. The effect of
fluorescence quenching was accounted for in each sample by preparing MUF
calibration curves in the same soil suspension as used for the analysis. The
limits of detection and quantification were determined to be 1.1
and 3.3 µM MUF respectively, equivalent to 1.1 µM of
phosphate for the determination of phosphomonoesterase activities.
Results and discussion
Sediment characterization and evidence of bioturbation
Characterization of sediment cores revealed physical and chemical solid-phase
homogeneity within the top 10 cm, with a bulk density of
∼ 1.3 g cm-3, water content of ∼ 50 % (by weight), OM
of ∼ 3 % and a carbonate of ∼ 25 % (Fig. 2b). Between 10
and 15 cm depth increases in bulk density and decreases to the sediment
water content, OM % and carbonate fraction occurred as soft sediment
transitioned to clay.
A benthic macro-invertebrate density of approximately 49 500 individuals
per m2 was determined, consistent with previously reported values
(Pelegri and Blackburn, 1995). The community (Fig. 2a) was dominated by
aquatic earthworms (Tubificidae 60 % and Branchiura sowerbyi 8 %), which typically feed and mix sediment within the top
5–10 cm (Fisher et al., 1980; McCall and Fisher, 1980). Other groups
identified included Ceratopogonidae (no-see-ums or biting midges,
22 %) including Sphaeromias, Probezzia, Bezzia and Chironomidae
(midges, 6 %) including Cryptochironomus, Tanypus and
Nemotoda (round worms, 4 %), and a single Hyalella azteca (scud < 1 %).
Bioturbating organisms, such as those identified, have previously been shown
to alter biogeochemical cycling within surface sediments (Hölker et al.,
2015). Reported influences include increased solute fluxes (Furukawa et al.,
2001; Matisoff and Wang, 1998), mixing of solid sediment (Fisher et al.,
1980) and bioconveying of sediment particles (Lagauzère et al., 2009).
These processes have been shown to enhance sediment oxygen demand (McCall and
Fisher, 1980; Pelegri and Blackburn, 1995), degradation of OM (Aller, 1994),
rates of denitrification, transport of contaminants to surface water
(Lagauzère et al., 2009) and temporal fluctuation of redox conditions
(Aller, 1994). Efficient sediment mixing allows frequent re-oxidation of
reduced sediments and therefore regeneration of terminal electron acceptors
(TEAs) such as nitrate, ferric iron and sulfate, which often limit
mineralization of OM in sediments underlying hypereutrophic water bodies
(Reddy and DeLaune, 2008). Electron donors in the form of fresh autochthonous
necromass are also rapidly redistributed vertically within the zone of
bioturbation. This environment should therefore support a metabolically
diverse, abundant and highly active microbial community (DeAngelis et al.,
2010).
Quantitative XRD analysis of the top 12 cm of sediment (Fig. 2c) showed
close agreement with the carbonate fraction determined by TGA
(∼ 25 % by TGA vs. 27 % by XRD) indicating a calcite-dominated,
carbonate-buffered system. The remaining mineral assemblage was dominated by
quartz and clay minerals (illite 30 % and chamosite 2 %). No pyrite
or vivianite was detected by XRD, suggesting either their absence or presence
in low abundance (< 1 %) with poorly crystalline structures.
Experimental redox oscillation: aqueous chemistry
Eh within the bioreactor oscillated between +470 and -250 mV
(Fig. 3) consistent with Eh ranges of wetland sediments (Nikolausz et
al., 2008). A slight pH oscillation was also present between ∼ 7.4
during anoxic half-cycles (N2 sparging) and ∼ 7.7 during oxic half-cycles (N2 : O2 : CO2 sparging) this variation, shown in
Fig. 3, is consistent with calcite/dolomite-buffered sediment equilibrating
with changing pCO2 caused by sparging gas composition and
microbial respiration. After the 11-day equilibration, ionic strength of the
aqueous phase in the reactor suspension remained at
∼ 0.025 ± 0.004 M for the duration of the experiment. This range
of Eh / pH conditions and ionic strength is consistent with the
range measured within surficial sediment at the field site and transitions
across the thermodynamically predicted stability boundaries for multiple
redox couples, e.g. MnO2 / Mn2+,
NO3- / NO2- / NH4+,
Fe(OH)3 / Fe2+, SO42- / HS-, during each 14-day redox cycle. The upper Eh values recorded during oxic cycles are
significantly lower than predicted by the O2 / H2O couple
(820 mV at pH 7) but are consistent with the O2 / H2O2
couple (300 mV at pH 7), which is considered to control electrode measured
Eh under oxic conditions (Stumm and Morgan, 1996).
Aqueous chemistry data, shown in Fig. 3, demonstrate the consumption of TEAs
in order of decreasing nominal energetic yield, coupled to the oxidation of
labile OM. Upon physical removal and consumption of residual oxygen by
aerobic respiration, nitrate concentration decreased in the solution.
Decreases to nitrate concentration coincided with peaks of nitrite
concentration within the first hour of oxygen removal, indicative of
microbial denitrification. Subsequent increases to Mn(aq),
Fe(aq)2+ and HS(aq)- imply sequential reduction of
MnO2, Fe(OH)3 and SO42- as more energetically efficient
electron acceptors were depleted. Mn (predicted as Mn2+ by the
thermodynamic model) and Fe2+ were detected in solution earlier within
each subsequent anoxic cycle, but the apparent order of reduction
remained consistent across all five redox cycles (O2, NO3-,
NO2-, MnO2, Fe(OH)3, SO42-). The consistent order
and relative magnitude of reduction imply that the main biogeochemical
functioning of the sediment suspension did not change dramatically between
cycles during the experiment.
Aqueous chemistry and iron extraction data with time during reactor
experiments: the solid red line is Eh, solid black line is DO, solid
blue line is pH, full black circles are NO3-, white
squares are NO2-, inverted red triangles are Fe(aq),
inverted white triangles are Mn(aq), red
triangles are Fe2+ 0.5 M HCl extractable, white
triangles are Fe2+ / Fe3+ ratio in 0.5 M HCl extract, red
circles are S2- and black diamonds are DOC. Sampling points for
31P NMR and extracellular enzyme assays (EEA) are shown on the Eh
curve (31P NMR is indicated by open blue squares and EEA by open green circles).
Although relatively low concentrations of Fe2+ (up to 71 µM)
were measured in solution, 0.5 M HCl extractions revealed that significantly
greater Fe2+ was produced during each anoxic cycle than was measured in
the aqueous phase. Fe2+ generated by dissimilatory iron reduction has
been shown to sorb to mineral surfaces in sediment (Gehin et al., 2007; Klein
et al., 2010; Liger et al., 1999) or precipitate as ferrous carbonate (Jensen
et al., 2002), ferrous sulfide or other mixed ferrous/ferric phases (Rickard
and Morse, 2005). The 0.5 M HCl extractions targeted this sorbed or poorly
crystalline freshly precipitated Fe2+. During each anoxic cycle HCl
extractable Fe2+ concentration increased by
50–70 µmol g-1, equivalent to 12.31 to 17.29 mM of iron
reduction within the reactor as a whole. Thus, only 0.63 ± 0.4 % of
microbially reduced Fe2+ was measurable in solution. The aqueous phase
of the reactor was shown to be supersaturated with respect to mackinawite
(FeS), pyrite (FeS2), vivianite (Fe3(PO4)2 : 8H2O)
and siderite (FeCO3) during anoxic half-cycles, indicating thermodynamic
favorability for precipitation of these minerals. The kinetic constraints on
precipitation were not, however, considered. No significant cumulative change
to extractable Fe2+ / Fe3+ occurred after five full
reduction–oxidation cycles, indicating that solid-phase Fe redox cycling was
reversible, potentially due to rapid oxidation of solid Fe2+ in the
presence of O2 and carbonate (Caldeira et al., 2010).
DOC concentration also fluctuated systematically during oxic–anoxic cycles
(Fig. 3). Higher DOC concentrations were measured during anoxic conditions
than oxic conditions. DOC may be replenished by both enzymatic hydrolysis of
particulate organic matter (Vetter et al., 1998) and desorption of
mineral-associated OM (Grybos et al., 2009). The addition of algal matter at
the beginning of anoxic cycles resulted in observable sharp peaks of DOC,
which was rapidly removed from solution, probably due to a combination of
mineralization of labile DOC to HCO3- and sorption processes
(Chorover and Amistadi, 2001; Grybos et al., 2009). The peak of DOC supplied
by addition of algal matter represented labile DOC, which was readily
mineralized in comparison to the residual DOC, which persisted in the system
throughout the experiment. The differences in residual OM mobility between
oxic and anoxic cycles were unlikely due to oxide dissolution as differences
to DOC concentration are observed prior to increases in Mn and Fe
concentration in solution. We therefore postulate that solubility changes to
humified DOC were driven by pH changes (Fig. 3) between oxic and anoxic
conditions caused by changes in pCO2 between oxic and anoxic
conditions as previously shown in wetland sediments (Grybos
et al., 2009).
Lowest aqueous phosphorus concentrations (∼ 2.5–3 µM), shown
in Fig. 4, occurred during oxic half-cycles and increased dramatically during
anoxic half-cycles to a maximum concentration of 50–60 µM per
cycle, 88 % of which occurred as SRP. The range of TDP concentrations
within the aqueous phase of the reactor suspension is similar to that
reported in situ at the field site by Mayer et al. (2006). The timing of
phosphorus release to the aqueous phase corresponded well with increasing
Fe(aq)2+ concentration. This is reflected in a strong positive
correlation between TDP and Fe concentrations (n = 37, R2=0.93, p<0.0001). Increases to aqueous P concentration occurred only after
depletion of residual O2, NO3- and NO2-, after increases
to Mn(aq) and before increases to HS(aq)-. The timing
of P release is suggestive of an iron(III)-oxyhydroxide or ferric phosphate
control on phosphorus mobility (Bonneville et al., 2004; Hyacinthe and Van
Cappellen, 2004) and indicates that complete nitrate depletion was required
prior to phosphorus release to the aqueous phase.
Aqueous- and solid-phase phosphorus speciation from sequential
chemical extractions with time during the reactor experiment. White panels
correspond to time periods with air sparging, and grey panels correspond to time
periods with N2 : CO2 sparging. Black symbols are total P
concentration, and white symbols are SRP concentration.
Sequential chemical extractions and solid-phase P partitioning
The sum of P concentrations from all sequential extraction reservoirs
(61 ± 5 µmol g-1) was consistently within 10 % of a
total P extraction (57 ± 4 µmol g-1), indicating
acceptable analytical precision from the sequential extraction procedure.
Highest P concentrations were associated with the PHum
(∼ 26 % of TP) and PFe (∼ 24 %) fractions with
lower P concentrations in the PEx ∼ 16 %, PCFA
∼ 15 %, PDetr ∼ 12 % and
PRes ∼ 7 % fractions (Fig. 4). The largest variations in
P concentration were observed for PFe and decreased in the order
PFe (2.6) > PHum (1.5) > PEx
(0.4) > PDetr (0.4) > PResi
(0.35) > PCFA (0.25), where the numbers in parentheses are
standard deviations (µmol g-1). High variability over time
within the PFe and PHum fractions suggests that P was
exchanged to and from these fractions during redox oscillation, whereas
changes to the PEx, PResi and PCFA fractions
were comparatively insignificant (Fig. 4).
The PFe fraction was the only P pool in which concentration
consistently decreased during anoxic conditions and increased during oxic
conditions (Fig. 4). When a P mass balance (Fig. 5) was attempted to account
for increases to aqueous phosphorus (PAq) from the iron-bound
(PFe) pool during anoxic periods, it became evident that only
approximately 4.5 % of variability observed in the PFe pool
(total PFe variation of up to 4.5 µmol g-1 during
anoxic periods) was necessary to account for the changes to inter-cycle
PAq concentrations (50 µM). The remainder of
PFe lost during anoxic cycles appears to be reversibly
redistributed to the PEx (∼ 30 %) and PHum
(∼ 65 %) pools, which generally increased during anoxic conditions
and decreased during oxic conditions (Fig. 4).
According to Ruttenberg (1992) the PEx fraction corresponds to P
mobilized via the formation of MgPO4- complexes and(or) mass action
displacement by Cl-. It is therefore considered that MgCl2
effectively extracts P loosely associated with mineral surfaces. However,
Ruttenberg (1992) also demonstrated that plankton were efficiently extracted
by MgCl2 as well as ∼ 25 % of P associated with biogenic
CaCO3. Consequently, it is likely that P associated with microbial
biomass, CaCO3, and other loosely sorbed P contributed to the
PEx fraction. Slight increases to PEx concentration
during anoxic periods and corresponding decreases during oxic periods likely
reflect the combination of two processes: (1) equilibration between P surface
complexes and aqueous P species due to fluctuating aqueous concentrations
which were consistently higher during anoxic periods (Olsen and Watanabe,
1957); and (2) pH-driven sorption/desorption as pH was consistently slightly
higher during aerobic periods (Fig. 3), favouring desorption from mineral
surfaces including illite, which comprised 30 % of the crystalline
mineralogical fraction (Fig. 2c) (Manning, 1996).
The NaHCO3 extraction step was originally added to the SEDEX method to
target OM-associated P, which would otherwise be liberated during the
PFe extraction step (Baldwin, 1996). Baldwin noted a brown
coloration in PHum extracts and that absorbance at 250 nm was
positively correlated with SRP. Absorbance at 254 nm has been shown to be
indicative of aromatic OM, commonly associated with humic substances
(Weishaar et al., 2003). A light brown colour was also present in the
NaHCO3 extracts recovered during this experiment despite comparatively
low sediment OM content (Fig. 2b). Absorbance spectra for these extracts were
not determined.
Change in P distribution between the start and end of each oxic and
anoxic period (7-day change). Iron-bound P (PFe) appears to be
reversibly redistributed to the loosely sorbed (PEx) and humic-bound
(PHum) and aqueous fractions (PAq). The dashed line is
1 : 1. Linear regression of the data results in an R2 of 0.95, a slope
of -1.1 and p<0.0001).
Li et al. (2015) have recently demonstrated that the SEDEX PFe
extraction step co-extracted P associated with fine iron oxide–OM complexes
when a prior NaHCO3 step was not incorporated. However, Li et al. (2015)
also suggested that these complexes may be more recalcitrant than pure
minerals. Iron was present within the PHum extract at
concentrations between 18 and 25 µmol g-1, but the
original speciation of this iron is prior to extraction is unknown.
Chemically similar extractions used in soil sciences such as Hedley's
extraction (0.5 M NaHCO3, pH 8.5, 16 h) and the Olsen P test
(0.5 M NaHCO3, pH 8.5, 30 min) have been shown to extract Mg and Ca
phosphates as well as some organic P (Hedley et al., 1982; Olsen et al.,
1954). Approximately two-thirds of P extracted within PHum was present
as SRP, suggesting that approximately one-third of this fraction may be Po
species. The pH of the NaHCO3 extract used here was adjusted to 7.6 to
minimize the dissolution of Mg and Ca phosphates prior to the PCFA
extraction. Despite this, 30–44 µmol g-1 of Ca was extracted
within the PHum fraction. The origin of the extracted Ca is,
however, unknown and may have been complexed with OM or part of labile
Ca phosphates. It is still expected that the majority of Ca-phosphate
minerals were quantified as part of the PCFA or PDetri
extractions, which included ∼ 500–700 and
∼ 50–70 µmol Ca g-1 respectively.
Humic acids are known to compete with orthophosphate for surface-binding
sites on various minerals, including goethite (Sibanda and Young, 1986) and
poorly ordered Fe-oxides in the short term (Gerke, 1993). However, sorption
of natural OM to freshly precipitated Fe oxides may increase the long-term
sorption capacity of ferric oxides towards P by decreasing recrystallization
over time (Gerke, 1993) and through the formation of OM–Fe-P complexes
(Gerke, 1993). Although previous studies have provided evidence for ternary
complexes between ferric iron, OM and phosphate (Kizewski et al., 2010b)
there is currently no direct spectroscopic evidence for the existence of
mixed OM–Fe(III)-phosphate complexes in natural waters. Identification of
such complexes in natural environments is inherently challenging due to the
complexity of natural geochemical matrices (Kizewski et al., 2010a). Recent
studies have however successfully investigated the structure of synthetic
OM–Fe(III)-phosphate complexes (Kizewski et al., 2010a) and similar
OM–Fe(III)-arsenate complexes spectroscopically (Mikutta and Kretzschmar,
2011; Sharma et al., 2010). These studies suggest that similar and perhaps
more complex heterogeneous ternary complexes are also likely to be present in
natural freshwater environments (Kizewski et al., 2010a). This suggestion is
also supported by the observation that more than 80 % of soluble P in
some natural waters is associated with high molecular weight OM (Gerke,
2010). As spectroscopic characterization of the P associated with the
PHum fraction was not performed in this study, the PHum
pool is considered to represent a variety of OM-associated P in addition to
small amounts of P from labile Ca-phosphate minerals. OM-associated P
extracted within this fraction is likely coordinated with ferric iron
(18–25 µmol Fe g-1 co-extracted), Ca
(30–44 µmol Ca g-1 co-extracted) or Al
(0.8–1.5 µmol Al g-1 co-extracted). These metals may,
in turn, be associated with various mineral surfaces within the sediment.
Sequential extraction data demonstrate that the PHum fraction is
the dominant P fraction in all samples analysed, which highlights the
significance of this fraction. P mass balance also suggests that reversible
re-partitioning between this and the PFe fraction occurs during
redox condition changes. As the exact chemical nature of the PHum
fraction is not known, interpretation of concentration changes over time are
challenging. Speculatively, increases to the PHum fraction under
anoxic conditions may be due to the release of occluded OM–metal–P complexes
within iron(III) oxyhydroxides during reductive dissolution or simply
re-equilibration of solid-phase OM–metal–P complexes with increased aqueous
P.
We consider that the majority of P extracted within the PFe
fraction was co-precipitated with iron(III) oxyhydroxides which were
reductively dissolved by dithionite during the extraction (Ruttenberg, 1992).
This interpretation is supported by the relatively high concentrations of
iron extracted within the PFe fraction
(72–91 µmol Fe g-1). Aqueous Fe2+ produced during this
extraction is subsequently chelated by citrate and therefore solubility of Fe
and P is maintained. The bicarbonate component functions as a pH buffer to
ensure maximum preservation of apatite and CaCO3-bound P during this
reaction step (Ruttenberg, 1992).
Neither the PCFA, PDetri nor PResi fractions
varied systematically between oxic and anoxic conditions or changed
consistently during the course of the experiment, suggesting their
comparative stability during short periods of redox fluctuation. This is
supported by calculated SI for hydroxyapatite, which
remained between +0.86 and +6.24 for the duration of the experiment.
The P contribution from individual algal additions
(∼ 1.5 µmol P g-1) was relatively small compared to
total P in the reactor (61 µmol P g-1) and within the margin
of analytical error associated with solid extractions. Additionally, no
single fraction shows a clear increase over the course of the experiment,
and therefore quantification of the redistribution of P added with algal
additions is not possible.
(a) P speciation determined by 31P NMR for the initial
suspension (top left), the final suspension (bottom left) and the average of
samples from oxic conditions n=2 (top right), the average of samples from
anoxic conditions n=3 (bottom right). Pho-P is phosphonates,
Di-P is diester P and Mon-P is monoester P. Polyphosphate was not detected
at concentrations > 1 % in any of the samples analysed.
(b) Average extracellular enzyme activities under oxic and anoxic
conditions for MUP, DiMUP, PYRO-P and MUGb (n=5).
Fe : P ratios
Sequential extraction data (shown in Fig. 4), aqueous chemistry data (shown in
Fig. 3) and the correlation between aqueous Fe and P (n = 37,
R2 = 0.93, p<0.0001) all suggest that P released to solution
under anoxic conditions originated in the PFe pool. Although the
maximum molar ratio for phosphate incorporation within ferric oxides has been
shown to be 2 : 1 (Fe : P) (Thibault et al., 2009), it has been suggested
that much higher solid Fe : P of 15 (Jensen et al., 1992) to > 20
(Phillips et al., 1994) may be necessary to control phosphorus mobility under
oxic conditions. Results from bioreactor experiments suggest that phosphorus
is retained in the solid phase under oxic conditions at total Fe : P ratios
of just 4.1 : 1, potentially due to the association of P with other solid
sedimentary pools, particularly PHum. Fe : P ratios below the
stoichiometric limitation of 2 : 1, measured in the aqueous phase
(1.5–1.9), during anoxic conditions are therefore likely due to the removal
of Fe2+ from solution by secondary sorption and precipitation processes,
subsequent to reductive dissolution. Probable secondary Fe2+ removal
processes include the formation of amorphous FeS (SI of up to +2.27 for
mackinawite) (Rickard and Morse, 2005) and sorption of Fe2+ to clays
(Gehin et al., 2007; Klein et al., 2010). This is supported by increases to
Fe2+ / Fe3+ ratios in the 0.5 M HCl extractable fraction
during anoxic conditions. Frequent rapid reoxidation of ferrous sulfide due
to air sparging in the reactor experiment, and extensive in situ bioturbation, likely prevents formation of more recalcitrant and slow-forming iron
sulfide minerals such as pyrite, despite strong thermodynamic favorability
for pyrite formation (SI up to +13.13) (Caldeira et al., 2010; Peiffer et
al., 2015). This is consistent with the results of XRD analysis, which did
not identify pyrite (Fig. 2c). A similar effect has been previously
demonstrated in lake sediments (Gächter and Müller, 2003).
Hydrolytic enzyme activities
The activities of model phosphomonoesterase, phosphodiesterase and
pyrophosphatase were found to be systematically higher under oxic conditions
compared to anoxic conditions by 37 % (p<0.005), 8 % (not
significant) and 24 % (p = 0.08) respectively (Fig. 6b).
Phosphomonoesterases were found to have the highest activities despite the
inherent overestimation of phosphodiesterase activities when using MUF-tagged
substrates (Sirová et al., 2013). The opposite trend was observed for
glycopyranoside, part of the cellulose degradation pathway (Dunn et al.,
2013), which showed consistently higher activity (69 % p=<0.05) under
anoxic conditions (Fig. 6b). The different trends exhibited by cellulose and
phosphatase enzymes indicate that changes in activity were not universal but
specific to enzyme function. Phosphomonoesterase activities obtained in the
current study (1.76–2.4 mmol h-1 kg-1) are similar to those
previously reported in wetland sediments (Kang and Freeman, 1999) and suggest
the capacity for rapid hydrolysis of Po species in necromass.
Lowering of the water table in wetlands has previously been shown to increase
the activity of phosphatase enzymes and the hydrolysis of Po
species (Song et al., 2007). However, water table fluctuation results in
concomitant changes to moisture content and redox conditions, which prevents
isolation of the causal variable in field investigations (Rezanezhad et al.,
2014). Therefore, this is, to the best of our knowledge, the first direct
demonstration of phosphatase activity changes in response to changing redox
conditions. We postulate that under anoxic conditions, when phosphorus
availability in the aqueous phase is high, production of extracellular
phosphatase enzymes by the microbial community is down-regulated. Conversely,
when bioavailable phosphorus is removed from solution under oxic conditions,
extracellular phosphatase production is up-regulated in response. Adjustments
to enzyme production in response to changes in phosphate availability must
occur on short timescales (hours/days) for such trends to be observable
during the bioreactor experiment. An inverse relationship between phosphatase
activities and phosphate concentration has previously been shown spatially
in wetlands by Kang and Freeman (1999) but to our knowledge never temporally
in sediments.
31P NMR
Results from 31P NMR analyses (Fig. 6a) show that the majority of
phosphorus was present in the solid phase as orthophosphate (84–91 %)
with 4–8 % monoester P, 3–8 % diester P and < 1 %
phosphonates and polyphosphates with no clear trend in relative abundance
emerging during the experiment. The NaOH–EDTA extraction resulted in a
recovery of ∼ 27 % of TP, which is comparable to previous studies
with carbonate-buffered soils and sediments (Hansen et al., 2004; Turner et
al., 2003a). Alpha and beta glycerophosphates are commonly identified in
monoester spectral regions and have been demonstrated to be products of
diesters degraded during analysis (Doolette et al., 2009; Jørgensen et
al., 2015; Paraskova et al., 2015). As no glycerophosphates were identified
in any of the analysed samples, recalculation of monoester / diester
ratios was not performed. A higher mean monoester / diester ratio (2.31)
was found in reduced samples than oxidized samples (0.97) (a statistically
significant difference, p=0.04). This difference could indicate either that
monoester P was less efficiently extracted under oxic conditions due
to sorption to metal oxides or that monoesterase/diesterase activity
decreased under anoxic conditions, which is consistent with enzymatic
activity assays (Fig. 6b). Total Po determined by 31P NMR
varied between 9 and 16 % over time compared to 5–11 % in the
Presi from sequential extractions, indicating that not all
Po was extracted in the Presi fraction, which is commonly
referred to as the organic-P fraction. We postulate that the remaining
∼ 5 % of total phosphorus, identified as Po by 31P
NMR, was extracted during previous steps in the sequential extraction scheme,
particularly MgCl2, which has been shown to efficiently extract P
associated with microbial biomass (Ruttenberg, 1992) and NaHCO3. The
relative activities of phosphatase enzymes appear to correlate with the
relative abundances of Po species identified by 31P NMR, e.g.
monoesters > diesters > pyro-P.
Significant polyphosphate (> 1 %) was not detected by 31P NMR
during experiments. Previous studies focusing on WWTP tertiary treatment for
phosphate removal suggest that redox oscillating conditions promote
intracellular poly-P accumulation during aerobic conditions to be used as an
energy store under anoxic conditions in order to uptake short-chain fatty
acids (SCFA) in the absence of an electron acceptor (Hupfer et al., 2007;
Wentzel et al., 1991). Phosphate uptake during aerobic conditions therefore
requires P availability in excess of what is required for growth and
maintenance of the microbial community. However, phosphate availability under
aerobic conditions is limited by sorption and co-precipitation with iron
oxides, assuming sufficient Fe : P, despite high total phosphorus
concentration in sediment. The P requirements by the microbial community are
also likely to be high during the transition to aerobic conditions due to the
availability of O2 as an energetically efficient electron acceptor and
fermentation products (SCFA), further decreasing the probability of
polyphosphate accumulation. Additionally, polyphosphate accumulation and
release has shown to be inhibited by denitrification and sulfate reduction
due to competition for SCFA (Kortstee et al., 1994; Yamamoto-Ikemoto et al.,
1994).