Applied Geochemistry 26 (2011) 688–695
Contents lists available at ScienceDirect
Applied Geochemistry
journal homepage: www.elsevier.com/locate/apgeochem
Isotopic evidence for the source and fate of phosphorus in Everglades
wetland ecosystems
Xin Li a,b, Yang Wang b,⇑, Jennifer Stern b,c, Binhe Gu d
a
Institute of Hydrobiology, Jinan University, Guangzhou, Guangdong 510632, China
Department of Earth, Ocean and Atmospheric Science, Florida State University & National High Magnetic Field Laboratory, Tallahassee, FL 32306-4100, USA
c
NASA Goddard Space Flight Center, Greenbelt, MD 20771, USA
d
Everglades Division, South Florida Water Management District, West Palm Beach, FL 33406, USA
b
a r t i c l e
i n f o
Article history:
Received 26 November 2009
Accepted 11 January 2011
Available online 15 January 2011
Editorial handling by W.B. Lyons
a b s t r a c t
Phosphorus has historically been a limiting nutrient in the Florida Everglades. Increased P loading to the
Everglades over the past several decades has led to significant changes in water quality and plant communities. Stormwater runoff that drains agricultural lands and enters the Water Conservation Areas
(WCAs) are known to contain elevated levels of P, but the exact source of this P has not been fully determined. Here the results of an O isotope study of dissolved inorganic phosphate (DIP) in both polluted and
relatively pristine (or reference) areas of the Everglades are reported. The data reveal spatial and temporal variations in the d18O signature of DIP, reflecting the source and the degree of cycling of P. The d18O
values of DIP collected from the Everglades National Park were close or equal to the predicted d18O values
of DIP formed in situ in equilibrium with ambient water, indicating that P is quickly cycled in the water
column in oligotrophic ecosystems with very low P concentrations. However, most DIP samples collected
from areas impacted by agricultural runoff yielded d18O values that deviated from the predicted equilibrium DIP–d18O values based on the d18O of water and water temperature, suggesting that biological
cycling of P was not rapid enough to remove the fertilizer d18O signature in the DIP pool from areas
receiving high P loading. The d18O signature of DIP in impacted areas reflects a mixing of fertilizer P
and biologically cycled P, where the relative proportions of biologically cycled vs. fertilizer DIP are controlled by both biological (microbial activities and plant uptake) and hydrologic factors (loading rate and
residence time). Using a two-end-member (i.e., fertilizer P and biologically cycled P) mixing model, fertilizers were estimated to contribute about 15–100% of the DIP pool in the highly impacted areas of the
northern Everglades, whereas the DIP pool in the reference (i.e., relatively pristine) wetlands in the Everglades National Park was dominated by biologically cycled P. The study shows that O isotopic measurements of dissolved PO3
4 can be a useful tool for tracing the fertilizer P inputs to freshwater ecosystems.
Ó 2011 Elsevier Ltd. All rights reserved.
1. Introduction
Nutrient loading in wetlands is the subject of much concern due
to human dependence on the benefits offered by a healthy wetland. Wetlands serve to store and filter water, provide habitats
for a diverse array of species, and support recreational activities
such as fishing. Many coastal and freshwater wetlands have experienced eutrophication due to poor land-use and the contributions
of excessive nutrients from agricultural or urban runoff upstream
(Carpenter et al., 1998). Eutrophication of freshwater wetlands is
problematic as it causes a reduction of biodiversity and the invasion of exotic species (Sklar et al., 2005). While the sources of these
excess nutrients can sometimes be traced, the fate of these nutrients upon entering the wetland system is poorly understood.
⇑ Corresponding author. Fax: +1 850 644 0827.
E-mail address: [email protected] (Y. Wang).
0883-2927/$ - see front matter Ó 2011 Elsevier Ltd. All rights reserved.
doi:10.1016/j.apgeochem.2011.01.027
Phosphorus is a limiting nutrient in most freshwater ecosystems
(Jickells, 1998). Unlike C and N, P added to an aquatic ecosystem or
released during the decomposition of organic matter usually stays
within the system, resulting in an enrichment of P in detritus and
surface soil/sediment (Reddy et al., 1999). Phosphorus is present
in the water column as particulate organic P (POP), dissolved organic P (DOP), and dissolved inorganic P (DIP). DIP is the most bioavailable form of P and may be quickly taken up by organisms. It may
also be sequestered in sediments/soils by adsorption to or reaction
with Fe, Al, Ca and Mg minerals. Soil P undergoes various transformations as it cycles through inorganic P pools (associated with Fe,
Al, Ca and Mg minerals) and organic P pools (including plants,
animals, microbes, and soil organic matter) (Richardson, 1999).
Microbes transform organic P from decaying organic matter into
bioavailable inorganic P. These processes are largely governed by
nutrient content (C, N and P) of the soil and water as well as by
the size of the microbial pool (Richardson, 1999; Reddy et al., 1999).
X. Li et al. / Applied Geochemistry 26 (2011) 688–695
Phosphorus loading is a particular problem in the Florida Everglades, historically a P-limited wetland (Davis, 1994; Reddy et al.,
1999). The runoff from the Everglades Agricultural Area is high in
concentrations of inorganic P (on average 20 times higher than
P-levels in the relatively pristine Everglades), directly influencing
water quality throughout the Everglades. Plants and algae endemic
to the Everglades are adapted to low P concentrations, with organic
P accounting for the largest fraction of soil P in this area (Koch-Rose
et al., 1994). Typically, inorganic P exists only in very small
quantities in natural waters; natural cycling ensures that P is efficiently utilized during photosynthesis by higher plants and algae
and subsequently regenerated by microbial action for later use
(Poister et al., 1994; Caraco et al., 1992; Benitez-Nelson and
Buesseler, 1999). However, when P loading occurs due to the use
of PO4-fertilizer, a surplus of bioavailable P (mostly inorganic) is
689
created, resulting in changes in water quality and higher plant
assemblages.
The results of this P enrichment have been well documented in
the Everglades. Perhaps the most noticeable ecological change has
been the vegetation shift from P-limited sawgrass to P-adapted
cattail stands in the more polluted areas of the marsh (Davis,
1994). In addition, increased net primary productivity and P storage by wetland vegetation (Craft et al., 1995; Chiang et al., 2000),
increased decomposition of detritus (Davis, 1991), and increased
organic soil accretion (Craft and Richardson, 1993, 1998) have
been documented. Increased decomposition rates indicate a faster
rate of microbial cycling in the marsh. The decomposition of
organic matter by microbes remineralizes organic P, making it
bioavailable to plants and further increasing the production of
organic matter.
Fig. 1. Map showing the study sites in the Everglades. Stromwater Treatment Area-1 West (STA-1W), which represents the highly polluted environment in the Everglades, is
routinely sampled by SFWMD for total P and for other water quality parameters. Sites in the Everglades National Park (ENP) are Pa-hay-okee and Nine Mile Pond, representing
the relatively pristine environment in the Everglades.
690
X. Li et al. / Applied Geochemistry 26 (2011) 688–695
This study was conducted to investigate the O isotopic systematics of dissolved PO3
4 and water in both polluted and relatively
pristine (or reference) wetlands in the Florida Everglades. The
objectives were (1) to determine whether the O isotopes in dissolved PO3
(DIP) could be used as ‘‘fingerprints’’ for identifying
4
the source of DIP in the Everglades wetland ecosystems and (2)
to assess the role of microbial cycling in the P cycle. Comparison
of the O isotopic compositions of DIP to the O isotopic composition
of water allows determination of whether microbial cycling of P results in a complete or partial re-equilibration of DIP oxygen with
the O in ambient water, providing insights into the time scales over
which P is cycled. Investigation of the source/fate of P and the
underlying processes controlling the P concentration will allow
characterization and improved understanding of the P cycle in
aquatic ecosystems.
Although the effects of P loading and hydrology on plant communities are clearly evident, very little is known about how these
factors influence the biogeochemical processes that regulate nutrient availability and cycling in impacted and non-impacted areas
(Reddy et al., 1999; Richardson, 1999). The source of the P in these
areas has also not been fully determined. It may come from the dissolution of fertilizers applied in the agricultural land and/or from
the accelerated decomposition of organic matter in the peat soils
of the agricultural areas due to drainage that has transformed
the soil environment from anaerobic to aerobic (Wang et al.,
2002; Stern et al., 2007).
In the natural environment, P is primarily found as orthophosphate (PO3
4 ) and its derivatives (Bieleski, 1973). The P–O bond in
PO3
is
resistant
to hydrolysis in inorganic systems. Studies
4
(Kolodny et al., 1983; Shemesh et al., 1983, 1988) have shown that
the O in PO3
4 retains its original isotopic signature even after harsh
chemical treatments and is practically inert to isotopic exchange
with water in low temperature inorganic chemical systems
(Lecuyer et al., 1999). However, the P–O bond can be easily broken
in enzyme-mediated biochemical reactions, resulting in rapid O
isotope exchange with surrounding water by organisms (Boyer,
1978; Bieleski, 1973; Blake et al., 1997, 2005; Paytan et al.,
2002). The PO3
precipitated in microbial culture experiments
4
has also been shown to be in O isotopic equilibrium with the
surrounding water (Blake et al., 1997, 1998). Recent field studies
suggest that the O isotopic composition of dissolved PO3
4 can be
used as a tracer to determine the source and the degree of cycling
of DIP in marine, estuarine and coastal environments (Colman
et al., 2000, 2005; Colman, 2002; McLaughlin et al., 2004,
2006a,b; Blake et al., 2005; Jaisi and Blake, 2010) and possibly in
freshwater ecosystems (Stern and Wang, 2002; Stern, 2005).
filter
0.45 µ
freshwater
MgCl2
NaOH
Mg(OH)2
DIP adsorbed
centrifuge
discard
supernatant
2. Study sites
The study area is located in the Everglades of south Florida
(Fig. 1). The Everglades region of Florida encompasses most of
the southern Floridian peninsula and represents the largest subtropical freshwater wetland ecosystem in the USA. Prior to settlement, the hydrology of this region was controlled by seasonal
cycles of rainfall causing sheetflow from Lake Okeechobee southward to the Florida Bay and flooding low lying south Florida. The
hydrologic regime of the Everglades has since been drastically altered. The once expansive freshwater marshes (known as the ‘‘River of Grass’’) are now dissected by drainage canals, levees and
water control structures into many sub-basins, including the Everglades Agricultural Area (EAA) which is now completely drained,
the Water Conservation Areas (WCA-1, WCA-2 and WCA-3), and
Mg(OH)2DIP residue
10M HNO3
acetic acid + KOH
Dissolved DIP
bufffered solution
pH~5.5
10M HNO3
dissolved DIP
NaOH
0.2 µ filter
dissolved DIP
cerium
phosphate
0.5g cerium
nitrate
centrifuge
discard
supernatant
cerium
phosphate
in 50mL tube
rinse
cerium
phosphate
precipitate
10 times
0.5M
CH3COOK
0.5g cerium nitrate
centrifuge
recycled process
one direction process
phosphate in
solution pH~5.5
discard
supernatant
drops of NH4OH
phosphate
in solution
seperate resin
retain solution
batch
seperation of
cerium ions
from solution
cation exchange
resin
BIORAD AG50
cerium
phosphate
dissolved in
0.2M HNO3
1M HNO3
DI water
cerium
phosphate
in 50mL tube
NH4OH
2M NH4NO3
0.5g AgNO3
silver phosphate
pH~8
vacuum filter
silver phosphate on filters
dry in oven at 50ºC
To TC/EA mass spectrometer
for oxygen isotope analysis
Fig. 2. Schematic diagram showing the procedure used to extract dissolved inorganic P (DIP) from water for O isotopic analysis (modified from McLaughlin et al., 2004).
691
X. Li et al. / Applied Geochemistry 26 (2011) 688–695
d18O (‰)
Note
Average
d18O
Standard
deviation (±r)
FTL-1C
FTL-1D
EAA7
EAA8
Esp 5
Esp 6
Scotts 6
Kmart1
Kmart2
Sunni-1
Sunni-2
St-Gr 6
St-Gr 7
P-Pro 3
P-Pro 4
24.4
25.3
23.9
23.9
23.4
23.2
24.6
23.3
24.9
23.8
23.7
24.7
25.3
20.9
20.9
EAA fertilizer
EAA fertilizer
EAA fertilizer
EAA fertilizer
Esposito’s
Esposito’s
Scotts
Kmart
Kmart
Sunniland
Sunniland
Stay green
Stay green
Peter’s professional
Peter’s professional
24.4
0.7
23.3
0.2
24.1
1.2
23.8
0.1
25.0
0.4
20.9
0.0
Average
23.8 ± 1.3
O (‰ vs. VSMOW)
26
24
22
20
18
Surface water samples for DIP analysis were collected in April
and July of 2005 and in March of 2006 using the MAGIC method
(Karl and Tien, 1992). The MAGIC method removes P from water
via co-precipitation with Mg(OH)2. The water samples were filtered through 5 and 0.45 lm cartridge filters consecutively and
the filtrate collected in a pre-cleaned bucket. Aliquots of 150 mL
NaOH and 100 g MgCl2 were added into 8 L of the filtered water
and the Mg(OH)2 flocculate was allowed to settle for 2 h. After
siphoning off the overlying water, the Mg(OH)2 precipitate was
transferred into a 1 L bottle. For every 8 L of water, two 1 L bottles
of Mg(OH)2 precipitate were collected. These bottles were stored in
a cooler with ice and were processed for O isotopic analysis of PO3
4
after arriving in the lab. Water samples were also collected after filtering through a 0.45 lm filter for O isotopic analysis for comparison with the O isotope ratios in DIP.
DIP in the sample was isolated in the laboratory as Ag3PO4 by
using the procedure shown in Fig. 2. The procedure is modified
from a method developed by McLaughlin et al. (2004). Although
the method described in McLaughlin et al. (2004) has been successfully applied to measure the O isotope ratios of DIP in seawater
and estuary waters (McLaughlin et al., 2006a,b), it is not directly
applicable to organic-rich freshwaters in the Everglades (Li,
2009). The modified procedure allows for the extraction of DIP
from freshwater with high concentrations of dissolved organic
matter for accurate O isotope measurements (Li, 2009). Due to
small sample size, Ag3PO4 yield was not determined for every sample. The Ag3PO4 yield based on selected samples ranged from 92%
to 98% and had no effect on the measured d18O values. The Ag3PO4
was then analyzed for O isotopic composition by using a High Temperature Conversion Elemental Analyzer (TC-EA) connected to a
Finnigan MAT Delta Plus XP continuous-flow isotope ratio mass
spectrometer (IRMS) at Florida State University (FSU). Oxygen liberated by the decomposition of Ag3PO4 at 1450 °C was converted to
CO by reaction with graphitic and glassy carbon inside the TC-EA
reactor. The CO gas was then analyzed for mass 30/28 ratio against
a reference CO gas. At least two sets of three different standards
were run in triplicate with each batch of samples. Fertilizer samples were prepared for O isotope analysis using an older method
(O’Neil et al., 1994). This involved dissolving the fertilizer in dilute
Sample ID
DIP-
3. Methods
Table 1
d18O of phosphate in commercial fertilizers.
18
STA-1w, Inlet
STA-1W, interior
STA-1W, outlet
ENP (Pa-hey-okee)
ENP (Nine Mile Pond)
16
Apr 05
Jun 05
Aug 05
Oct 05 Dec 05
Feb 06
Apr 06
Feb 06
Apr 06
Sampling Date
300
P concentration ( µg/L)
the Everglades National Park (ENP). Stormwater Treatment Areas
(STAs), which are constructed wetlands, have been established by
the South Florida Water Management District (SFWMD) to treat
runoff from the EAA and have proven effective in reducing P in
effluent waters. The study sites selected for this project are located
in the Storm Water Treatment Area-1 West (STA-1W), part of
which was previously known as the Everglades Nutrient Removal
Project (i.e., ENR), and the ENP (Fig. 1).
Samples were collected for chemical and isotopic analyses from
the supply canal, the inlet, the interior and the outlet of STA-1W,
and from two localities (Pa-hay-okee and Nine Mile Pond) within
the ENP (Fig. 1). These sites represent two contrasting biogeochemical conditions in the Everglades: highly polluted aquatic systems
near the EAA and relatively pristine wetlands in the ENP (Fig. 1).
The STA-1W is routinely sampled by SFWMD for total P analysis
and for measuring other water quality parameters. The inlet area
of STA-1W is highly impacted by agricultural runoff and directly
receives runoff water from the canal. The P level at this site is
typically greater than 120 lg/L. Sites in the ENP represent ‘‘unpolluted’’ or reference conditions that are least affected by agricultural
runoff and typically have P levels <10 lg/L (Stern et al., 2007).
These sites provide an excellent opportunity to compare relatively
pristine and impacted ecosystems.
STA-1W, Inlet, P
STA-1W, interior, P
STA-1W, outlet, P
ENP (Pa-hay-okee), P
ENP (Nine Mile Pond), P
250
200
150
100
50
0
Apr 05
Jun 05
Aug 05
Oct 05
Dec 05
Sampling Date
Fig. 3. Spatial and temporal variations in DIP-d18O and P concentration in Storm
Water Treatment Area-1 West (STA-1W) and Everglades National Park (ENP).
HNO3 and treating the solution with HF to remove Ca2+ cations. Silver ammine solution was added and samples were heated at no
more than 60 °C for approximately 4 h to precipitate Ag3PO4. This
solid was then combusted with a stoichiometric amount of graphite in a sealed quartz tube at 1200 °C to produce CO2 for analysis on
a Finnigan Delta S dual-inlet IRMS at FSU (O’Neil et al., 1994). All
phosphate samples were analyzed in triplicate and had a precision
(1r) of ±0.7‰ or better.
Phosphorus concentrations were measured using a single
collector ICP-MS at the National High Magnetic Field Laboratory,
Florida State University (FSU). The O isotopic compositions of
692
X. Li et al. / Applied Geochemistry 26 (2011) 688–695
water samples were determined using a Gas Bench Auto-WaterEquilibration device connected to the IRMS and the analytical precision (based on replicate analyses of standards processed with
each batch of samples and on sample replicates) is ±0.1‰ (1r).
The O isotope data are reported in the standard d notation relative
to Vienna Standard Mean Ocean Water (VSMOW) as:
d18 O ¼
18
O=16 Osample
1
1000 ð‰Þ
18 O=16 O
VSMOW
4. Results and discussion
The major sources of P in the Everglades area include fertilizers
directly leached from soils in the Everglades Agricultural Area
(EAA) and biologically cycled P derived from the decomposition
of organic matter. The analysis of commercial fertilizers shows that
fertilizer phosphate d18O values range from 20.9‰ to 25.0‰, with a
mean d18O of 24 ± 1‰ (Table 1), and are very similar to those of
marine phosphates (Longinelli and Nuti, 1968; Shemesh et al.,
1988). This is expected because fertilizer phosphate is manufactured from marine apatite.
The d 18O values of DIP are affected by multiple factors such as
the sources of P, the degree of in situ biological cycling of P due to
microbial action, temperature and the d18O of the ambient water.
The d18O value of the fertilizer used in the EAA is 24.4 ± 0.7‰
(Table 1). The d18O values of biologically cycled DIP (which is
assumed to be in equilibrium with environmental water) at the
study sites can be calculated from the d18O of water and water
temperature using the following temperature – O isotope fractionation equation for PO3
and water established by Longinelli and
4
Nuti (1973):
Tð CÞ ¼ 111:4 4:3ðd18 OP d18 OW Þ
ð1Þ
18
18
where T is temperature of PO3
4 formation; d OP and d OW are the
d18O of PO3
and
water,
respectively.
4
The data show significant temporal and spatial variations in the
18
d O of DIP and P concentrations in the Everglades ecosystems
Table 2
d18O values of DIP and water, total P concentration, water temperature and estimated contribution of fertilizer P to the DIP pool at the study sites.
a
b
Sample ID
Sampling location
Sampling
date
d18ODIP
(‰)
Standard
deviation (±r)
Water
d 18O
P level
(lg/L)
Average
water T (°C)
Estimated
%fertilizer P
LX02-02
LX02-03
LX02-19
LX02-20
LX02-07
LX02-08
LX02-13
LX02-14
LX03-02
LX03-04
LX03-05
LX03-09
LX03-10
LX03-11
LX03-12
LX03-14
LX03-17
LX03-19
LX03-21
LX03-23
LX03-25
LX03-26
LX03-27
LX04-2-14
LX04-16-28
LX04-30
LX04-32
LX04-35
LX04-36
STA-1W, inflow G302
STA-1W, inflow G302
C-51 (canal) near STA-1W inlet
C-51 (canal) near STA-1W inlet
STA-1W, interior, G253
STA-1W, interior, G253
STA-1W, outflow G310
STA-1W, Outflow G310
STA-1W, outlet, G310
STA-1W, outlet, G311
STA-1W , outlet, G312
STA-1W, interior, G253
STA-1W, interior, G253
STA-1W, interior, G253
STA-1W, interior, G253
STA-1W, interior, G253
STA-1W, inlet
STA-1W , inlet
STA-1W, inlet
Canal, near STA-1W inlet
Canal, near STA-1W inlet
Canal, near STA-1W inlet
Canal, near STA-1W inlet
STA-1W G302
Canal
Pa-hay-okee
Pa-hay-okee
Nine Mile Pond
Nine Mile Pond
4/13/05
4/13/05
4/13/05
4/13/05
4/13/05
4/13/05
4/13/05
4/13/05
7/28/05
7/28/05
7/28/05
7/28/05
7/28/05
7/28/05
7/28/05
7/28/05
7/28/05
7/28/05
7/28/05
7/28/05
7/28/05
7/28/05
7/28/05
3/7/06
3/7/06
3/8/06
3/8/06
3/9/06
3/9/06
21.4
21.4
21.7
21.5
22.5
22.2
21.7
21.6
21.6
20.6
20
20.1
20.6
21.2
21.6
20.9
21
21.1
20.2
20.7
22
20.4
21.2
25.2
25.5
25.1
24.5
21.6
22.7
0.2
0.2
0.3
0.6
0.6
0.5
0.5
1.6
1.6
0.5
0.5
0.8
0.8
0.8
0.6
0.6
0.6
0.6
0.6
0.2
0.2
0.2
0
0
0
0
0.3
0.3
2.9
2.9
0.8a
0.8a
113.5
113.5
23 ± 2
23 ± 2
23 ± 2
23 ± 2
23 ± 2
23 ± 2
23 ± 2
23 ± 2
30 ± 1
30 ± 1
30 ± 1
30 ± 1
30 ± 1
30 ± 1
30 ± 1
30 ± 1
30 ± 1
30 ± 1
30 ± 1
30 ± 1
30 ± 1
30 ± 1
30 ± 1
21 ± 2
21 ± 2
21 ± 2
21 ± 2
21 ± 2
21 ± 2
15
15
22
15
21
2
22
19
46
22
7
14
26
40
49
33
45
47
29
37
63
31
47
100
100
0
0
0
1
0.3
0.3
0.2
0.1
0.2
0
0.2
0.2
0.2
0.2
0
0.1
0.1
0.2
0.4
0.1
0.2
0.2
0.2
0.2
0.1
0.1
0.1
0.2
60.9
60.9
82.7
82.7
27.3
27.3
27.3
131.2
131.2
131.2
131.2
131.2
71.7
71.7
71.7
137.7
137.7
137.7
137.7
90
153.2
<10b
<10b
<10b
<10b
Water d18O data from previous years.
From South Florida Water Management District & USGS database.
Table 3
Two-tailed t-test results for significant differences between different locations and times.
ENP P levels vs. STA-1W P levels
ENP DIP-d18O (2006) vs. STA-1W Inlet DIP-d18O (April and July 2005)
ENP DIP-d18O (March 2006) vs. STA-1W inlet DIP-d18O (April 2005)
ENP DIP-d18O (March 2006) vs. STA-1W inlet DIP-d18O (July 2005)
STA-1W inlet DIP-d18O (April and July 2005) vs. STA-1W inlet DIP-d18O (March 2006)
STA-1W Inlet DIP-d18O (April 2005) vs. STA-1W inlet DIP-d18O (March 2006)
STA-1W inlet DIP-d18O (July 2005) vs. STA-1W inlet DIP-d18O (March 2006)
STA-1W DIP-d18O (April 2005) vs. STA-1W inlet DIP-d18O (March 2006)
STA-1W DIP-d18O (July 2005) vs. STA-1W inlet DIP-d18O (March 2006)
STA-1W DIP-d18O (April 2005) vs. STA-1W DIP-d18O (July 2005)
Mean difference
df
t
p
Significant difference at 95%?
88
2.3
2
2.5
4.2
3.9
4.4
3.6
4.5
0.9
25
13
6
9
11
4
7
8
15
21
4.3
4.4
2.4
3.85
10.5
27.4
9.9
12.04
10.4
3.7
0.00022
0.00074
0.05
0.004
<0.0001
<0.0001
<0.0001
<0.0001
<0.0001
0.0012
Yes
Yes
Yes
Yes
Yes
Yes
Yes
Yes
Yes
Yes
693
X. Li et al. / Applied Geochemistry 26 (2011) 688–695
ðd18 ODIP d18 Orecycled Þ
18
18
ðd Ofertilizer d Orecycled Þ
April 13, 2005
DIP-
18
O (‰ vs. VSMOW)
25
100ð%Þ;
ð2Þ
where F is the estimated percentage of DIP derived from fertilizer;
d18ODIP, d18Orecycled and d18Ofertilizer are the d18O of DIP, biologically
cycled DIP and fertilizer DIP, respectively. It is important to note
that this simple model would not be appropriate for systems dominated by organic P.
As shown in Fig. 5, the amount of fertilizer-derived DIP varied
spatially and temporally, likely controlled by the timing of fertilizer applications in the EAA and the amount of runoff from the
agricultural lands as well as the biological activities in the area.
The fertilizer-derived DIP generally decreased in the downstream
direction from STA-1W inlet, interior, to STA-1W outlet and to
the ENP (Fig. 5). Fertilizer P accounted for about 15% to nearly
100% of the DIP in the supply canal and at the inlet of the
24
23
22
21
23°C
20
STA-1W, April 2005
T=21°C (Longinelli & Nuti, 1973)
T=23°C (Longinelli & Nuti, 1973)
T=25°C (Longinelli & Nuti, 1973)
19
18
-4
-3
-2
-1
0
1
2
3
4
18
Water
O (‰ vs. VSMOW)
(b) 26
July 28, 2005
22
20
DIP-
18
O (‰ vs. VSMOW)
24
30°C
18
STA-1W, July 2005
T=29°C (Longinelli & Nuti, 1973)
T=30°C (Longinelli & Nuti, 1973)
T=31°C (Longinelli & Nuti, 1973)
16
-4
-2
0
Water
(c)
2
4
6
18
O (‰ vs. VSMOW)
26
March 7-9, 2006
O (‰ vs. VSMOW)
25
DIP-
F¼
(a) 26
18
(Fig. 3, Tables 2 and 3). The d18O values of DIP samples from STA1W and ENP ranged from 20.0‰ to 25.5‰ (Fig. 3). The P concentration varied from <10 to 138 lg/L, with highest P levels found in the
STA-1W and its supply canal near the EAA in the northern Everglades and lowest P concentration in the ENP (Table 2 and
Fig. 3). In the same season, P concentration normally decreased
along the flow direction from the STA-1W supply canal, the inlet,
to the outlet due to the removal of P from the water by plant uptake and adsorption onto Fe, Al, Ca and Mg minerals in sediment/
soil (Reddy et al., 1999). There was no clear relationship demonstrated between d18O values of DIP and P concentrations (Fig. 3).
As shown in Fig. 3, the d18O values of DIP were generally lower
in the summer than in the winter at a given site, reflecting the d18O
of water, water temperature and the sources of DIP. However, if the
DIP pool consists entirely of biologically cycled DIP, the d18O values
of DIP should only reflect the d18O of ambient water and temperature as described by Eq. (1). Water temperatures for the sampling
period (2005–2006) were 21 ± 2 °C (with a range of 18–23 °C),
23 ± 2 °C (20–26 °C), and 30 ± 1 °C (29–32 °C) for February–March,
March–April, and July, respectively (USGS Surface Water Database:
http://waterdata.usgs.gov/nwis/dvstat/). Comparison of measured
d18O values of DIP with predicted d18O values of biologically cycled
DIP (using Eq. (1)) shows that the DIP samples collected in July
2005 and March 2006 from the STA-1W were not in O isotopic
equilibrium with ambient water (Fig. 4b and c), suggesting that
d18O values of DIP were influenced by fertilizer d18O, and that biological activity was not sufficient in these highly impacted areas to
completely remove the fertilizer signature. DIP samples from April
2005 from STA-1W appeared to be near or in O isotopic equilibrium with ambient waters (Fig. 4a). In the reference sites (i.e., relatively pristine wetlands) in the ENP, the limited data indicate that
the DIP samples collected from Pa-hay-okee (a marsh area with
sawgrass growing in water less than 50 cm deep) and Nine Mile
Pond (with deeper water) were in or close to O isotopic equilibrium with ambient water (Fig. 4c), likely due to low P concentration and faster P cycling. This is consistent with radiocarbon data
from the area that suggest a rapid cycling of C and nutrients predominantly controlled by biological processes in the ENP as evidenced by the ‘‘modern’’ radiocarbon signatures of dissolved
organic matter (DOM) in the system (Stern et al., 2007). In comparison, in STA-1W, a significant source of C and nutrients was runoff
from the agricultural land as indicated by the ‘‘old’’ radiocarbon
ages of DOM (Wang et al., 2002; Stern et al., 2007).
Assuming DIP is a mixture of biologically cycled P with a d18O
value reflecting equilibrium with ambient water and fertilizer P
with a d18O value of 23.8‰, the relative contribution of fertilizer
to the DIP pool (Table 2) was estimated using the following mass
balance equation:
24
23
22
21
21°C
20
STA-1W, Mar. 2006
ENP (Pa-hay-okee) DIP, Mar. 2006
ENP (Nine Mile Pond) DIP
T=19°C (Longinelli & Nutti, 1973)
T=21°C (Longinelli & Nuti, 1973)
T=23°C (Longinelli & Nuti, 1973)
19
18
-4
-3
-2
-1
Water
18
0
1
2
3
4
18
O (‰ vs. VSMOW)
Fig. 4. A comparison of d O values of DIP and water collected on April 13 (2005),
July 28 (2005), and March 7–9 (2006) from the Everglades area. Solid and dashed
lines represent the d18O relationships between biologically cycled DIP and ambient
water calculated using the equation given in Longinelli and Nuti (1973) for the
mean monthly temperature (solid line) and one standard deviation from the mean
monthly temperature (dashed lines) for March–April 2005 (a), July 2005 (b), and
February–March 2006 (c). The shaded area represents one standard deviation from
the mean d18O of fertilizer samples from EAA.
694
X. Li et al. / Applied Geochemistry 26 (2011) 688–695
biological activities, in controlling the P cycle in the Everglades
wetland ecosystems.
120
STA-1W, i n l e t
STA-1W, interior,
STA-1W, outlet,
ENP (Pa-hay-okee)
ENP (Nine Mile Pond)
Fertilizer DIP (%)
100
5. Conclusions
80
60
40
20
0
Apr 05
Jun 05
Aug 05
Oct 05
Dec 05
Feb 06
Apr 06
Sampling Date
Fig. 5. Relative contribution of fertilizer to the DIP pool, estimated using a two-end
member mass balance model. Error bars represent one standard deviation from the
mean value.
Widespread use of fertilizers has overloaded many aquatic
ecosystems with excess amounts of P and has caused a marked
increase in eutrophication (Carpenter et al., 1998). In order to
develop effective management practices to preserve water quality
and remediation plans for sites that are already polluted, it is
important to identify P sources and to understand the processes
affecting P cycling in aquatic systems. Measurements of the O isotopic compositions of DIP from both impacted and reference areas
of the Everglades wetland ecosystems show that the d18O value of
DIP varied both spatially and temporally, reflecting the water
temperature, the d18O value of water, rate of P cycling, and the
source/load of P. The d18O data suggest that biological cycling of
P in the polluted area was not rapid enough to completely
remove the fertilizer d18O signature and that the DIP pool in these
areas consisted of biologically cycled P as well as fertilizer P, with
fertilizer P accounting for about 15–100% of the total DIP. In contrast, the d18O values of DIP collected from reference sites or relatively pristine areas of the Everglades were close to or equal to
the predicted equilibrium d18O values, indicating a faster cycling
of P in the areas with very low P concentrations. The data presented in this study show that biological (bacterial activity) and
hydrologic factors (e.g., runoff, water residence time) exert an
important control on the d18O values of DIP. This study demonstrated that the O isotopic signature of PO3
is a useful tool for
4
studying the source and the degree of microbial cycling of P in
freshwater ecosystems.
Acknowledgements
Fig. 6. Comparison of d18O of DIP with monthly flow (acre feet, 1 acre
foot = 1233.5 m3), total P (TP) load (tons) and flow weighted mean concentration
(FWMC) (lg/L) in the STA-1W (from the USGS database, http://sofia.usgs.gov/
metadata/). All data except d18O are plotted on a natural logarithmic scale (vertical
axis) for easy comparison.
STA-1W, but less than 50% of the DIP pool at the outlet of the
STA-1W during the study period (Fig. 5). In contrast, the DIP-d18O
values in reference sites in the ENP were in or close to O isotopic
equilibrium with local water, suggesting that the DIP in relatively
pristine wetlands with very low P concentrations was dominated
by biologically cycled P and little or no fertilizer P. The data also show
that a high proportion of fertilizer-derived DIP generally corresponded to a high P concentration (Figs. 3 and 5).
Comparisons of the d18O values of DIP with hydrological data
(available from the USGS database: http://sofia.usgs.gov/metadata/) do not show a strong relationship between DIP-d18O and either
total P (TP) load or flow rate (p-value > 0.05) (Fig. 6), but a slightly
positive correlation between DIP-d18O and TP was observed in the
samples collected in the same season from STA-1W. Correlation
analyses also show that the d18O value of DIP is positively correlated with the flow weighted mean TP concentration (FWMC) (pvalue = 0.04) within STA-1W (Fig. 6). Such correlations would be
expected if runoff was a significant source of DIP and if the DIP
from the runoff had not been completely processed by microbes
within STA-1W. Although correlation does not mean causation,
these relationships observed between d18O values of DIP, TP and
FWMC underscore the importance of hydrology, in addition to
We would like to thank Dr. Adina Paytan and Karen McLaughlin
for allowing Xin Li to work in their lab to learn their phosphorus
extraction procedure and for their helpful suggestions. We are
grateful to two anonymous reviewers for their valuable suggestions and comments regarding the revision of the manuscript. This
work received financial support from South Florida Water Management District (Contract Number PCP-502588) and from National
Science Foundation (EAR-0073851).
References
Benitez-Nelson, C., Buesseler, K., 1999. Variability of inorganic and organic
phosphorus turnover rates in the coastal ocean. Nature 398, 502–505.
Bieleski, R., 1973. Phosphate pools, phosphate transport, and phosphate availability.
Ann. Rev. Plant Physiol. 24, 225–252.
Blake, R.E., Oneil, J.R., Garcia, G.A., 1997. Oxygen isotope systematics of biologically
mediated
reactions
of
phosphate.
1.
Microbial
degradation
of
organophosphorus compounds. Geochim. Cosmochim. Acta 61, 4411–4422.
18
Blake, R.E., O’Neil, J.R., Garcia, G.A., 1998. Effects of microbial activity on the d Op of
dissolved inorganic phosphate and textural features of synthetic apatite. Am.
Mineral. 83, 1516–1531.
Blake, R.E., O’Neil, J.R., Surkov, A., 2005. Biogeochemical cycling of phosphorus:
insights from oxygen isotope effects of phosphoenzymes. Am. J. Sci. 305, 596–
620.
Boyer, P., 1978. Isotope exchange probes and enzyme mechanisms. Acc. Chem. Res.
11, 218–224.
Caraco, N.F., Cole, J.J., Likens, G.E., 1992. New and recycled primary production in an
oligotrophic lake: insights for summer phosphorus dynamics. Limnol.
Oceanogr. 37, 590–602.
Carpenter, S.R., Caraco, N.F., Correll, D.L., Howarth, R.W., Sharpley, A.N., Smith, V.H.,
1998. Nonpoint pollution of surface waters with phosphorus and nitrogen. Ecol.
Appl. 8, 559–568.
Chiang, C., Craft, C.B., Rogers, D.W., Richardson, C.J., 2000. Effects of 4 years of
nitrogen and phosphorus additions on Everglades plant communities. Aquat.
Bot. 68, 61–78.
Colman, A., 2002. The Oxygen Isotope Composition of Dissolved Inorganic
Phosphate and the Marine Phosphorus Cycle. Dissertation, Yale Univ.
X. Li et al. / Applied Geochemistry 26 (2011) 688–695
Colman, A., Karl, D., Fogel, M., Blake, R., 2000. A new technique for measurement of
phosphate oxygen isotopes of dissolved inorganic phosphate in natural waters.
EOS, Trans. (Abstr.) 81 (48), 176 (Fall Meeting of Am. Geophys. Union, San
Francisco).
Colman, A., Blake, R., Karl, D., Fogel, M., Turekian, K., 2005. Marine phosphate
oxygen isotopes and organic matter remineralization in the oceans. Proc. Nat.
Acad. Sci. USA 102, 13023–13028.
Craft, C.B., Richardson, C.J., 1993. Peat accretion and N, P, and organic C
accumulation in nutrient-enriched and unenriched everglades peatlands. Ecol.
Appl. 3, 446–458.
Craft, C.B., Richardson, C.J., 1998. Recent and long-term organic soil accretion and
nutrient accumulation in the everglades. Soil Sci. Soc. Am. J. 62, 834–843.
Craft, C.B., Vymazal, J., Richardson, C.J., 1995. Response of everglades plant
communities to nitrogen and phosphorus additions. Wetlands 15, 258–271.
Davis, S.M., 1991. Growth, decomposition and nutrient retention of Cladium
jamaicense Crantz and Typha domingensis Pers. in the Everglades. Aquat. Bot.
40, 203–224.
Davis, S.M., 1994. Phosphorus inputs and vegetation sensitivity in the Everglades.
In: Davis, S.M., Ogden, J.C. (Eds.), Everglades: The Ecosystem and its Restoration.
St. Lucie Press, pp. 357–378.
Jaisi, D., Blake, R., 2010. Tracing sources and cycling of phosphorus in Peru Margin
sediments using oxygen isotopes in authigenic and detrital phosphates.
Geochim. Cosmochim. Acta 74, 3199–3212.
Jickells, T.D., 1998. Nutrient biogeochemistry of the coastal zone. Science 281, 217–
221.
Karl, D.M., Tien, G., 1992. MAGIC: a sensitive and precise method for measuring
dissolved phosphorus in aquatic environments. Limnol. Oceanogr. 37, 105–116.
Koch-Rose, M.S., Reddy, K.R., Chanton, J.P., 1994. Factors controlling seasonal
nutrient profiles in a subtropical Peatland of the Florida Everglades. J. Environ.
Qual. 23, 526–533.
Kolodny, Y., Luz, B., Navon, O., 1983. Oxygen isotope variations in phosphate of
biogenic apatite, I. Fish bone apatite – rechecking the rules of the game. Earth
Planet. Sci. Lett. 64, 398–404.
Lecuyer, C., Grandjean, P., Sheppard, S., 1999. Oxygen isotope exchange between
dissolved phosphate and water at temperatures &lt; 135 °C: inorganic versus
biological fractionations. Geochim. Cosmochim. Acta 63, 855–862.
Li, X., 2009. Tracing the Flow of Phosphorus, Carbon and Nitrogen in Aquatic
Ecosystems. Dissertation, Florida State Univ.
Longinelli, A., Nuti, S., 1968. Oxygen isotopic composition of phosphorites from
marine formations. Earth Planet. Sci. Lett. 5, 13–17.
Longinelli, A., Nuti, S., 1973. Revised phosphate–water isotopic temperature scale.
Earth Planet. Sci. Lett. 19, 373–376.
695
McLaughlin, K., Silva, S., Kendall, C., Stuart-Williams, H., Paytan, A., 2004. A precise
method for the analysis of d18O of dissolved inorganic phosphate in seawater.
Limnol. Oceanogr.-Methods 2, 202–212.
McLaughlin, K., Cade-Menun, B.J., Paytan, A., 2006a. The oxygen isotopic
composition of phosphate in Elkhorn Slough, California: a tracer for
phosphate sources. Estuar. Coast. Shelf Sci. 70, 499–506.
McLaughlin, K., Kendall, C., Silva, S.R., Young, M., Paytan, A., 2006b. Phosphate
oxygen isotope ratios as a tracer for sources and cycling of phosphate in North
San Francisco Bay, California. J. Geophys. Res.-Biogeosci. 111, G03003.
O’Neil, J., Roe, L., Reindhard, E., Blake, R., 1994. A rapid and precise method of
oxygen isotope analysis of biogenic phosphate. Isr. J. Earth Sci. 43, 203–212.
Paytan, A., Kolodny, Y., Neori, A., Luz, B., 2002. Rapid biologically mediated oxygen
isotope exchange between water and phosphate. Global Biogeochem. Cycles 16,
1–8.
Poister, D., Armstrong, D.E., Hurley, J.P., 1994. A 6-yr record of nutrient element
sedimentation and recycling in three North Temperate Lakes. Can. J. Fish. Aquat.
Sci. 51, 2457–2466.
Reddy, K.R., White, J.R., Wright, A., Chua, T., 1999. Influence of phosphorus loading
on microbial processes in the soil and water column of wetlands. In: Reddy, K.R.,
O’Connor, G.A., Schelske, C.L. (Eds.), Phosphorus Biogeochemistry in Subtropical
Ecosystems. Lewis Publishers, Boca Raton, pp. 249–273.
Richardson, C.J., 1999. The roles of wetlands in storage, release and cycling of
phosphorus on the landscape: a 25 year retrospective. In: Reddy, K.R., O’Connor,
G.A., Schelske, C.L. (Eds.), Phosphorus Biogeochemistry in Subtropical
Ecosystems. Lewis Publishers, Boca Raton, pp. 47–68.
Shemesh, A., Kolodny, Y., Luz, B., 1983. Oxygen isotope variations in phosphate of
biogenic apatite, II. Phosphate rocks. Earth Planet. Lett. 64, 405–416.
Shemesh, A., Kolodny, Y., Luz, B., 1988. Isotope geochemistry of oxygen and carbon
in phosphate and carbonate of phosphorite francolite. Geochim. Cosmochim.
Acta 52, 2565–2572.
Sklar, F.H., Chimney, M.J., Newman, S., McCormick, P., Gawlik, D., Miao, S.L., McVoy,
C., Said, W., Newman, J., Coronado, C., 2005. The ecological–societal
underpinnings of Everglades restoration. Front. Ecol. Environ. 3, 161–169.
Stern, J., 2005. Biogeochemical Cycling of Carbon, Phosphorus, and Trace Metals.
Dissertation, Florida State Univ.
Stern, J.C., Wang, Y., 2002. Using oxygen isotopes to determine the source of
phosphate in the Everglades. Eos Trans. AGU, Spring Meeting Suppl. 83, 19.
Stern, J., Wang, Y., Gu, B., Newman, J., 2007. Distribution and turnover of carbon in
natural and constructed wetlands in the Florida Everglades. Appl. Geochem. 22,
1936–1948.
Wang, Y., Hsieh, Y.P., Landing, W.M., Choi, Y.H., Salters, V., Campbell, D., 2002.
Chemical and carbon isotopic evidence for the source and fate of dissolved
organic matter in the northern Everglades. Biogeochemistry 61, 269–289.