Phosphorus Flux from the Pelham Bay Wetlands, Bronx, New York


Phosphorus (P) released and transported from wetlands affects wetland ecosystems and surrounding water systems. Sediment samples from the Pelham Bay wetlands in the Bronx, NY, have been collected and analyzed. Analysis of the sorption characteristics, P compound identification and P mineralization, showed significant correlations between the sorption maximum (Smax) and Ox-Fe (r = 0.894), ash-TP and HCl-Ca (r = 0.94), ash-TP and TOC (r = 0.96), as well as TOC vs HCl-Ca (r = 0.93). These results indicate that mineral content affects the OP content and the sorption process. P sorption maxima Smax ranged from 70.4 to 1667 mg/Kg, and the equilibrium P concentration EPC0 ranged from 0.09 to 0.2 mg/L. The high Smax in most of the sites, fairly high EPC0 and high percentages of Pr (>96%), indicate substantial amounts of P could be bioavailable for plant uptake in the water column under changing hydro-climatic conditions. Dominant P compounds are dihydroxyacetone phosphate (DHAP), phosphoenolpyruvates (PEP), inosine monophosphate (IMP), glucose-6-phosphate (G6P), nucleoside monophosphates (NMP), glycerophosphate (GlyP), polynucleotides (PolyN), and pyrophosphates (Pyrop). The active P pool could maintain substantial P bioavailability and potentially cause eutrophication. The mineralization of 7, 15 and 30 days in laboratory experiments indicates a decrease of NaHCO3 and HCl, and an increase of NaOH.

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Wang, J. (2022) Phosphorus Flux from the Pelham Bay Wetlands, Bronx, New York. Journal of Environmental Protection, 13, 360-383. doi: 10.4236/jep.2022.135023.

1. Introduction

Wetland sediments, soils, and pore waters are sources of Phosphorus (P) and are a factor influencing the eutrophication of water bodies [1] (Massen and Balla, 2010). Wetlands help improve water quality, and artificial wetlands have been built to treat wastewater and seepage, mill waste and animal waste [2] (Nairn and Mitsch, 2000). Goose Creek, the Hutchinson River and the Lagoon feed the three major marshes in the Pelham Bay Park, which has the largest assemblage of tidal wetlands in New York City, outside of Jamaica Bay. Salt marshes trap pollutants/contaminants, acting as a natural filtration system to reduce contamination in bays and oceans. Wetlands have the ability to absorb fertilizer, reduce erosion, and improve water quality [3] (NYC Parks). P dynamics in wetlands are influenced by sediment physicochemical characteristics, and P release from wetland sediment is critical for wetland restoration; P management in wetlands is important to reduce nutrient loading to downstream water bodies [4] (Dunne et al., 2010). Hydrological factors affect water quality in wetlands, and riparian wetlands help improve the water quality in urban watersheds [2] (Nairn and Mitsch, 2000). Wetland is a unique ecosystem providing water resources and protecting biodiversity. Wetland type plays an important role on the P biogeochemical cycle [5] (Hu et al., 2022).

P is an essential nutrient for plants and animals; however, excessive P can be released into water bodies and wetlands from anthropogenic sources, e.g. storm water, fertilizer application, and sewage overflow. These anthropogenic factors cause eutrophication from algae blooming and oxygen depletion [6] (Adam et al., 2007), and degraded water quality in fresh water systems such as rivers, wetlands, and saline waters system in estuary areas. P can be sorbed or precipitated by Fe, Al, Ca, Mg; and these metal ions play important roles on P absorption and desorption. In the Pelham Bay wetlands, Ca is especially dominated by P concentrations and equilibrium phosphorus concentration-EPC0. The sediments from the freshwater wetlands contain Ca that precipitates P in a stable form, apatite. Sediment size also influences the sorption process, because fine-grained sediments, including silt and clay, have a large surface area [6] (Adam et al., 2007).

P can be released from sediment/soil pore water into the water column and incoming water bodies in the wetland that contain lower concentrations of P [4] (Dunne et al., 2010). Organic P accumulates in wetlands, and it can be released as inorganic phosphate by microbial decomposition [7] (Newman and Robinson, 1997; [8] Diaz et al., 2006). Physicochemical factors, such as pH and temperature, as well as other hydrological conditions, influence P release and organic matter decomposition [9] (Wright and Reddy, 2001). The internal supply of P in wetlands causes P release from soil/sediment to overlying water. Sediment/soil physicochemical conditions, such as pH, redox, organic matter, nutrient content, calcium/magnesium content in alkaline soil, and iron/aluminum content in acidic soil, as well as overlying floodwater temperature, oxygen content/ionic strength can affect P release [4] (Dunne et al., 2010).

Armorphous acid oxalate-extractable aluminum and iron content in the soil/sediments are naturally associated with P retention capacity in wetlands, bogs, swamps, and rivers. The extractable aluminum content of the soil predicts the P adsorption potential in wetland ecosystems [10] (Richardson, 1985). Microbial and plant uptake, adsorption and incorporation of organic P (OP) into soil peat can remove dissolved inorganic P (DIP) from the water, consequently determining the P storage in wetlands. P is released after tissue death of the plant [10] (Richardson, 1985).

Anthropogenic sources impact P loading in a nutrient-limited ecosystem (i.e., northern Everglades) [7] (Newman et al., 1997). Reducing P loading is a strategy to help environmental restoration [11] (Daroub et al., 2003). Sediment P plays an important role in eutrophication in shallower water bodies [12] (Bostrom et al., 1982). The physiochemical properties of sediments impact P internal cycling, and P released from sediments has a significant impact on P concentrations in the water column [12] (Bostrom et al., 1982). Sediment could be either sink or source for P depending on the water column and sediment physicochemical properties [13] (Reddy et al., 1995; [10] Richardson, 1985). Analyzing TP concentrations and major P compounds in the sediments is an approach used to estimate the potential impact of sediments on shallow water bodies such as wetlands, streams, and lakes [8] (Diaz et al., 2006).

Organic P can comprise more than 50% of sediment TP in wetland ecosystems and could be as much as 90% of the water column TP. The accumulation and mineralization of OP is important in P cycle [14] (Reddy et al., 1998; [15] Rigler, 1964, [7] Newman and Robinson, 1997). OP, typically identified as the difference between TP and IP, could be either a sink or a source for bioavailable P. The mineralization of labile OP is critical for internal P cycling and productivity of an ecosystem [7] (Newman and Robinson, 1997). Sequential extraction is used to identify OP [16] (Hedley et al., 1982). Chromatographic fractionation, 31P nuclear magnetic resonance (NMR), and enzymatic hydrolysis are techniques used to isolate and identify specific OP compounds [7] (Newman and Robinson, 1997). There are six common forms of P using 31P-NMR: phosphonates, inorganic orthophosphate, orghophosphate monoesters, orthophosphate diesters, pyrophosphates, and polyphosphates [17] (Dai et al., 1996). The 31P-NMR extraction solution that has been commonly used is 0.5 M NaOH [18] (Bowman and Moir, 1993); some studies also used 0.25 M NaOH-0.05 M EDTA [19] (Cade-Menun and Preston, 1996).

The objective of this research is to estimate the nutrient impact on water quality in the wetlands in Pelham Bay Park, estimate internal P loading to the water column and sediments, identity P compounds and relative composition in wetland sediments, and estimate their potential bioavailability. This research sponsored by a grant from the New York Park Division was conducted to analyze the wetland sediments characteristics on sorption, in order to predict the potential bioavailability of phosphorus, the sources of phosphorus, the causes of algae blooms and the toxic algae’s deleterious effects on ecosystems and human health. The research also helps form an effective strategy for water quality improvement in Pelham Bay estuary area. An improvement in water quality will also benefit New Yorkers’ health.

2. Material and Methods

2.1. Study Area

Pelham Bay Park, the largest municipal park in the New York City, is located on Long Island Sound, in the northeastern corner of the Borough of the Bronx. It contains nearly 500 acres of tidal marshes [20] (NYS DOS, 1992), comprised of high marshes, intertidal marshes, salt flats, lagoons, a narrow bay, and a wetland complex. The area is managed by the City of New York Department of Parks and Recreation [20] (NYS DOS, 1992). Wetland and aquatic habitats in Pelham Bay Park have been seriously degraded by discharges of sewage, stormwater runoff, and industrial effluents containing heavy loads of nutrients, sediments, or chemical pollutants into contiguous water bodies; or water pollution caused by oil spills, waste disposal, and excessive turbidity causing by construction/dredging [20] (NYS DOS, 1992).

The Pelham Bay wetlands are located in Pelham Bay Park, Bronx, NY. The Pelham Bay Park Wetlands, one of the largest natural estuarine habitats on the north shore of western Long Island Sound, is the only undisturbed tidal wetlands areas remaining in the Bronx [20] (NY DOS, 1992). For this study we collected samples at six sampling sites within the wetlands area. The six fresh water wetland sites locations were (Figure 1; Table 1): sites 1, 2, 3, and 4, located in the

Figure 1. Pelham bay freshwater wetlands study area.

Table 1. Geographical locations of the sampling sites and sediment texture in Pelham Bay freshwater wetlands.

wetlands close to the Bartow-Pell Mansion Museum; site 5 in the wetland at the Pelham Golf Course; and site 6 located in a ditch in the Split Rock Golf Course (Figure 1). The major contents in the sediment are silt and clay. The sediment at sites 1, 2, and 3 were very sticky and silty; site 4 was more sandy, with detritus materials such as plant roots; sites 5 and 6 had a putrid smell, very sticky, composed mostly of silt and clay. The putrid odor came from fertilizer that had applied on the golf courses.

2.2. Sediment Sampling in Freshwater Wetlands

Sediment samples were collected from the six wetland sites in July 2006. Each site was located with a Global Positioning System (GPS) unit, and the coordinates are provided in Table 1. A Core Sampler (diameter = 8 cm; length = 17 cm) was used to obtain the bed sediments. The sediment samples were sealed in one gallon zipper bags, and transported to the Environmental Laboratory of the Department of Environmental, Geographic, and Geological Sciences, Lehman College of The City University of New York at the end of each sampling day, and stored at 4˚C until further experimentation. The sediment samples were immediately homogenized and saved in waterproof double-track zipper bags (10.2 × 15.2 cm; made by Fisher Scientific Co., USA), and stored at 4˚C until they were used for further analysis. A portion of each homogenized sediment sample was dried at 70˚C for 72 h, and then finely ground and used for selected physico-chemical analysis.

2.3. Physcio-Chemical Analysis of the Sediments

The wet sediment samples were oven dried at 105˚C for 24 h and then analyzed for their water content; sediment samples were also dried at 70˚C for 72 h and finely ground for total phosphorus (TP) analysis. To determine TP, 0.5 to 1.0 g of finely ground dry sediments were combusted at 550˚C in a muffle furnace for 5 h and the ash was dissolved in 6 M HCl. The digestate was filtered to a 100 ml volumetric flask with Fisher filter paper using an automated ascorbic acid method (ESS Method 310.1; USEPA, 1992) [21]. Meanwhile, the total organic matter (TOM) in sediments was determined by combustion in a muffle furnace. The total organic carbon (TOC) is equal to 0.58 times of the amount of TOM (TOC = 0.58*TOM).

To determine the different poorly crystalline Fe and Al effect on phosphorus sorption, 0.5 g of finely ground dry sediment was extracted with 30 ml 0.175 M ammonium oxalate + 0.1 M oxalic acid at a ratio of 1:60 (0.5 ml/30ml = 1:60) by shaking for 4 h on an end-over-end mechanical shaker [22] (Pant and Reddy, 2001). Then the suspensions were centrifuged at 10,000 rpm for 10 min and filtered through a 0.45 µm membrane filter in the dark. The filtrates were analyzed for oxalate-extractable Al and Fe, plus oxalate-extractable P by an inductively coupled argon plasma spectrometry (ICP).

To analyze Ca and Mg effects on P sorption, a similar procedure as the previous step was used where the finely ground dry sediment of 0.5 g was extracted with 30 ml 1 M HCl at a ratio of 1:60 (0.5 ml/30ml = 1:60) by shaking on an end-over-end mechanical shaker for 4 h [22] (Pant and Reddy, 2001). The suspensions were centrifuged at 10,000 rpm for 5 min and filtered through a 0.45 µm membrane filter in the dark. The filtrates were then analyzed for HCl-extractable Ca, Mg, and P by inductively coupled argon plasma spectrometry (ICP-MS).

2.4. Phosphorus Sorption Experiments

A batch incubation was conducted to determine P absorption and desorption properties. All of the sediments were considered under aerobic conditions. Five grams of wet sediments were placed in a 50 ml centrifuge tube and shaken for 24 h in a 10 ml solution with orthophosphate-potassium phosphate monobasic (KH2PO4) in 0.01 M (mol/L) potassium chloride (KCl) solution with concentrations of 0, 0.1, 0.2, 0.5, 1, 2, 5, 10, 25, 50, and 100 mg/L (ppm) P, using an end-over-end mechanical shaker at 25˚C ± 2˚C. This was the absorption procedure. Thereafter, the homogenized samples were centrifuged for 20 min at 20˚C, 10,000 rpm. Then the supernatants were decanted to 25 ml vials, acidified with one drop of concentrated sulfuric acid H2SO4 and stored at 4˚C for P absorption analysis. The sediment residues were then weighted and filled with 10 ml 0.01 M KCl solutions for desorption procedure, shaking for 24 h, using an end-over-end mechanical shaker at 25˚C ± 2˚C. Then the samples were centrifuged at 10,000 rpm for 20 min, decanted to 25 ml vials, and acidified with concentrated H2SO4. The concentration of P in the supernatants from both absorption and desorption were measured at 880 nm by UV-2501PC UV-VIS Recording Spectrophotometer using an automated ascorbic method (ESS Method 310.1; USEPA, 1992) [21].

2.5. 31P-NMR Identification

A wet sediment sample of 40 g was extracted twice with 80 ml 0.4 M NaOH, each for 4 h by shaking in an end-over-end shaker at 25˚C + 2˚C, 230 l/min. The suspensions were centrifuged at 21˚C + 2˚C, 10,000 ×g for 30 min after each extraction. The supernatants were combined and concentrated 10 times from 100 ml to 10 ml using a Vacuum Rotatory Evaporator (R-205, BUCHI) at 25˚C. Then 0.45 ml concentrated extracts were put in a 0.5 mm diameter Wilmad NMR tube, one drop of D2O was added to lock the signal, and then the solution was scanned at 121.4944448360 MHz on a BRUKER 300 UltraShield 31P NMR for 15 h, and the spectra were processed using TOPSIN 2.0 software. The solution composition was determined with an external standard of 85% phosphoric acid, and the NMR spectra peaks were identified by comparing the sample peaks with the differences from the reference peaks: 0.6 ml samples were spiked with 0.02 ml 10,000 mg·l1 sodium pyrophosphate (Na4HP2O7·10H2O) as an internal standard [23] (Gadian et al., 1979; [24] Pant et al., 1999).

2.6. Phosphorus Mineralization Experiments

Wet sediments (2 g) were incubated under a nonflooding condition at 37˚C for 7, 15, 30, and 45 days (d), and extracted with NaHCO3, NaOH, and HCl sequentially as mentioned previously. The changes in NaHCO3-P, NaOH-P, HCl-P, residue-P in the sediments during incubations at 37˚C for 7, 15, 30, 45 d from without incubation were to determine the extent of P mineralization at physiological/optimum temperature.

2.7. Statistical Analysis

All the P mineralization experiments were carried out in triplicate and analyzed with SAS JMP Version 13.0 [25] (SAS Institute Inc., 2008) using two-way analysis of variance (ANOVA). A Tukey’s HSD (honestly significant difference) test was used for statistical differences at p < 0.05 level as described by [26] Tao et al. (2008). Similarly, correlation coefficients were tested on sorption parameters using Pearson’s correlation (r) significant test at the p < 0.05 or p < 0.01 level using SPSS Version.15.0 [27] (SPSS Inc., 2006). All regression analyses were performed using the SPSS 15.0 software.

3. Results and Discussion

3.1. Physio-Chemical Characteristics of the Wetland Sediments

No significant difference in pH was found among the six freshwater wetland sites in the Pelham Bay (Table 2). EC ranges from 87.9 at site 2 to 540 µs/cm at site 6 Split Rock Golf Course. Total organic carbon (TOC) ranges from 3.4 to 9.2. Ash-TP ranges from 337.6 to 666.1 mg/Kg (Table 2). The highest TOC%, Ox-P, Ox-Al, HCl-Ca, HCl-P and ash-TP, and the second highest Ox-Fe, HCl-Mg appeared at Site 4 beside Turtle Creek close to the Bartao-Pell Mansion Museum. The four sites close to Bartao-Pell Mansion Museum (sites 1, 2, 3 and

Table 2. Pelham fresh water wetlands chemical characteristics.

OM: total organic carbon, %. ash-P: the total phosphorus in bed sediments from Anderson ignition method, mg·kg−1. EC: electrical conductivity, µs·cm−1.

4) have relatively high Ox-Fe, Ox-Al (other than site 2, with a value of that was lower than site 6, Ox-P, Fe:Al, TOC, HCl-Ca, and total ash-P, which might be related to the sticky silty and clay type of texture, and OM could be attached to ferric (Fe3+) in aerobic conditions.

3.2. Sorption

Smax ranges from 70.4 to 1666.7 mg/Kg (Table 3), with the highest value at sites 1-3. EPC0 (ranging from 0.086 to 0.235), were also high indicating a substantial amount of P that remained in the sediments could be bioavailable to the water column under seasonal conditions (Table 3). Pearson’s correlation results show that Smax (P sorption maximum) is strongly correlated to Ox-Fe (after removing outlier Smax at site 4) (r = 0.89, significant probability level p < 0.05) (Table 4). The maximum Smax at site 2 was 1666.7 mg/Kg (Figure 2 and Table 3), and the Ox-Fe at site 2 was 19,166 mg/Kg (Table 2), both were highest among the six sites. The strong correlation between Smax and Ox-Fe (r = 0.894) indicates that the sediments that had high Fe content tended to have high maximum sorption capacity (Table 4). Site 1 and site 2 were mostly controlled by fine clay and silt; the soil texture was sticky for both sites. Redox has an impact on sorption. A high content of amorphous and poorly crystalline Fe under aerobic conditions results in high Smax values [22] (Pant and Reddy, 2001; [28] Wang and Pant, 2010a). The wetland area was exposed to air and ferric iron as Fe II converted to ferrous Fe as FeIII. The FeIII in both sties bound P tightly and had a high capacity to bind P [28] (Wang and Pant 2010a). EPC0 at site 2 was 0.086 mg/L, which was the lowest among the six sites (Figure 3 and Table 3), indicating that EPC0

Table 3. Phosphorus sorption characteristics of the Pelham fresh water wetlands.

Kd: distribution coefficient, L·kg1. S0: P originally sorbed on the solid phase, mg·kg1. EPC0: equilibrium P concentration, mg·L1. Smax: P sorption maximum, mg·kg1. k: a constant related to the bonding strength, L·mg1. Kf: P absorption energy, L·kg1.

Table 4. P sorption parameters and selected chemical characteristics coefficient r-values.

*Correlation is significant at the 0.05 level (2-tailed), p < 0.05. **Correlation is significant at the 0.01 level, p ≤ 0.01.

was not impacted by Ox-Fe [28] (Wang and Pant, 2010a). The higher EPC0 is mostly associated with anaerobic conditions, and ferrous FeIII is reduced to ferric Fe II [22] (Pant and Reddy, 2001). Sediment would absorb P if the P concentration in the water column was more than EPC0 in the wetland sediment; sediment would desorb P if the P concentration in the water column was less than sediment’s EPC0 ( [29] Kunishi et al., 1972; [30] Wang et al., 2009; [28] Wang and Pant, 2010a). S0 (P originally sorbed on the solid phase) is strongly correlated to Kd (adsorption energy), and the Pearson correlation coefficient r was 0.936 (Table 5). Pr% at the six sites ranges from 96% to 100% (Table 3), indicating that most adsorbed P was retained at a higher level than desorbed P [28] (Wang and Pant, 2010a). EPC0 is not significantly correlated to Pr% ( Pr % =

Pr Pad × 100 % ), however, site 2 has the lowest EPC0 but the highest Pr% and

Table 5. Pearson’s correlation matrix (r) of sorption parameters.

*Correlation is significant at the 0.05 level (2-tailed), p < 0.05. **Correlation is significant at the 0.01 level, p ≤ 0.01.

Figure 2. Pelham wetland 2 (close to Bartao-Pell mansion museum) Smax calculation.

Figure 3. Pelham wetland 2 (close to Bartao-Pell Mansion Museum). The langmuir isotherms batch experiment for EPC0 and S0 calculation.

Smax. The high retention ratio in the sediment made it saturated with P, therefore the equilibrium P concentration decreased when Smax and Pr% increased. Changes of water chemistry in sediments, such as pH and redox, could potentially cause massive P release to the water column ( [30] Wang and Pant, 2009; [28] 2010; [31] 2010; [32] 2011; [33] 2011; [34] 2012).

The significant correlations at the 0.01 level (2-tailed) between TOC (total organic carbon) and HCl-Ca, HCl-Ca vs ash-TP and TOC vs ash-TP (Table 4) indicate that Ca and OM bounded P could become a potential source of P in the water column under different redox conditions [8] (Diaz et al., 2006). HCl-Pi (HCl extracted IP) could range from 46% - 64% of TP in water conservation area canals in the Everglades [8] (Diaz et al., 2006). HCl extracted SRP (comparatively equal to HCl-Pi) ranged from 12% to 60%, and concentrations ranged from 73.9 to 255.9 mg/Kg in Pelham Bay wetland sediments. HCl-Pi, Ca, and Mg bound P is relatively stable and unavailable for biological assimilation [35] (Sonzogni et al., 1982). Organic P extracted by NaOH can be hydrolyzed to bioavailable P, which is associated with humic and fulvic acid ( [36] Bowman and Cole, 1978; [37] Ivanoff et al., 1998).

3.3. P Compound Identification-31P-NMR

Organic P compounds were found in the 6 sampling sites, which could be related to the OM attached to Fe3+ [30] (Wang and Pant, 2009). The dominant P compounds G6P and DHAP are organic P compounds in sites 1, 2, 3, and 4, and those four sites also have the highest Ox-Fe, Ox-Al, Ox-P, TOC, HCl-Ca, HCl-P and ashP, indicating organic P is dominant in those fine sediments [30] (Wang and Pant, 2009). G6P (glucose-6-phosphate, major peak at 5.34 ppm, 81.5%), NMP (nucleoside monophosphates, peak at 4.42 ppm, 7.2%), IMP (inosine monophosphate, peak at −1.11, trace amount of 3.8%) and trace amount of PyroP (pyrophosphate) were found in site 1 (Figure 4 & Figure 5). The percentage proportion was determined by the peak area %. The major peak at site 1 was G6P (Table 6), and it had the highest TOC among the six sites. G6P appeared in all sites. DHAP (dihydroxyacetone phosphate) was the dominant compound (peak at 5.59 ppm, nearly 100%), and trace amounts of PEP (phosphoenopyruvates, peak at 0.11, 0.1%) and IMP (peak at −0.23, 0.1%) at site 2 (Figure 6 & Figure 7) were also present. There is around 100% DHAP (99.97%) at site 3 (peak at 5.16 ppm), plus trace amounts of orthophosphate (peak at 8.01 ppm, 0.01%)), GPEA (glycerophosphoethanolamine, peak at 0.47 ppm, 0.01%) and PEP (phosphoenolpyruvates, peak at −0.10 ppm, 0.01%) (Figure 8 & Figure 9). Major P compound at site 4 was G6P (4.47 ppm, 97.40%) (Figure 10 & Figure 11), trace peaks are NMP (3.57 and 3.23 ppm, 2.4%), IMP (−1.86 ppm, 0.1%), and PyroP (−5.99 ppm, 0.1%). G6P is about 100% at site 5 (Figure 12 & Figure 13). The spectra at site 6 showed an interesting pattern, with diverse P compounds (Figure 14 & Figure 15): trace amount DHAP (peak at 6.49 ppm, 0.1%), PolyN (peak at 3.65 ppm, 0.1%), and major P compounds are GlyP (peak at 5.19 ppm, 26.3%), NMP (peak at 4.25 ppm, 26.3%) and PyroP (peak at −5.14, 25.6% and peak at −6.07, 21.5%).

Compared to the Bronx River, P compounds in the Pelham Bay freshwater wetlands are quite different. The P compounds in the river bed sediments are mostly GlyP, NMP and PolyN, few DHAP and IMP. Major OP compounds in wetlands are G6P, DHAP, and small amounts of PEP [30] (Wang and Pant,

Figure 4. PW 1 concentrated sample. Site 1 PW1-Pelham Bay wetland close to Bartow-Pell Mansion Museum. There were three peaks in this sites concentrated sample, the major peak at 5.34 was G6P, the one at 4.42 was NMP, and the one at −1.11 was IMP. They represented G6P-81.5%, NMP-7.2%, and IMP-3.8% of relative composition respectively.

Figure 5. PW1 spiked sample, Pyrop spike: −5.1697.

Figure 6. PW2 concentrated sample. Site 2 PW2-Pelham Bay wetland close to Bartow-Pell Mansion Museum. PW2 had DHAP for 99.8%, PEP 0.1%, and IMP 0.1% showing in Figure 6.

Figure 7. PW 2 spiked sample, pyrop spike: −5.6057.

Figure 8. PW3 concentrated sample. Site 3. PW3-Pelham Bay wetland close to Bartow-Pell Mansion Museum. The major peak at 5.16 ppm was DHAP 99.97%. The rest of three peaks have not been identified yet, peak at 8.01 ppm orthophosphate (0.01%), 0.47 ppm GPEA (0.01%) and −1.02 ppm PEP (0.01%).

Figure 9. PW3 spiked sample, pyrop spike: −6.3514.

Figure 10. PW4 concentrated sample. Site 4. PW4-Pelham bay wetland close to Bartow-Pell mansion museum. W4 had five different P compounds: peak at 4.47 ppm G6P of 97.4%, peaks of 3.57 and 3.23 were both NMP of 1.2% each, −1.86 ppm was IMP of 0.1%, and −5.99 ppm was PyroP of 0.1%.

Figure 11. PW4 spiked sample, pyrop spike: −5.9869.

Figure 12. Site 5 concentrated sample. Site 5. Pelham golf course. Site 5 had one major peak at 6.67 ppm, which was G6P, accounted for 100% of relative composition.

Figure 13. Site 5 spiked sample, pyrop spike: −4.1397.

Figure 14. Site 6 concentrated sample. Site 6. Split rock golf course. Site 6 spiked sample showed a very interesting pattern, the first double peaks include peak at 5.19 ppm of GlyP for 26.3% and peak at 4.25 ppm of NMP for 26.3%. Peak at further left 6.49 ppm of DHAP for 0.1%, and further right 3.65 ppm was PolyN for 0.1%; the second double peaks were composed of the Pyrop spike at −5.15 ppm, around 25.6%, and another Pyrop from the sample at −6.07 ppm of 21.5%.

Figure 15. Site 6 spiked sample. DHAP 6.49 ppm: 0.1%, PolyN 3.65 ppm: 0.1%, GlyP 5.18 ppm: 26.3%, NMP 4.25 ppm: 26.3%, PyroP −5.14 ppm: 25.6%, PyroP −6.07 ppm: 21.5%.

Table 6. NMR results.

G6P: glucose-6-phosphate; NMP: nucleoside monophosphates; IMP: inosine monophosphate; Pyrop: pyrophosphate; DHAP: dihydroxyacetone phosphate; PEP: phosphoenopyruvate; GPEA: glycerophosphoethanolamine; GlyP: glycerophophate.

2009), IMP, besides GlyP, PolyN and NMP. Here, G6P might not exist in the water system because it was soluble, and it only could exist in wetland or soil environment. There are many other organic P compounds in wetlands rather than in the river. The P bioavaiability prediction in wetlands is quite meaningful; because under certain conditions, such as favorable pH, temperature, redox, the organic P (OP) could be mineralized to inorganic P (IP) and become bioavailable to plants [34] (Wang and Pant, 2012) and potentially cause eutrophication. The active P pool in the wetland sediments could maintain substantial bioavailability of P for seasonal algal growth under changing hydro-climatic conditions.

3.4. P Mineralization

The phosphorus pool changed during the mineralization for 7, 15, 30, and 45 days under flooding incubation. NaHCO3 increased from 43 to 57 mg/Kg during the 7, 15, 30, and 45 day flooding incubation (Table 7, Table 8). The high content of NaOH-P in sites 1 - 4 is associated with the high organic matter in those sites, correspondingly dominant OP compounds, and those sticky, silt, and clay, and sandy sediments that are associated with Fe, Al, and under changing redox conditions, could have resulted in high Smax or high EPC0 ( [30] Wang and Pant, 2009; [32] 2011).

3.4.1. Spatial Variation of 7, 15, 30, 45 Day Mineralization

The 7-day nonflooding mineralization after sequential extraction by NaHCO3, NaOH and HCl showed that site 3 had the highest average value of 66.26 mg/Kg (Table 6) for NaCO3-P, and it was significantly higher than sites 1, 5, and 6. Site 5 had a significantly lower NaOH-P value than the other sites. HCl-P was the highest on site 6 and it was significantly higher than sites 1, 3 and 5. Total cumulative P was highest at site 4 (542 mg/Kg), and that is significantly different from site 5. The ashTP was significantly different among the six sites, and site 4 had the highest one of 875 mg/Kg that is significantly higher than the other five sites. Site 5 had the lowest of 337 mg/Kg (Table 6). Residue P had the highest value at site 4 of 333 mg/Kg, which was significantly higher than sites 2, 3, 5 and 6. The negative values at sites 3 and 6 were because ashTP was lower than total cumulative P. Combined sewer overflows (CSOs) in the discharge of untreated sewage into NYC rivers during rainstorms, elevate nutrients and bacteria levels and contaminate the waterways, which could impact P composition, P concentration, P pool, and P mineralization [38] (Wang, 2014) differently from river bed sediments.

The 15 day nonflooding incubation showed variation among the six sites. Sites 1-4 are significantly higher than sites 5 and 6 for NaHCO3-P (Table 6). Site 2 had the highest NaOH-P of 378 mg/Kg that is significantly higher than other five sites. Site 6 again had the highest HCl-P of 220 mg/Kg, and it was significantly higher than other sites (similar as 7 day incubation). Cumulative P appeared highest at site 2, but ashTP was the highest at site 4, which had the highest residue P of 412 mg/Kg.

The 30 day nonflooding incubation showed similar patterns as 7 and 15 days. Site 4 had all the highest values other than HCl-P (Table 6). The 45-day flooding incubation also showed similar patterns. The highest values of NaHCO3-P, TP-added up, ashTP, and residue occur at site 4, whereas the highest NaOH-P was at site 2 (344 mg/Kg, which was higher than the 7, 15, and 30 day flooding incubation). HCl-P of 206 mg/Kg was highest at site 6, which was lower than the 7, 15, and 30 day flooding incubation.

Table 7. Mineralization of 7, 15, 30 and 45 days.

Values in the same column with different letters are significantly different (Tukey’s HSD test, p ≤ 0.5).

Table 8. All sites 7, 15, 30 and 45 days.

Values in the same column with the same letters are not significantly different (Tukey’s HSD test, p ≤ 0.5).

3.4.2. Temporal Variation of 7, 15, 30, 45 Day Incubation

No significant variation exists among 7, 15, 30 and 45 day incubation (Table 7) even though thought there is a slight difference in concentrations of NaHCO3-P, NaOH-P, HCl-P, total cumulative P, ash TP, and residue P. It showed a different pattern of mineralization from river bed sediments in the Bronx River [31] (Wang and Pant, 2010). For all of the six sites, NaHCO3-P increased from day 7 to day 15, 30, and 45. For site 1, NaOH-P increased from day 7 to day 15, then decreased on day 30 and increased again on day 45; all other sites had similar patterns other than site 5 at the Pelham golf course and site 6 at the Split Rock Golf Course. The NaOH-P indicates there is organic matter in the sediments, and high Al and Fe content in humus matter [32] (Wang and Pant, 2011). For HCl-P no significant variation was found at the different sites. Site 1 increased from day 7 to day 15, decreased on day 30, and increased on day 45, and a similar temporal pattern occurred at other sites. Residue-P increased on day 30 for most of the sites.

There is no significant difference for all sites on flooding incubation for 7, 15, 30 and 45 days on NaHCO3-P, NaOH-P, HCl-P, cumulative TP, ashTP, and residue P (Table 8).

4. Conclusion

Phosphorus flux in the Pelham Bay wetlands showed a significant sorption capacity in wetland sediments, showing potential bioavailability and nutrient enrichment. Sites 1, 2, and 3 had significantly higher values of Smax, Ox-Fe, and Pr%, and there was a significant correlation between Smax and Ox-Fe that indicated that ferrous Fe sediments in the wetlands have a large capacity for sorption of P. G6P and DHAP are the dominant P compounds at sites 1, 2, and 3, and NaOH-P values are highest in these three sites on 7 day incubation, which showed that organic P played a significant role in the sediment in these sites. Site 4 had the highest values of EPC0, Ox-Al, Ox-Fe, Ox-P, TOC, HCl-Ca, HCl-Mg, HCl-P, and ashP. The dominant P compound is G6P, indicating organic matter is highest at this site, and organic matter associated with amorphous and poorly crystalline Fe and Al formed suitable conditions for P sorption. The OM and Ca bounded P could release bioavailable P under different redox conditions. However, the high EPC0 produced the lower Smax value. The changing hydro-climatic conditions such as temperature rise could potentially release more bioavailable P in the wetlands and water system and cause eutrophication that degrades water quality. Organic matter bounded tightly with fine and sticky wetland sediments could have a higher potential to release bioavailable P under different redox conditions. P composition in these wetland sites are dominated by OP and P compounds that could be distinguished from the river bed sediments.

Conflicts of Interest

The author declares no conflicts of interest regarding the publication of this paper.


[1] Massen, S. and Balla, D. (2010) Impact of Hydrodynamics (Ex- and Infiltration) on the Microbially Controlled Phosphorus Mobility in Running Water Sediments of a Cultivated Northeast German Wetland. Ecological Engineering, 36, 1146-1155.
[2] Nairn, R.W. and Mitsch, W.J. (2000) Phosphorus Removal in Created Wetland Ponds Receiving River Overflow. Ecological Engineering, 14, 107-126.
[3] NYC Parks (2008) Pelham Bay Park.
[4] Dunne, E.J., Clark, M.W., Mitchell, J., Jawitz, J.W. and Reddy, K.R. (2010) Soil Phosphorus Flux from Emergent Marsh Wetlands and Surrounding Grazed Pasture Uplands. Ecological Engineering, 36, 1392-1400.
[5] Hu, M., Sardans, J., Le, Y., Yan, R., Zhong, Y. and Penuelas, J. (2022) Effects of Wetland Types on Dynamics and Couplings of Labile Phosphorus, Iron and Sulfur in Coastal Wetlands during Growing Season. Science of the Total Environment, 830, 1-10.
[6] Adam, K., Krogstad, T., Vrale, L., Sovik, A.K. and Jenssen, P. (2007) Phosphorus Retention in the Filter Materials Shellsand and Filtralite P®-Batch and Column Experiment with Synthetic P Solution and Secondary Wastewater. Ecological Engineering, 29, 200-208.
[7] Newman, S., Reddy, K.R., DeBusk, W.F., Wang, Y. Shih, G. and Fisher, M.M. (1997) Spatial Distribution of Soil Nutrients in a Northern Everglades Marsh: Water Conservation Area 1. Soil Science Society of America Journal, 61, 1275-1283.
[8] Diaz, O.A., Daroub, S.H., Stuck, J.D., Clark, M.W., Lang, T.A. and Reddy, K.R. (2006) Sediment Inventory and Phosphorus Fractions for Water Conservation Area Canals in the Everglades. Soil Science Society of America Journal, 70, 863-871.
[9] Wright, A.L. and Reddy, K.R. (2001) Phosphorus Loading Effects on Extracellular Enzyme Activity in Everglades Wetland Soils. Soil Science Society of America Journal, 65, 588-595.
[10] Richardson, C.J. (1985) Mechanisms Controlling Phosphorus Retention Capacity in Freshwater Wetlands. Science, 228, 1424-1427.
[11] Daroub, S.H., Stuck, J.D., Lang, T.A., Diaz, O.A. and Chen, M. (2003) Implementation and Verification of BMPs for Reducting P Loading in the EAA. Final Project Report Submitted to Everglades Agricultural Area Environmental Protection District and the Florida Department of Environmental Protection, Tallahassee, FL.
[12] Bostrom, B., Jasson, M. and Forsberg, C. (1982) Phosphorus Release from Lake Sediments. Archiv Fuer Hydrobiologie, 18, 5-59.
[13] Reddy, K.R., Diaz, O.A., Scinto, L.J. and Agamini, M. (1995) Phosphorus Dynamics in Selected Wetlands and Streams of Lake Okeechobee Basin. Ecological Engineering, 5, 183-207.
[14] Reddy, K.R., O’Connor, G.A. and Gale, P.M. (1998) Phosphorus Desorption Capacities of Wetland Soils and Stream Sediments Impacted by Dairy Effluent. Journal of Environmental Quality, 27, 438-447.
[15] Rigler, F.H. (1964) The Phosphorus Fractions and Turnover Time of Inorganic Phosphorus in Different Types of Lakes. Limnology Oceanography, 6, 165-174.
[16] Hedley, M.J., Stewart, J.W.B. and Chauhan, B.S. (1982) Changes in Inorganic and Organic Soil Phosphorus Fractions Induced by Cultivation Practices and by Laboratory Incubations. Soil Science Society of America Journal, 46, 970-976.
[17] Dai, K.H., David, M.B., Vance, G.F. and Krzyszowska, A.J. (1996) Characterization of Phosphorus in a Spruce-Fir Spodosol by Phosphorus-31 Nuclear Magnetic Resonance Spectroscopy. Soil Science Society of America Journal, 60, 1943-1960.
[18] Bowman, R.A. and Moir, J.O. (1993) Basic EDTA as an Extractant for Soil Organic Phosphorus. Soil Science Society of America Journal, 57, 1516-1518.
[19] Cade-Menun, B.J. and Preston, C.M. (1996) A Comparison of Soil Extraction Procedures for 31P-NMR Spectroscopy. Soil Science, 161, 770-785.
[20] New York State Department of State, NYDOS. (2012) Coastal Fish and Wildlife Rating Form, Pelham Bay Wetlands. Pelham Bay Park Wetlands Designated.
[21] US EPA ESS Method 310.1: Ortho-Phosphorus, Dissolved Automated, Ascorbic Acid. Environmental Science Section, Inorganic Chemistry Unit, Wisconsin State Lab of Hygiene, 465 Henry Mall, Madison, WI 53706. Revised October 1992.
[22] Pant, H.K. and Reddy, K.R. (2001) Phosphorus Sorption Characteristics of Esturarine Sediments under Different Redox Conditions. Journal of Environmental Quality, 30, 1474-1480.
[23] Gadian, D.G., Radha, G.K., Richards, R.E. and Seeley, P.J. (1979) 31P NMR in Living Tissue: The Road from a Promising to an Important Tool in Biology. In: Shulman, R.G., Ed., Biological Application of Magnetic Resonance, Academic Press, New York, 463-535.
[24] Pant, H.K., Warman, P.R. and Nowak, J. (1999) Identification of Soil Organic Phosporus by 31P Nuclear Magnetic Resonance Spectroscopy. Communications in Soil Science and Plant Analysis, 30, 757-772.
[25] JMP Version 7.0. (2016) SAS Institute Inc., Cary, NC.
[26] Tao, G.S., Tian, S.J., Cai, M.Y. and Xie, G.H. (2008) Phosphate-Solubilizing and -Mineralizing Abilities of Bacteria Isolated from Soils. Pedosphere, 18, 515-523.
[27] SPSS (2006) Version 15.0. SPSS, Chicago, IL.
[28] Wang, J. and Pant, H.K. (2010) Enzymatic Hydrolysis of Organic Phosphorus in River Bed Sediments. Ecological Engineering, 36, 963-968.
[29] Kunishi, H.M., Taylor, A.W., Heald, W.R., Gburek, W.J. and Weaver, R.N. (1972) Phosphate Movement from an Agricultural Watershed during Two Rainfall Periods. Journal of Agricultural and Food Chemistry, 20, 900-905.
[30] Wang, J. and Pant, H.K. (2009) Identification of Organic Phosphorus Compounds in the Bronx Bed Sediments by Phosphorus-31 Nuclear Magnetic Resonance Spectroscopy. Environmental Monitoring and Assessment, 171, 309-319.
[31] Wang, J. and Pant, H.K. (2010) Phosphorus Sorption Characteristics of the Bronx River Bed Sediments. Chemical Speciation & Bioavailability, 22, 171-181.
[32] Wang, J. and Pant, H.K. (2011) Assessments of Potential Spatial-Temporal Variations in Phosphorus Distribution and Fractionation in River Bed Sediments. Clean-Soil Air Water, 39, 148-156.
[33] Wang, J. and Pant, H.K. (2011) Land Use Impact on Bioavailable Phosphorus in the Bronx River, New York. Journal of Environmental Protection, 2, 342-358.
[34] Wang, J. and Pant, H.K. (2012) Estimation of Phosphorus Bioavailability in the Water Column of the Bronx River, New York. Journal of Environmental Protection, 3, 316-323.
[35] Sonzogni, W.C., Chapra, S.C., Armstrong, D.E. and Logan, T.J. (1982) Bioavailability of Phosphorus Inputs to Lakes. Journal of Environmental Quality, 11, 555-563.
[36] Bowman, R.A. and Cole, C.V. (1978) Transformations of Organic Phosphorus Substrates in Soils as Evaluated by NaHCO3 Extraction. Soil Science, 125, 49-54.
[37] Ivanoff, D.B., Reddy, K.R. and Robinson, S. (1998) Chemical Fraction of Organic Phosphorus in Selected Histosols. Soil Science, 163, 36-45.
[38] Wang, J. (2014) Combined Sewer Overflows (CSOs) Impact on Water Quality and Environmental Ecosystem in the Harlem River. Journal of Environmental Protection, 5, 1373-1389.

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