Phosphite flux at the sediment–water interface in northern Lake Taihu

Science of the Total Environment 543 (2016) 67–74

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Phosphite flux at the sediment–water interface in northern Lake Taihu Huimin Qiu, Jinju Geng ⁎, Hongqiang Ren, Zhaoyi Xu State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, China

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• Phosphite migrates from Lake Taihu water into sediment most of the year. • Higher annual average negative fluxes appeared in site with higher P level. • No correlation was found between phosphite flux and sedimentary ORP. • Phosphite flux showed a significant negative correlation with its concentration. • Flux showed significant negative correlations with sedimentary OM and Ca–Ps.

a r t i c l e

i n f o

Article history: Received 29 August 2015 Received in revised form 23 October 2015 Accepted 27 October 2015 Available online xxxx Keywords: Phosphite Flux Sediment–water interface Lake Taihu P cycle

a b s t r a c t 2− Phosphite (H2PO− 3 , HPO3 , +3 valence), a reduced form of phosphorus (P), has been widely detected in water environments. The role of phosphite in the P biogeochemical cycle has not been investigated systematically and quantitative results on phosphite fluxes are lacking. In this study, intact sediment core simulation was employed to measure the flux of phosphite at the sediment–water interface in northern Lake Taihu. Phosphite fluxes (μmol m−2 d−1) ranged from −38.21 ± 1.14 to 7.10 ± 2.18, with an annual average of −4.72 ± 10.40. On the whole, phosphite migrated from water into sediment and the sediment was primarily a sink. The highest seasonal negative phosphite fluxes (μmol m−2 d−1) occurred in winter (−10.44 ± 18.63), followed by summer (−8.04 ± 5.61) and spring (−2.61 ± 4.17). In autumn, phosphite flux was 2.20 ± 4.07. Higher annual average negative fluxes of phosphite (μmol m−2 d−1) appeared in site ZSB (− 12.70 ± 17.96), which contained the highest content of total soluble P. The average yearly migration of phosphite in Lake Taihu from water to sediment was estimated to be (4.04 ± 8.88) × 106 mol y−1. The transfer of phosphite from water into sediment usually occurs in winter may due to the season's natural tendency to create more favorable conditions for phosphite biogeochemical reactions. Phosphite fluxes showed significant negative correlations with the original phosphite concentration in water (r = −0.840, p b 0.01), as well as organic matter (r = −0.720, p b 0.01) and phosphate bound to Ca (Ca-Ps) (r = −0.632, p b 0.05) in sediment. These results indicate that microbiological processes and P species bound to Ca may play an important role in the P redox cycle. No significant correlations between phosphite fluxes and dissolved oxygen or oxidation-reduction potential were observed. © 2015 Published by Elsevier B.V.

⁎ Corresponding author at: State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing 210023, China. E-mail address: [email protected] (J. Geng).

http://dx.doi.org/10.1016/j.scitotenv.2015.10.136 0048-9697/© 2015 Published by Elsevier B.V.

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1. Introduction Conventionally it is considered that orthophosphate is the only form that can be directly involved in biological phosphorus (P) utilization, 2− but researchers have found that phosphite (H2PO− 3 , HPO3 , + 3 valence) exists and plays important roles in early life (Pasek et al., 2008, 2013). Phosphite, a form of reduced P species (with a valence lower than +5), has been found to widely exist in water environment. Pech et al. (2009) confirmed the presence of 0.06 ± 0.02 μmol L−1 phosphite in a geothermal pool. Pasek et al. (2014) detected that 0 to 33% of dissolved P was phosphite in freshwater samples from central Florida. Morton et al. (2005) detected a high concentration (mmol L−1 level) of phosphite in process water samples from an elemental P production plant. Han et al. (2012) reported 0.01 to 0.17 μmol L−1 of phosphite in water of Lake Taihu. In a full-scale municipal wastewater treatment plant, the concentration of phosphite was detected ranging from 0.15 ± 0.03 μmol L−1 to 1.11 ± 0.11 μmol L−1 in influent and from 0.04 ± 0.02 μmol L− 1 to 0.14 ± 0.03 μmol L− 1 in effluent (Yu et al., 2015). Phosphite also has been found in the oxidation products of schreibersite (Fe,Ni)3P in water (Bryant et al., 2009; Bryant and Kee, 2006; Pasek, 2008). In agriculture, synthetic phosphite has been widely marketed as a fungicide or as a potential P fertilizer (Thao and Yamakawa, 2009). It has been shown that many kinds of microorganism can uptake and convert phosphite into phosphate (Cao et al., 2012; Stone and White, 2012; White and Metcalf, 2007). For instance, in Lake Taihu, a phosphite degrading bacterial strain was identified by Cao et al. (2012). Martinez et al. (2012) found that phosphite could be utilized by the marine picocyanobacterium Prochlorococcus MIT9301, the numerically dominant primary producer in the oligotrophic ocean. Phosphite can boost cell numbers and chlorophyll-a content of Microcystis aeruginosa (one of the predominant species involved in algal blooms in freshwater lakes of China), as long as phosphate is provided simultaneously (Zhang et al., 2011). P is the limiting nutrient in many ecosystems, for phosphate is poorly soluble and slow to dissolve in natural conditions (Pasek et al., 2014). However, phosphite is more soluble and more reactive than orthophosphate, capable of forming condensed phosphates, organic C–P compounds, and even C–O–P compounds (Pasek et al., 2007). As a P form with + 3 valence, phosphite can be oxidized to phosphate (+ 5 valence) or reduced to gaseous phosphine (PH3 , − 3 valence) (Bolduc and Goe, 1974; Pasek et al., 2008; Pasek et al., 2014; Ratjen and Gerendas, 2009). These results strongly indicate that phosphite is a considerable component of the biogeochemical P cycle. While limited research is focused on the source and migration of phosphite in aquatic environments. In China, many lakes are suffering from increasing nutrients such as P and nitrogen and other chemical inputs, especially Lake Taihu, which is becoming increasingly eutrophied and has experienced annual lakewide cyanobacterial blooms in recent decades (Duan et al., 2014; Yang and Liu, 2010). P is the limiting nutrient in Lake Taihu (Jin et al., 2006; Paerl et al., 2011). The phosphate flux, in Lake Taihu, was estimated at about − 3.88 ± 1.26 mg m−2 d− 1 to 5.67 ± 1.91 mg m−2 d−1 (Fan et al., 2006). The migration of gaseous PH3, has been detected with a flux of 0.0138 ± 0.005 pg dm−2 h− 1 from lake sediment to water. That causes water PH3 concentration up to 0.178 ± 0.064 pmol dm−3 (Geng et al., 2005). Han et al. (2012) first detected 0.01 ± 0.01 to 0.17 ± 0.01 μmol L− 1 of phosphite in Lake Taihu. Those previous works mainly focused on the field determination and the flux of PH3 at the water environment (Geng et al., 2005; Han et al., 2011a, b). The role of phosphite in the biogeochemical cycle of P has not been investigated systematically. Thus far, reports on phosphite are mainly restricted to phosphite detection (Han et al., 2012; Han et al., 2013), whereas quantitative results on phosphite fluxes are currently lacking. Knowledge of phosphite's migration is necessary for understanding its roles in global biogeochemical cycling.

The purpose of this study was to measure the flux and estimate the emission level of phosphite at the water–sediment interface by sediment cores simulation, and to investigate the potential correlations between phosphite flux and other environment factors in lake water and sediment. 2. Materials and methods 2.1. Study area Lake Taihu, the third largest fresh water lake of China, is a large shallow lake with an area of 2338 km2 with an annual average water depth of 1.9 m and a volume of 4.66 × 109 m3 (Yang and Liu, 2010; Ye et al., 2009). The area surrounding Lake Taihu is the most economically developed region, and it is also the most densely populated area (Hu et al., 2006; Lu et al., 2010). It is the primary drinking water source for 30 million residents in the lake basin. However, Lake Taihu has been in a eutrophic state since the 1980s, and it is currently considered hypertrophic (Stone, 2011; Ye et al., 2009). Northern Lake Taihu was selected in this study, given that the northern part of Lake Taihu has a heavier pollution level (Bai et al., 2009; Paerl et al., 2011) and a higher phosphite concentration (Han et al., 2012; Han et al., 2013). This choice also makes it comparable with other studies about other forms of phosphorus like phosphate and phosphine which have been studied in this area (Fan et al., 2006; Geng et al., 2005; Han et al., 2011b). Four sites located in Dapu Harbor (DPH), Zhushan Bay (ZSB), Meiliang Bay (MLB) and Gonghu Bay (GHB) in northern Lake Taihu were selected (Fig. 1) for sample extraction. The collections carried out in the spring did not include site GHB. The specific latitudes and longitudes of each site are listed in Table 1. 2.2. Sample collection On 21 March, 22 June, 2 November and 31 December of 2013, two parallel surficial intact sedimentary column cores with a height of 25 cm were vertically collected from the selected four sites with a 6.0 × 50 cm cylindrical sediment tube sampler (purchased from Nanjing Institute of Geography and Limnology, China). Simultaneously, in situ overlying water in each site was collected by a plexiglass water sampler with a volume of 5 L. After fully covered with overlying water, the tubes filled with fresh sediment column cores and overlying water were

Fig. 1. Sampling sites in Lake Taihu.

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2.4. Analytical methods

Table 1 Latitude and longitude of each sampling site in Lake Taihu. Site

Location

MLB ZSB DPH GHB

31o27′12 N, 120o10′00 E 31o27′12 N, 120o02′38 E 31o18′19 N, 119o57′44 E 31o24′10 N, 120o20′07 E

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2.4.1. Physicochemical properties of the lake water The physicochemical properties of lake water in situ were measured by a multi parameter analyzer (HACH HQ40d, America) which is portable and can afford the measurement of multiple environmental parameters, such as temperature (T), oxidation-reduction potential (ORP), dissolved oxygen (DO), pH and conductivity. The results were represented as “mean values ± standard deviation” based on four sites (Table 2).

sealed at both ends with rubber stoppers and care was taken not to introduce air bubbles. The tubes with samples were set in a vertical position during the transfer process. The sediment column cores and overlying water were not disturbed during the field coring and transport process. Another surficial sedimentary sample at four sites was also extracted and packed into black plastic bags for various physicochemical analyses. In addition, bottom water samples were collected about 0.5 m above the water–sediment interface to replenish water during lab experiments. These water samples were collected in situ into fully filled HDPE (high density polyethylene) bottles. All the samples were transported to the laboratory within 6 h. The replenishment water samples were filtered through 0.45 μm filter to remove particles immediately after transfer to the laboratory. Then all the samples were stored at 4 °C in darkness before analysis and experiment.

2.3. Simulated phosphite flux experiment at the sediment–water interface The simulated phosphite flux experiments were carried out using methods applied to the research of phosphate fluxes by Fan et al. (2006). The intact sediment cores' incubation simulation experiment, by which the sediments can maintain the original state without disturbing inner structure (Boers and van Hese, 1988; Fan et al., 2002), was used to measure the phosphite flux at sediment–water interface. Original overlying water above the sediment column cores was replaced by carefully siphoning the filtered in situ water. The liquid level was kept 20 cm above the sediment–water interface. Plexiglass tubes filled with filtered in situ water without sediment were set as control samples. Then, all of the tubes with their upper opening were placed in a constant temperature incubator in darkness. The temperature was set according to the actual temperature when sampling. 10 mL of water was sampled 10 cm above the interface by pipet at 0, 0.25, 0.5, 1, 1.5, 2, 3 and 4 day to analyze the concentration of phosphite. After each sampling, in situ filtered water was replenished to the original scale immediately.

2.4.2. Determination of phosphite concentration in water Since phosphite is unable to react with the molybdenum-antimony reagent, it cannot be determined by the colorimetric method as phosphate (Han et al., 2012; Morton et al., 2005). The two-dimensional ion chromatography (ICS 5000, Thermo Fisher, USA) was chosen to determine the concentration of phosphite. The advantage of ICS 5000 is the series connection of the analytical system and the capillary system. Ions get the preliminary separation in analytical column of Dimension 1. According to the different retention times in Dimension 1, the target ions can then be switched to the capillary column of Dimension 2 for further separation and detection. With the removal of many interfer2− ence ions (Cl−, NO− 3 , SO4 and so on) in Dimension 1 and the lower detection limit of the capillary column in Dimension 2, the sample can now be tested for trace materials in natural water with complex matrix and small injection volume (Han et al., 2012; Qiu et al., 2013). To analyze phosphite in water, 5 mL of water was filtered through a 0.22 μm membrane, treated with the OnGuard RP cartridge (1.0 cm3, Thermo Fisher, USA) and the OnGuard H cartridge (1.0 cm3, Thermo Fisher, USA) to remove the hydrophobic organic substances and heavy metal ions. 1 mL of treated samples was prepared for analysis by ICS 5000 to separate and determine the concentration of phosphite. The ICS 5000 consisted of pumps, electrolytic eluent generators, carbonate removal devices and suppressed conductivity detectors. The analytical system (Dimension 1) was equipped with an IonPac AS11-HC column (250 mm × 2 mm), and the capillary ion chromatography system (Dimension 2) was equipped with a capillary MAX-100 column (250 mm × 0.25 mm). The process in Dimension 1 for gradient elution with potassium hydroxide (KOH) solution at a rate of 0.38 mL min−1 was set as follows: 0–10 min, 10–20 mM; 10–20 min, 20–50 mM; 20–25 min, 50–20 mM; 25–30 min, 20–10 mM. The injection volume was 25 μL and the detector cell temperature was held at 30 °C. According to the retention time in the Dimension 1, phosphite was switched into Dimension 2 from 8.10 to 8.80 min. The process of gradient elution

Table 2 Physicochemical properties of in-situ measurement of the lake water. Site

Season

T (°C)

Conductivity (μS cm−1)

ORP (mv)

DO (mg L−1)

pH

C0 (μmol L−1)

TSP (μmol L−1)

MLB

Spring Summer Autumn Winter Spring Summer Autumn Winter Spring Summer Autumn Winter Summer Autumn Winter

13.0 28.0 18.2 4.4 12.6 28.0 18.2 4.6 12.1 27.8 18.2 5.1 26.7 17.6 4.6

278 608 477 418 389 570 498 562 315 575 461 298 557 413 243

220.6 201.6 168.3 46.0 187.7 204.8 172.8 21.8 192.1 209.8 230.5 −11.5 211.6 168.4 53.2

9.97 5.41 / 15.05 9.14 4.02 / 11.81 9.42 5.93 / 13.14 / / 14.44

8.25 6.53 8.28 8.31 7.97 6.50 8.29 8.13 8.11 6.60 8.60 8.18 6.55 8.33 8.5

0.07 ± 0.01 0.65 ± 0.00 0.49 ± 0.05 0.19 ± 0.00 0.63 ± 0.03 0.62 ± 0.01 0.50 ± 0.02 1.55 ± 0.05 0.13 ± 0.01 0.87 ± 0.01 0.47 ± 0.03 n.d. 0.58 ± 0.00 0.42 ± 0.05 0.07 ± 0.00

5.24 ± 0.00 Lost 0.91 ± 0.00 0.51 ± 0.07 6.15 ± 0.03 Lost 1.08 ± 0.03 2.74 ± 0.00 1.79 ± 0.00 Lost 0.64 ± 0.00 0.44 ± 0.00 Lost 0.47 ± 0.03 0.51 ± 0.00

ZSB

DPH

GHB

C0: the initial concentration of phosphite of in situ lake water. TSP: total soluble P. /: not detected because of instrument failure.

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with KOH solution in Dimension 2 at a rate of 0.01 mL min−1 was set as follows: 0–5 min, 2–4 mM; 5–35 min, 4 mM; 35–40 min, 30 mM; 40–45 min, 2 mM. The detection limit (S/N = 3) was 0.18 nM, the relative standard deviation (RSD, n = 6) of peak area was 2.7% and the recoveries were 101.0%–118.4% (0.1, 0.5 and 1 μM of phosphite). More details can be seen in our earlier study (Qiu et al., 2013). 2.4.3. Determination of sedimentary P fractions and organic matter content Samples of sediment were also subjected to various physicochemical analyses. About 10 g of sedimentary samples were freeze-dried, ground and homogenized by passing through an 80 mesh sieve. By the chemical sequential extraction scheme, sedimentary P fractions were divided into inorganic phosphate (IPs), exchangeable phosphate (EPs), dissolved phosphate (DPs), phosphate bound to Al (Al–Ps), phosphate bound to Fe (Fe–Ps), phosphate bound to Ca (Ca–Ps), and organic phosphate (Org–Ps) and were then measured using the molybdenumantimony spectrophotometric method with the advantage of simple operating, accurate results and low cost (Agricultural Chemistry Specialty Council of Edaphic Society in China, 1989). Organic content was measured by comparing the weight before and after calcination at 550 °C in a muffle furnace for 5 h. The sedimentary P fractions and organic matter content are shown in Table 3. 2.5. Data analysis Phosphite flux (Fi) at the sediment–water interface was calculated according to Fan et al. (2002). h i n F i ¼ V ðC n  C 0 Þ þ ∑n¼1 V n1 ðC n1  C a Þ =ðS  t Þ

Fi is the phosphite flux (μmol m−2 d− 1) for experimental group or corresponding control group, separately, and the net phosphite flux is the difference of the experimental group minus the control group. V is the volume of overlying water (L). Vn − 1 is the sampling volume at time (n − 1). C0, Cn and Cn − 1 are the phosphite concentrations at times 0, n and n − 1 in the overlying water (μmol L−1), respectively, Ca is the phosphite concentration in the replenishment filtered water (μmol L−1). S is the surface of the sediment (m2), and t is the experiment time (d). Correlation analyses were performed to determine relationships between phosphite fluxes and environmental variables. All statistical analyses were conducted using SPSS Statics 22 (IBM, USA). Correlations were symbolized as * at p b 0.05 and ** at p b 0.01.

Fig. 2. Seasonal and spatial variations of phosphite flux across the sediment–water interface in northern Lake Taihu.

3. Results 3.1. Simulated seasonal phosphite fluxes at sediment–water interface in northern Lake Taihu Fig. 2 shows the seasonal fluxes of phosphite across the sediment– water interface in northern Lake Taihu. The fluxes (μmol m−2 d− 1) ranged from − 38.21 ± 1.14 to 7.10 ± 2.18, with the average of − 4.72 ± 10.40. The fluxes of phosphite were negative for the most part, which means the migration direction of phosphite was from water to sediment. The average yearly migration value of phosphite from Lake Taihu water into sediment was estimated to be (4.04 ± 8.88) × 106 mol y− 1. The highest annual average negative fluxes (μmol m−2 d−1) of phosphite appeared in site ZSB (−12.70 ± 17.96), followed by site MLB (− 4.48 ± 5.40), site DPH (− 3.59 ± 5.68) and site GHB (1.87 ± 3.50). The highest average seasonal negative fluxes (μmol m− 2 d− 1) occurred in winter (− 10.44 ± 18.63), followed by summer (−8.04 ± 5.61), spring (−2.61 ± 4.17) and autumn (2.20 ± 4.07). The seasonal and spatial variations of phosphite concentration in water of northern Lake Taihu are shown in Fig. 3. The concentrations (μmol L−1) were between 0 and 1.55 ± 0.05, with an average concentration of 0.48 ± 0.39. The highest annual average phosphite concentration (μmol L−1) appeared in site ZSB (0.83 ± 0.49), more than twice the concentrations of other sites: DPH (0.37 ± 0.39), GHB (0.36 ± 0.26) and MLB (0.35 ± 0.27). The highest seasonal concentration of phosphite

Table 3 Sedimentary P fractions and organic matter content. Site

MLB

ZSB

DPH

GHB

Season

Sediment moisture (%)

Organic matter (g kg−1)

IPs (mg kg−1)

EPs (mg kg−1)

DPs (mg kg−1)

Al–Ps (mg kg−1)

Fe–Ps (mg kg−1)

Ca–Ps (mg kg−1)

Org–Ps (mg kg−1)

TPs (mg kg−1)

Spring Summer Autumn Winter Spring Summer Autumn Winter Spring Summer Autumn Winter Summer Autumn Winter

48.9 ± 1.6 42.7 ± 0.9 44.2 ± 1.6 37.1 ± 4.2 48.9 ± 3.3 51.9 ± 0.4 49.3 ± 3.3 43.6 ± 0.2 41.7 ± 1.1 38.8 ± 5.7 42.9 ± 0.3 39.5 ± 05 41.7 ± 1.8 39.4 ± 1.6 58.5 ± 0.4

47.0 ± 0.3 46.6 ± 3.8 27.9 ± 0.1 29.4 ± 1.0 67.8 ± 1.3 49.6 ± 0.6 38.9 ± 4.3 76.0 ± 22.0 41.2 ± 3.6 38.5 ± 8.1 36.2 ± 0.6 32.1 ± 0.1 37.1 ± 3.6 26.8 ± 0.4 55.8 ± 0.9

264.5 ± 2.5 377.0 ± 3.7 190.3 ± 0.9 174.0 ± 2.0 639.5 ± 14.7 555.7 ± 2.8 414.0 ± 69.2 404.5 ± 17.8 341.7 ± 6.2 518.7 ± 2.8 293.6 ± 5.4 227.2 ± 8.4 218.7 ± 2.8 211.2 ± 2.7 272.0 ± 7.9

2.0 ± 0.3 1.3 ± 0.1 1.1 ± 0.3 1.5 ± 0.3 2.2 ± 0.1 1.5 ± 0.3 1.1 ± 0.1 1.8 ± 0.0 1.3 ± 0.4 2.3 ± 0.1 1.3 ± 0.0 1.1 ± 0.1 1.2 ± 0.3 1.3 ± 0.0 1.7 ± 0.1

2.4 ± 0.6 1.3 ± 0.6 1.1 ± 0.1 0.9 ± 0.1 3.4 ± 0.6 3.3 ± 0.6 1.5 ± 0.4 1.0 ± 0.1 2.6 ± 0.4 1.0 ± 0.1 1.3 ± 0.0 1.2 ± 0.1 2.2 ± 1.6 1.3 ± 0.3 1.2 ± 0.1

3.7 ± 0.4 2.4 ± 0.1 3.9 ± 0.5 0.5 ± 0.0 4.3 ± 0.4 2.0 ± 0.2 2.6 ± 0.4 0.6 ± 0.2 2.6 ± 0.4 2.1 ± 0.3 1.1 ± 0.2 0.0 ± 0.0 2.9 ± 0.5 1.7 ± 0.8 0.1 ± 0.1

74.9 ± 3.0 99.2 ± 0.0 58.2 ± 2.8 58.1 ± 0.0 180.7 ± 4.4 343.9 ± 2.1 230.9 ± 2.1 72.1 ± 2.2 57.9 ± 2.2 124.3 ± 2.1 138.5 ± 5.0 42.6 ± 0.7 76.2 ± 2.9 76.6 ± 1.4 73.6 ± 0.7

118.3 ± 6.6 257.4 ± 4.7 141.0 ± 1.1 110.0 ± 1.1 265.7 ± 7.7 182.0 ± 2.5 180.5 ± 5.6 261.1 ± 1.5 161.0 ± 46.6 341.7 ± 0.7 188.3 ± 2.1 163.9 ± 1.8 145.8 ± 2.1 137.8 ± 0.7 132.1 ± 5.5

288.3 ± 35.3 51.9 ± 11.4 42.0 ± 2.3 198.0 ± 12.8 84.9 ± 26.4 120.0 ± 8.5 100.2 ± 30.9 146.8 ± 50.2 50.3 ± 2.0 60.4 ± 67.8 66.4 ± 20.7 205.9 ± 58.1 130.8 ± 3.9 0.65 ± 0.0 317.2 ± 13.8

552.7 ± 35.3 428.9 ± 7.6 232.3 ± 3.2 372.0 ± 14.8 724.4 ± 26.4 675.8 ± 5.7 514.3 ± 38.3 551.3 ± 68.0 392.0 ± 26.4 579.1 ± 50.7 360.0 ± 3.8 433.1 ± 64.0 349.5 ± 6.7 213.7 ± 23.7 489.3 ± 21.7

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Table 5 Correlation between phosphite and environmental conditions in lake water. Phosphite fluxes (y, μmol m−2 d−1)

Environmental conditions (x)

T (°C) Conductivity (μS cm−1) ORP (mv) DO (mg L−1) pH TSP (umol L−1) C0 (μmol L−1)

r

p

Equation

n

−0.105 −0.469 0.271 0.144 0.216 −0.336 −0.840

0.710 0.078 0.329 0.692 0.440 0.312 0.000⁎⁎

y = 0.126x − 7.048 y = −0.042x + 13.511 y = 0.036x − 10.467 y = 0.433x − 12.671 y = 2.969x − 28.084 y = −1.925x − 0.439 y = −24.291x + 7.399

15 15 15 10 15 11 15

⁎⁎ Indicates the correlation is significant at the 0.01 level.

Fig. 3. Seasonal and spatial variations of phosphite in water of northern Lake Taihu.

(μmol L−1) occurred in summer (0.71 ± 0.13), followed by winter (0.58 ± 0.74), autumn (0.49 ± 0.04) and spring (0.28 ± 0.31). 3.2. Correlation analysis between phosphite fluxes and environmental factors Evaluating the possible connection of phosphite fluxes and sedimentary characters at all sites throughout the year, a significant negative correlation between phosphite fluxes and sedimentary organic matter (r = − 0.720, p b 0.01, n = 15) and a negative correlation between phosphite fluxes and Ca–Ps (r = − 0.632, p b 0.05, n = 15) were observed (Table 4). Evaluating the influence of environmental conditions in lake water on phosphite fluxes at all sites throughout the year, a negative correlation between phosphite fluxes and the original phosphite concentration in situ water (r = −0.840, p b 0.01, n = 15) was observed (Table 5). No significant correlations between phosphite flux and other factors (T, DO, ORP etc.) were observed. 4. Discussion 4.1. Seasonal and spatial variations of phosphite fluxes at the sediment–water interface The time dependent variations of phosphite fluxes from water to sediment were winter N summer N spring N autumn (Fig. 2). The highest average flux in winter may be caused by the unexpected highest flux in ZSB in winter. Overall, the fluxes in summer at the four sites were commonly at a higher level. Since the experiments were carried out in

darkness with the overlying water pre-filtered, there were no photosynthetic reactions in the overlying water. So the faster migration of phosphite from water into sediment in summer may have been caused by rapid microbial growth in sediments during this period. Rapid microbial growth would lead to the depletion of P in sediment (Fan et al., 2006). Since phosphite could also be a P source for many organisms (Cao et al., 2012; Stone and White, 2012), the consumption of phosphite in sediments could increase and pump phosphite into sediments. The transfer of phosphite from water to sediment usually occurs in winter may due to the season's natural tendency to create more favorable conditions for phosphite biogeochemical reactions. As a reduced product of phosphite, phosphine (PH3) is found higher in Arctic areas, and a higher production rate has been found at 4 °C compared to 25 °C after 72 h of incubation of three ornithogenic soil samples (Zhu et al., 2006). And thermodynamic calculations also proved that lower temperature promoted the production of phosphine from the disproportionation reaction of phosphite (Pasek et al., 2014). A higher annual average negative flux of phosphite appeared in site ZSB (−12.70 ± 17.96 μmol m−2 d−1). Meanwhile, site ZSB had higher concentrations of TSP and phosphite in water (Table 2, Fig. 3). A positive correlation was found between the concentrations of phosphite and TSP in water (r = 0.152, p = 0.655). Furthermore, a significant negative correlation was found between phosphite fluxes and phosphite concentration in water (r = −0.840, p b 0.01). So, the fluxes may be partly linked to the level of P pollution in water. Previous study has demonstrated that high concentrations of phosphite occurred along with a high P level (Han et al., 2013). The phosphite fluxes (− 38.21 ± 1.14 to 7.10 ± 2.18 μmol m−2 d−1) were smaller than the phosphate fluxes in Lake Taihu (− 0.13 ± 0.04 to 0.18 ± 0.06 mmol m−2 d− 1) (Fan et al., 2006) and a shallow small estuary of Palmones River in Southern Spain (with an average of 0.68 mmol m−2 d−1 (Clavero et al., 2000), and higher than the fluxes of PH3 in marsh and paddy fields (− 227 ± 57 to 276 ± 100 ng m−2 h− 1) (Han et al., 2011a). Given that phosphite accounts for 1.34–73.4 percentage of the TSP in Lake Taihu water (calculated by Table 2), the migration of phosphite is noteworthy. Its transformation can be an important supplement for

Table 4 Correlation between phosphite fluxes and sedimentary character and P fractions. Sedimentary character and P fractions (x)

Sediment moisture content (%) Organic matter (g kg−1) IPs (mg kg−1) EPs (mg kg−1) DPs (mg kg−1) Al-Ps (mg kg−1) Fe-Ps (mg kg−1) Ca-Ps (mg kg−1) Org-Ps (mg kg−1) TPs (mg kg−1) ⁎ Indicates the correlation is significant at the 0.05 level. ⁎⁎ Indicates the correlation is significant at the 0.01 level.

Phosphite fluxes (y, μmol m−2 d−1) r

p

Equation

n

0.071 −0.720 −0.424 −0.452 0.088 0.168 −0.019 −0.632 −0.003 0.437

0.803 0.002⁎⁎ 0.115 0.091 0.756 0.550 0.946 0.012⁎ 0.992 0.103

y = 9.102x − 9.110 y = −0.539x + 18.357 y = −0.032x + 5.762 y = −11.916x − 13.034 y = 1.115x − 9.110 y = 1.307x − 7.693 y = −0.003x – 4.755 y = −0.102 x + 13.842 y = −0.0003 x – 5.079 y = −0.032 x + 9.571

15 15 15 15 15 15 15 15 15 15

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phosphate or PH3 gas. Phosphate can be a P source, and phosphine can release from water into the ambient atmosphere (Han et al., 2011b). Since the calculation has cut off the phosphite fluxes of control groups without sediment, the negative fluxes meant that the sediment was the internal sink of phosphite. Han et al. (2013) found that lower levels of phosphite were found in deeper sediment material, suggesting that phosphite may migrate from surficial sediment to deep layer. Fan et al. (2006) thought that the adsorption by sediment and uptake by organisms was the main reason for the migration of phosphate from water to sediment. With an intermediate valence, phosphite can be transformed to phosphate or PH3 through chemical or biological process (Bolduc and Goe, 1974; Pasek et al., 2014; White and Metcalf, 2007). 10–67% culturable bacteria were capable of using phosphite as a sole P source (Stone and White, 2012). Therefore the migration of phosphite from water to sediment was possibly accelerated by the adsorption of sediment particles, and by phosphite consumption by microorganisms or other oxidants and reductants in sediment. This special property changes its migration through environments, especially in contrast to phosphate (with the highest oxidation state) and phosphine (with the lowest reduction state). Normally, phosphate and phosphine are released from sediment to water (Fan et al., 2006; Ozkan and Buyukisik, 2012; Zabel et al., 1998; Zhang et al., 2006). Geng et al. (2005) has determined an average yearly emission value of PH3 from Taihu Lake sediment to water of 28.3 ± 10.2 g y−1. 4.2. Possible relevance between the phosphite fluxes and lake water environment conditions/sedimentary characters 4.2.1. Correlation between phosphite fluxes and sedimentary character and P fractions There was a significant negative correlation between phosphite fluxes and organic matter (r = −0.720, p b 0.01, n = 15) in sediment (Table 4). In sediments, organic matter is potential sources of carbon and energy for microorganism metabolism. Higher organic matter reflects active microbiological processes that link to uptake and transform phosphite. The resulting concentration difference promotes the migration of phosphite from water to sediment. Many microorganisms can utilize and transform phosphite (Hanrahan et al., 2005; Pasek et al., 2014; White and Metcalf, 2007). Strain FiPS-3, a phosphite-oxidizing sulphate-reducing bacterium, can oxidize phosphite to phosphate

(Schink and Friedrich, 2000). Some microorganisms can reduce phosphite to PH3 (Pasek et al., 2014; Tsubota, 1959). Microbial uptake at the sediment–water interface was the most important sink for soluble reactive P (SRP) (Qu et al., 2001). Furthermore, there also may be a reduction pathway of phosphite to phosphine by microorganisms in sediment. Hou et al. (2009) reported that higher matrix PH3 was generally detected in the intertidal sediment enriched in organic P. They hypothesized that the microbial degradation of organic P probably contributes to the production of PH3 in the intertidal sediment, while Pasek et al. (2014) stated that reduction from phosphate is energetically unfeasible unless phosphite or hypophosphite are present in the environment. Roels and Verstraete (2001) reported that reduced P compounds including phosphite can be transformed through microbial mediation or disproportionation reactions into gaseous P compounds. A significant negative correlation was also observed between phosphite fluxes and Ca−Ps in sediment (r = − 0.632, p b 0.05, n = 15) (Table 4). Greater negative fluxes occurred in sites with greater Ca− Ps, a main fraction of IPs, showing phosphite might have an affinity for calcium. A significant positive relationship also exists between phosphine and Ca (p = 0.008) in soils (Han et al., 2011c). Phosphine has exhibited positive traits in its significant correlation to Ca−Ps other than other P fractions in Lake Taihu sediments (Geng et al., 2005), intertidal sediments (Hou et al., 2009), and paddy soils (Han et al., 2011c). Those results imply that P species bound to Ca may play an important role in the P redox cycle, even though the mechanism is still unclear. 4.2.2. Correlation between phosphite fluxes and environmental conditions in lake water A negative correlation was found between phosphite fluxes and original phosphite concentration in the in situ lake water samples (r = −0.840, p b 0.01, n = 15) (Fig. 4). This means that the migration of phosphite from water into sediment more likely occurred with a high concentration of phosphite in water. Positive correlations were observed between phosphite fluxes and DO or ORP (Table 5), even though they were not significant. With lower ORP or DO, the faster decrease of phosphite in water appeared (r = 0.271, p = 0.329, n = 15). This is different from the migration of phosphate, which is usually thought to trend toward release under anaerobic conditions (Ahlgren et al., 2011; Jin et al., 2006). The difference indicates that, the migration of phosphite might have a different

Fig. 4. Correlation between phosphite flux and environmental factors: (a) ORP; (b) DO; (c) T; (d) C0.

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mechanism from phosphate. With lower ORP, reduction is more likely to happen. Pasek et al. (2014) hypothesized that phosphite and hypophosphite are the ultimate source of phosphine in the atmosphere. A significant negative correlation was observed between phosphine concentration and ORP (Feng et al., 2008). And phosphine gas has been detected widely around the world including the Antarctic (Zhu et al., 2006), aquatic sediment of the Yangtze Estuary (Hou et al., 2009) and Lake Taihu (Geng et al., 2005). As a result, the transformation of phosphite to phosphine may also partly account for the loss of phosphite in the aquatic environment. In real environments, there may be many other factors besides organic matter and Ca-Ps in this study, which can affect the migration of phosphite, such as hydrodynamics, sunlight, wind, rainfall and so on. Or its migration may be a result of the combination of some above factors. The combination of real conditions and theoretical model which have done good job in environment area (Wang et al., 2014; Wu et al., 2009; Zhao et al., 2006), may provide us more information in the P cycle.

5. Conclusion The fluxes of phosphite (μmol m−2 d−1) at sediment–water interface in northern Lake Taihu ranged from − 38.21 ± 1.14 to 7.10 ± 8.18, with the average of −4.72 ± 10.40. The highest seasonal negative fluxes occurred in winter, and then followed by summer, spring and autumn. The sediment was primarily a sink. The average yearly migration value of phosphite from Lake Taihu water to sediment was estimated to be (4.04 ± 8.88) × 106 mol y−1. Even though the values of phosphite flux were smaller than the phosphate fluxes, the migration and transformation of phosphite should be a part in the P cycle. The sedimentary organic matter content and Ca–Ps showed significant negative correlations with phosphite fluxes. Greater negative fluxes occurred in conditions with more organic matter and Ca–Ps in sediment. Greater negative fluxes are also accompanied by higher levels of phosphite in water. As a P species with intermediate valence, phosphite migration regularity and influence factors could be more complicated, and different from phosphate (with the highest oxidation state) and phosphine (with the lowest reduction state). As an important part of P biogeochemical cycle, the microcosmic mechanism of phosphite migration and transformation in natural environmental needs more attention, especially the roles that microbiological processes and P species bound to Ca play.

Acknowledgment This work was supported by the National Natural Science Foundation of China (51278241, 41203062).

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