Journal of Hazardous Materials 266 (2014) 1–9

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Oxygen and phosphorus dynamics in freshwater sediment after the deposition of flocculated cyanobacteria and the role of tubificid worms Lei Zhang a,∗ , Qianjiahua Liao b , Xiaozhi Gu a , Wei He a , Zhe Zhang b , Chengxin Fan a a State Key Laboratory of Lake Science and Environment, Nanjing Institute of Geography and Limnology, Chinese Academy of Sciences, Nanjing 210008, PR China b Department of Environmental Science, China Pharmaceutical University, Nanjing 211198, PR China

h i g h l i g h t s • • • •

Algae deposition increased sediment O2 uptake and decreased O2 penetration depth. Algae deposition altered SRP flux, pore-water SRP profiles and P fractions. Tubificid worms transferred algae cells deeper in the sediment and mitigate their degradation. Worms enhanced the increase of SRP in pore water and loosely adsorbed P in sediment.

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Article history: Received 15 August 2013 Received in revised form 23 November 2013 Accepted 8 December 2013 Available online 12 December 2013 Keywords: Harmful algal blooms (HABs) Flocculation Oxygen uptake rate Phosphorous fractionation Taihu

a b s t r a c t Flocculation is a promising method for controlling harmful algal blooms; however, little is known about the effects of algae deposition by flocculation on benthic oxygen (O2 ) and nutrient dynamics. In this study, we aimed to investigate the influence of cyanobacteria flocculation deposition on benthic O2 and phosphorus (P) dynamics and the role of tubificid worms in the process. Chitosan and sediment particles were used to flocculate and deposit cyanobacteria cells onto lake sediment. The impulse deposition of algal flocculation degraded the deposited algal cells, which decreased the O2 penetration depth in sediment and increased the O2 uptake rate. Algae deposition also increased the soluble reactive P (SRP) in pore water and loosely adsorbed P in sediment, and changed SRP flux. Tubificid worms transported algal cells deeper into the sediment, mitigated their degradation, and altered the O2 penetration depth, but not the O2 uptake rate. Tubificid worms enhanced the increase in pore-water SRP and loosely adsorbed P in sediment. Therefore, the deposition of algal flocculation modifies the benthic O2 and P dynamics, and tubificid worms can mitigate or enhance some of these processes. © 2013 Elsevier B.V. All rights reserved.

1. Introduction Harmful algal blooms (HABs) are increasing world-wide due to anthropogenic nutrient enrichment [1,2]. They affect public health, tourism, fisheries, and ecosystems because they can lead to adverse tastes and odors and the presence of toxic materials [3,4]. Thus, determining methods for controlling HABs has become a challenging aquatic science task. Several technologies have been studied, including flocculation [5], coagulation–flotation [6], oxidation [7], and ultrasonic irradiation [8]. Of these control methods, flocculation is popular because it can quickly flocculate algal cells for

∗ Corresponding author at: Nanjing Institute of Geography and Limnology, CAS. 73# East Beijing Road, Nanjing 210008, PR China. Tel.: +86 25 86882210; fax: +86 25 57714759. E-mail addresses: [email protected], [email protected] (L. Zhang). 0304-3894/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.jhazmat.2013.12.008

deposition onto sediment. For example, modified soils [9], sediments [10], sands [11], and mineral materials [12] have been verified to mitigate algae blooms effectively through the flocculation of algal cells. Chitosan-modified soil and sediment are able to quickly flocculate and deposit algal cells, and more than 90% algal cells were removed in 1 h using 1 mg chitosan and 10 mg soil or sediment in 1 L lake water containing 4.86 × 109 Microcystis aeruginosa cells [9]. Take the price of chitosan in China for example, 1 kg chitosan is about 25 USD and it is able to dispose 1000 m3 water; and sediment and soils are able to be obtained locally, so the cost is low using chitosan modified soil and sediment to flocculate algal cells. This method has been used in a small bay in Lake Taihu, China [13]. Previous studies on algal flocculation have focused on algae removal efficiency, whereas little research has focused on the influence of algal flocculation on benthic oxygen (O2 ) and nutrient dynamics. However, natural algae or other organic matter can influence benthic element dynamics. For example, O2 penetration depth

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was decreased and volume-specific O2 consumption was increased by intense sedimentation during a seasonal algae bloom [14]. In addition, algal bloom deposition increased sediment respiration and altered the benthic nitrogen and phosphorus (P) dynamics [15,16]. The deposition of artificial algal flocculation is an imposed organic impulse for sediment. Will this artificial intense deposition influence the benthic O2 and nutrient dynamics? Benthic animals live in or on the sediment surface and are an important constituent of the benthic sediment ecosystem. In addition, the activities of benthic animals alter the sediment matrix and solute exchange across the sediment–water interface (SWI) [17]. Benthic animals can mix sediment particles [18], alter sediment stratification [19], and increase sediment porosity [20]. Their movements in sediment produce the heterogeneity of O2 , pH, and microbes [21,22]. Moreover, organic matter degradation [23,24], nutrient cycling [25–27], O2 dynamics [28], and ferrous iron concentrations [29] are altered by benthic animals. Benthic animals can burrow through the deposited flocculation layer [30] and alter the degradation of pulse-settled organic matter and nutrient fluxes across the SWI [31]. In addition, algal flocculation has been verified to influence benthic animal community structure and diversity after 11 months [32]. Thus, the questions of whether benthic animals will be influenced by the deposition of algal flocculation in short time and whether benthic animals will influence the degradation and survival of deposited algae need to be studied in detail. Phosphorus plays an important role in freshwater eutrophication and HABs [2,33]. Therefore, the control of P input is critical for the prevention of eutrophication in freshwater ecosystems [34]. In addition, the release of P from sediment to overlying water can influence algae development [33]. Oxygen is a central molecule for global element cycling and plays a key role in P cycling [28]. In freshwater ecosystems, cyanobacteria are the most common algal taxa that induce HABs [3]. In the present study, we aimed to investigate the influence of cyanobacteria flocculation deposition on benthic O2 and P dynamics. The role of tubificid worms in the process was examined simultaneously because they accumulate in high densities in eutrophic aquatic ecosystems. 2. Materials and methods 2.1. Field sampling Lake Taihu is a eutrophic lake in Eastern China, with an area of 2338 km2 . The cyanobacteria blooms in Taihu have received increasing attention in recent years because this lake is essential to the local fishery, water supply, environment, and tourism [1,35]. On July 9, 2012, sediment cores and lake water were sampled at the Dapu River estuary (31◦ 18 19.1 N, 119◦ 55 58.2 E) in western Taihu. Sediment cores were collected using plexiglas tubes (11 cm I.D., 50 cm long) and an 11 cm × 50 cm gravity corer. In addition, lake water was collected in plastic barrels. On August 2, surface sediment was collected by a Petersen Grab sampler from the same site and was screened with a 0.5 mm net to collect the tubificid worms (Limnodrilus hoffmeisteri) for this experiment. Cyanobacteria cells were collected along the shore of the Dapu River estuary using a plankton net on August 28. All samples were transported to the laboratory immediately after sampling. 2.2. Microcosms In the laboratory, the top 12 cm of each sediment core was sectioned into 0–4 cm, 4–8 cm, and 8–12 cm portions; the same portions from different cores were pooled together. Each pool was sieved using a 0.6 mm mesh to exclude macroinvertebrates and large particles and was then homogenized with a dough mixer. The

sediment pools were transferred into 24 Plexiglas tubes (11 cm I.D., 17 cm long) according to their original sequences and depths. Lake water was added to the sediment surface in each microcosm using intravenous needles, resulting in 24 microcosms with 12 cm of sediment and 5 cm of water. The microcosms were randomly separated into three groups with eight replicates. Three groups were assigned as the control treatment (no algae or tubificid worm addition, C), algae treatment (only algae addition, A), and algae and tubificid worm treatment (both algae and tubificid worm addition, A + T). The three microcosm groups were transferred to three corresponding water tanks and submerged in lake water. The water in each tank was aerated by a mini aerator to maintain O2 saturation. The microcosms were pre-incubated for 3 weeks before the tubificid worms were introduced. 2.3. Experimental design For each treatment, three microcosms were selected randomly for the measurement of O2 uptake rate and soluble reactive phosphorus (SRP) flux across the SWI, whereas the remaining five microcosms were used for pore-water sampling and for analysis of chlorophyll a (chl a) in the sediment. On August 3 and 4, O2 uptake rate and SRP flux were measured separately. On August 5, 285 tubificid worms were added to each microcosm (30,000 ind. m−2 ) in the A + T treatment. The addition density of tubificid worms was based on their density in our sampling area in Lake Taihu [36]. All microcosms were then incubated for 24 additional days (until August 29). To investigate the influence of tubificid worms on sediment, O2 and pH profiles were measured on August 24 using microsensors (Unisense, Denmark), and O2 uptake rate and SRP flux were examined again on August 26 and 27. On August 29, the A and A + T treatment microcosms were removed from the water tanks. A 30 mL concentrated cyanobacteria “solution” (the dry algal cells weighed 2.7 ± 0.04 g) was then gently dispersed into the overlying water of each microcosm. One milliliter of 1 g L−1 chitosan solution and 10 mL of mud that contained 3 g of dry sediment from the field sampling site, were added to the overlying water with a gentle stir to flocculate and deposit the algal cells. After the sedimentation of flocculated algae, 15 mL of mud (same as above) was gently dispersed on the flocculation surface to prevent resuspension of the algae. All microcosms were returned to their original tank and carefully submerged into the lake water. This time point was defined as day 0 of the experiment. The following incubation experiment lasted for 80 days, with water replacement every two weeks; a mini aerator was used in each tank to supply a sufficient amount of O2 to water. During the experiment, the O2 uptake rate was measured on days 6, 13, 20, 27, 41, 55, and 69, whereas the SRP flux was examined on days 7, 14, 21, 28, 42, 56, and 70. The O2 and pH profiles were measured on days 2, 7, 14, 21, 28, 42, 56, and 70. Pore water was acquired using a minipeeper with a vertical resolution of 4 mm [37]. Three minipeepers were prepared at a time, deoxidized under nitrogen, and then separately inserted into microcosms from the three treatment groups on days 6, 13, 27, 41, and 76. Three days later (on days 9, 16, 30, 44, and 79), the minipeepers were removed from the microcosms and carefully flushed with oxygen-free deionized water. Ferrous iron and SRP in the pore water were analyzed immediately. After the minipeeper was removed, the surface sediment (0–2 cm) was sectioned into layers and freeze-dried for analysis of the chl a content. For the first two time points, the sediment was sectioned at depths of 0–5 mm, 5–10 mm, and 10–20 mm; for the remaining time points, the sediment was sectioned into 10 layers with depths of 2 mm to illustrate the transportation of algae by tubificid worms. At the end of the experiment, one of the three microcosms used for O2 uptake rate and SRP flux measurements was sectioned into 0–2 cm, 2–4 cm,

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4–6 cm, and 6–8 cm layers for the examination of sediment P fractions. The other two microcosms were screened with a 0.5 mm net to collect the surviving tubificid worms. The measurements of O2 uptake rate and SRP flux were based on the change of O2 and SRP in the overlying water before and after the microcosm was sealed (with an O2 saturation of at least 80%). When the O2 uptake rate was measured, the O2 in the overlying water was examined using an oxygen microsensor before and after the microcosm was sealed. For the SRP flux measurement, 10 mL of overlying water was sampled before and after the microcosm was sealed.

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Fig. 1. Water temperature during the incubation experiment.

2.5. Statistical analysis Prior to the addition of algae, the differences in the O2 uptake rate and SRP flux among the three different treatments were examined using one-way analysis of variance (ANOVA) followed by a Tukey post hoc test. After the algae addition, two-way ANOVA followed by a Tukey post hoc test was used to examine the difference in the O2 uptake rate and SRP flux using treatment and time as the two factors. In addition, the differences in sediment P fractions were examined by two-way ANOVA with treatment and depth as the two factors. All of the statistical analyses were conducted using the SPSS 13.0 software (SPSS, USA).

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The Fe2+ and SRP in pore water were analyzed using the miniaturized photometrical method [38] with a microtiter plate and microtiter plate reader (BioTek Epoch, USA), which was modified from the ferrozine [39] and molybdenum blue [40] methods. Sediment samples for chl a analysis were freeze dried, ground, extracted with 90% acetone at 4 ◦ C [41], and examined using a UV–vis spectrophotometer (Shimadzu UV-2550, Japan). The sediment P fraction was analyzed using a sequential extraction, in which sediment was successively extracted with 0.46 mol L−1 NaCl, 0.11 mol L−1 NaBD (0.11 mol L−1 Na2 S2 O4 in 0.11 mol L−1 NaHCO3 ), 0.1 mol L−1 NaOH, and 0.5 mol L−1 HCl; the residue was then ashed at 550 ◦ C for 2 h and extracted with 1 mol L−1 HCl [42]. The P fractions were sequentially labeled NaCl-P, NaBD-P, NaOH-P, HClP, and residual-P. The resulting NaOH extract was digested with potassium persulfate (K2 S2 O8 ) to enable measurement of the total P extracted by NaOH. The difference between total P and SRP in the NaOH extract was referred to as the organic-P in the sediment. All P fractionation extracts and SRP in the overlying water were analyzed with a UV–vis spectrophotometer according to the molybdenum blue method [40].

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Fig. 2. Oxygen uptake rate (OUR, (a)) and soluble reactive phosphorus (SRP, (b)) flux across the sediment–water interface (SWI) in microcosms after the addition of tubificid worms and before the addition of algae. Results are expressed as the mean ± 1 S.D. (n = 3). For SRP fluxes, positive and negative values indicate the release of SRP from the sediment to the overlying water and the adsorption of SRP from the overlying water to the sediment, respectively. Values with different letters are significantly different at the p < 0.05 level (Tukey post hoc tests). C = control; A = algae addition; A + T = algae and tubificid worm addition.

3. Results 3.1. General observations Tubificid worms gradually burrowed into the sediment after being introduced into the A + T treatment microcosms. They excreted fecal pellets onto the sediment surface throughout the experiment. The added algae floated on the overlying water surface, whereas flocculation was produced immediately after the addition of chitosan and mud, which quickly deposited onto the sediment surface. The dispersal and sedimentation of mud compressed the deposited algal flocculation layer. After 3 h, the overlying water appeared clear. During the incubation experiment, the water temperature decreased from 28.0 ◦ C on day 0 to 17.2 ◦ C on day 80 as the outdoor temperature changed (Fig. 1). At the end of the experiment, the average surviving worms in treatments C and A were 26 (37 and 15) and 45 (36 and 54) individuals per microcosm, despite

no worms being added to these two groups. In treatment A + T, the mean survival in one microcosm was only 186 (166 and 205). 3.2. O2 uptake rate and SRP flux Prior to the addition of tubificid worms, the O2 uptake rate ranged from 542 ± 117 to 563 ± 34 ␮mol m−2 h−1 in the three treatments, and no significant difference was detected (p > 0.05). Moreover, the SRP flux, which varied from 1.57 ± 0.76 to 2.10 ± 0.50 ␮mol m−2 h−1 , was similar among the three treatments (p > 0.05). The presence of tubificid worms increased the sediment O2 uptake rate (p < 0.001) and caused SRP flux from the overlying water to the sediment (p < 0.001) (Fig. 2). The deposition of algal flocculation increased the sediment O2 uptake rate (p < 0.001) (Fig. 3(a)). The O2 uptake rate was increased

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(a)

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day 9 to day 44, and the two treatments became similar on day 79. Compared with treatment A, worm resulted in a higher Fe2+ concentration in approximately the top 1 cm of sediment on day 9. Then, the Fe2+ in the pore water of treatment A + T was generally less than that of treatment A. Similar to Fe2+ changes, the SRP was higher in treatment A than that in treatment C until day 79. Worms did not enhance the SRP in pore-water on day 9 compared with that in treatment A. From day 16, worms increased the pore-water SRP noticeably.

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The deposition of algal flocculation increased the concentration of chl a in the surface sediments in treatments A and A + T (Fig. 6). The chl a concentration decreased rapidly from day 9 to day 30 and thereafter decreased less rapidly. On day 79, a small amount of chl a was preserved in the sediment. Tubificid worms delayed the decomposition of algae in the sediment and transferred some algae into deeper sediment compared with treatment A.

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Algal flocculation and tubificid worms increased the NaCl-P in sediment (p < 0.001, Fig. 7). More specifically, treatments A and A + T exhibited significantly increased NaCl-P compared with treatment C (p < 0.05); treatment A + T caused higher NaCl-P than that in treatment A (p < 0.05) (Tukey post hoc test). In addition, the NaBD-P was significantly influenced by algal flocculation and by the addition of tubificid worms (p < 0.05). In contrast, the Tukey post hoc test indicated no significant difference among the three treatments (p > 0.05). The remaining four P fractions – NaOH-P, organic-P, HClP, and residual-P – were not significantly different among the three treatments (p > 0.05).

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Time (d) Fig. 3. Oxygen uptake rate (OUR, (a)) and soluble reactive phosphorus (SRP, (b)) flux across the SWI after the deposition of algae flocculation. Results are expressed as the mean ± 1 S.D. (n = 3). For SRP fluxes, positive and negative values indicate the release of SRP from the sediment to the overlying water and the adsorption of SRP from the overlying water to the sediment, respectively. C = control; A = algae addition; A + T = algae and tubificid worm addition.

in treatments A and A + T compared with treatment C (p < 0.05), whereas treatment A exhibited an O2 uptake rate similar to that of treatment A + T (p > 0.05) (Tukey post hoc tests). The SRP flux was also significantly influenced by the deposition of algal flocculation (p < 0.001) (Fig. 3(b)). Specifically, the SRP flux in treatment A changed significantly in comparison to that in treatment C (p < 0.05), whereas the SRP flux in treatment A + T was similar to that in treatment C (p > 0.05) (Tukey post hoc tests). 3.3. O2 and pH profiles Before algal flocculation, the O2 and pH profiles of treatments C and A were similar, whereas tubificid worms raised pH level in sediment, and altered O2 profiles in treatment A + T (Fig. 4). The deposition of algal flocculation decreased the O2 penetration depth in treatments A and A + T, which was strong in the first half of the experiment and weakened after day 42. In addition, worms changed the O2 profile compared with that in treatment A; however, this change varied throughout the experiment. In general, algae deposition decreased the pH in approximately the top 10 mm of sediment and elevated the pH in the deeper sediment. In contrast to treatment A, worms raised pH level in most times of the experiment. 3.4. Pore-water profiles The deposition of algal flocculation and tubificid worms altered the Fe2+ and SRP pore-water profiles (Fig. 5). The overall pore-water Fe2+ concentration in treatment A was greater than that in C from

4. Discussion 4.1. Effects of algal flocculation deposition Oxygen is the most favorable electron acceptor available during the degradation of organic matter, and it is quickly depleted in sediment rich in organic matter [23,28]. The addition of frozen diatoms and pulse input of labile material have been reported to enhance the benthic O2 uptake rate [15,43]. Thus, the degradation of pulsed organic matter is an important reason for the O2 uptake rate increase in treatment A (Fig. 3), which is confirmed by the decrease in the chl a concentration in sediment over time (Fig. 6). In contrast, the presence of excess chl a in treatment A, compared with treatment C, may demonstrate the existence of live algae, which also consume O2 and contribute to the enhancement of the O2 uptake rate. The rapid consumption of O2 by deposited algal flocculation strongly decreased the O2 penetration depth in the first half of the experiment, and the O2 penetration depth then was prone to recover as the algae degradation became slowly in the last half of the experiment (Fig. 4). The degradation of reactive organic matter produces some organic acids (such as acetic acid), which would cause a reduction in pH [44,45]. In contrast, an increase in organic matter on a sediment surface accelerates sulfate reduction and increases the H2 S concentration in the sediment [45]. Thus, we observed a significant decrease in pH to depths of approximately 10 mm (Fig. 4), and the most significant decrease occurred on day 7, when the highest O2 uptake rate was observed (Fig. 3). Accompany the decrease in O2 penetration depth and pH, the Fe2+ concentration in pore water increased with the deposition of flocculated algae (Fig. 5, upper panel), which reflects the decrease of redox potential in the sediment.

L. Zhang et al. / Journal of Hazardous Materials 266 (2014) 1–9

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0 5 10 15 20 25 30 Fig. 4. Oxygen and pH profiles across the SWI at different times. C = control; A = algae addition; A + T = algae and tubificid worm addition. SWI = sediment–water interface.

P bound to hydrous ferric oxides is considered reducible P, which may change with redox conditions, and increases in the Fe2+ concentration in sediment have been noted to frequently be accompanied by increases in the SRP [46]. The SRP in the pore water of treatment A increased in a manner similar to the increase in the Fe2+ concentration (Fig. 5). The degradation of algae can produce SRP [47], which supports the observed increase in pore-water SRP in treatment A. Algal flocculation increased the NaCl-P in the sediment (Fig. 7), which is the loosely adsorbed P on sediment particles and the P in pore water [42]. This result is in agreement with the pore-water SRP data. The increase in the SRP in the pore water indicates that the SRP flux from the sediment to the overlying water by molecular diffusion should be enhanced [48]. We did measure an enormous SRP release from the sediment to the overlying water on day 7, which is in accordance with the high SRP concentration around the SWI (Fig. 5). Hansen et al. [31] also observed a steep increase in P release in sediment–water microcosms on the eighth day of natural green algae sedimentation. In contrast, sediment from treatment A adsorbed more SRP than treatment C after day 14 (Fig. 3). This increased adsorption may be due to the effect of Fe2+ on the sediment. The increase in the Fe2+ concentration in the top

layer of sediment would enhance the migration of Fe2+ from lower depths to the sediment surface. After Fe2+ reached the oxic sediment surface, it would oxidize and precipitate, thereby resulting in the enrichment of hydrous ferric oxides in the sediment surface. The hydrous ferric oxides can adsorb SRP and act as a barrier that inhibits SRP release from the sediment to the overlying water [27]. 4.2. Effects of tubificid worms Several studies have reported an enhancement of the O2 uptake rate by tubificid worms [29,49], which can be attributed to the worms’ respiration, their stimulation of microbial activity, and their activities that expose otherwise anoxic sediments to O2 [28]. Therefore, the O2 uptake rate in treatment A + T would be greater than that in treatment A. However, the O2 uptake rate in treatments A and A + T were similar after algae deposition (Fig. 3 (a)). This result may be attributed to the following two reasons. First, algae respiration and degradation after deposition requires a large quantity of O2 , which is supported by the O2 uptake rate results for treatment A, whereas the O2 diffusion across the sediment surface is confined. Therefore, O2 consumed by tubificid worm respiration and activity would be limited. Second, the survival of worms under

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different treatments illustrates that some worms in treatment A + T died during the experiment. The loss of worms in treatment A + T would partially decrease the O2 uptake rate. In addition, the low O2 availability for tubificid worms, caused by algae respiration and degradation, was possibly why some worms died in treatment A + T.

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L. Zhang et al. / Journal of Hazardous Materials 266 (2014) 1–9

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Fig. 7. Phosphorus fractions in different sediment treatment groups at the end of the experiment. Results are expressed as the mean ± 1 S.D. (n = 3). C = control; A = algae addition; A + T = algae and tubificid worm addition.

illustrate that tubificid worms can increase sediment pH with or without organic matter input on the sediment. After algae deposition, worms caused a decrease in the Fe2+ concentration in comparison with that in treatment A (Fig. 5), because tubificid worms can increase the redox potential of sediment [50]. However, the raise of the SRP was not in agreement with the decrease in Fe2+ concentration in treatment A + T. This result is incongruent with our previously discussed analysis, where the changes in the Fe2+ concentration and in the SRP in treatment A were similar. Hansen et al. and Tuominen et al. [16,31] observed that Chironomus plumosus larvae and Monoporeia affinis separately decreased pore-water SRP with algae sedimentation. Our NaCl-P results shown in Fig. 7 confirm our pore-water SRP result. The difference to previous studies may be due to different experimental conditions, including differences in temperature, sediment characteristics, algae biomass, and benthic animals. Benthic animals likely play an important role in the diverse findings. Under the same conditions, chironomidae larvae were observed to significantly decrease SRP in pore water, whereas tubificid worms were not [29,51]. In addition, the ingestion, egestion, and undulating movement of tubificid worms migrated algae from the sediment surface to lower depths (Fig. 6), and the input of fresh organic matter likely promotes the degradation of old organic matter [24]. Therefore, the degradation of both fresh and old organic matter would produce SRP [24,52]. Lastly, the degradation of algae produced SRP, and the undulation of tubificid worms would transfer it to the deeper sediment, which would also increase the SRP, thereby explaining the increase in SRP in treatment A + T relative to that in treatment A. The SRP flux in treatment A and A + T shared the similar direction after algae deposition, whereas statistical analyses indicated the SRP flux in treatment A + T was significantly different from that in treatment A (Fig. 3(b)). This statistical result

might result from the huge release of SRP in treatment A on day 7. 4.3. Implications for algae removal by flocculation Initially, the deposited algae by flocculation died quickly when deposited onto the sediment surface, which eventually slowed (Fig. 6). A small amount of algae might be still alive in the surface sediment after 44 days. The benthic current, the wind wave, and the benthic animals can facilitate sediment resuspension, which will again bring deposited algae into the overlying water. In the case of our experiment, Lake Taihu experiences frequent wind, which can suspend the surficial 0.40–3.8 mm of sediment into the water column [53]. When resuspension is combined with SRP release (SRP release on day 7), algae recovery in the water column accelerates. Thus, special attention should be paid to lakes with frequent resuspension when using this technology to remove algae. Tubificid worms used in our experiment are upward conveyors: they burrow into the sediment, ingest the deeper sediment, and defecate onto the sediment surface [54]. Their undulation generates advective movement of surficial sediment particles around them from the top of the sediment to the bottom of their feeding zone [49,55], which is why tubificid worms can transport algae from the sediment surface to lower depths (Fig. 6). Their activities also mitigate the degradation of algae in sediment, which will provide more opportunities for algae to return to overlying water. Tubificid worms are small in body size, and their movements are primarily vertical. The larger benthic animals with latitudinal movement, such as mollusks, will substantially destroy the sediment surface [56], thereby exposing more deposited algae to the overlying water or increasing the suspension of algae in the overlying water. Our study indicates that algae sedimentation caused an approximate 35% mortality, although tubificid worms can tolerate low

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oxygen. For benthic animals that require high oxygen levels, algae deposition may increase mortality. Therefore, the effects of algae deposition on other benthic animals require further research. In addition, the effect of algal flocculation deposition on benthic animals will need to be specifically considered if this method is applied in controlling HABs. 5. Conclusions We examined the influence of cyanobacteria deposition by flocculation on the benthic O2 and P dynamics. The impulse algae deposition decreases the O2 penetration depth in sediment and greatly increases the sediment O2 uptake rate. The decrease in the O2 penetration depth was strong at the beginning of the algae deposition but recovered gradually during the latter half of the incubation experiment. Pore-water SRP and NaCl-P in the sediment were increased by algae deposition; however, we only observed an enormous SRP release from the sediment to the overlying water 7 days after the algae deposition, and the SRP flux was adsorbed by sediment for the remainder time of the experiment. Tubificid worms transported algal cells to the deeper sediment and mitigated their degradation. In addition, the worms enhanced the increase in pore-water SRP and NaCl-P in the sediment. When considering the use of flocculation to control HABs, special attention should be paid to a large release of SRP and to an enhancement of the O2 uptake rate after algae deposition. In addition, the interaction between benthic animals and the newly deposited algae layer should be considered. Acknowledgments This work was financially supported by the National Natural Science Foundation of China (41103033), the Basic research program of Jiangsu Province (BK2011879), and Nanjing Institute of Geography and Limnology, CAS (NIGLAS2011QD09). We would like to thank Shiguang Shao, Jingge Shang for their friend helps during field sampling. References [1] H.W. Paerl, N.S. Hall, E.S. Calandrino, Controlling harmful cyanobacterial blooms in a world experiencing anthropogenic and climatic-induced change, Sci. Total Environ. 409 (2011) 1739–1745. [2] J. Heisler, P. Glibert, J.M. Burkholder, D. Anderson, W. Cochlan, W. Dennison, Q. Dortch, C. Gobler, C. Heil, E. Humphries, Eutrophication and harmful algal blooms: a scientific consensus, Harmful Algae 8 (2008) 3–13. [3] H.W. Paerl, R.S. Fulton, P.H. Moisander, J. Dyble, Harmful freshwater algal blooms, with an emphasis on cyanobacteria, Sci. World J. 1 (2001) 76–113. [4] D.M. Anderson, A.D. Cembella, G.M. Hallegraeff, Progress in understanding harmful algal blooms: paradigm shifts and new technologies for research, monitoring, and management, Annu. Rev. Mar. Sci. 4 (2012) 143–176. [5] R.H. Pierce, M.S. Henry, C.J. Higham, P. Blum, M.R. Sengco, D.M. Anderson, Removal of harmful algal cells (Karenia brevis) and toxins from seawater culture by clay flocculation, Harmful Algae 3 (2004) 141–148. [6] S. Gao, J. Yang, J. Tian, F. Ma, G. Tu, M. Du, Electro-coagulation–flotation process for algae removal, J. Hazard. Mater. 177 (2010) 336–343. [7] J.J. Chen, H.H. Yeh, The mechanisms of potassium permanganate on algae removal, Water Res. 39 (2005) 4420–4428. [8] P. Rajasekhar, L. Fan, T. Nguyen, F.A. Roddick, A review of the use of sonication to control cyanobacterial blooms, Water Res. 46 (2012) 4319–4329. [9] H. Zou, G. Pan, H. Chen, X. Yuan, Removal of cyanobacterial blooms in Taihu Lake using local soils II. Effective removal of Microcystis aeruginosa using local soils and sediments modified by chitosan, Environ. Pollut. 141 (2006) 201–205. [10] G. Liu, C. Fan, J. Zhong, L. Zhang, S. Ding, S. Yan, S. Han, Using hexadecyl trimethyl ammonium bromide (CTAB) modified clays to clean the Microcystis aeruginosa blooms in Lake Taihu, China, Harmful Algae 9 (2010) 413–418. [11] G. Pan, J. Chen, D.M. Anderson, Modified local sands for the mitigation of harmful algal blooms, Harmful Algae 10 (2011) 381–387. [12] Y. Tang, H. Zhang, X. Liu, D. Cai, H. Feng, C. Miao, X. Wang, Z. Wu, Z. Yu, Flocculation of harmful algal blooms by modified attapulgite and its safety evaluation, Water Res. 45 (2011) 2855–2862. [13] G. Pan, B. Yang, D. Wang, H. Chen, B. Tian, M. Zhang, X. Yuan, J. Chen, In-lake algal bloom removal and submerged vegetation restoration using modified local soils, Ecol. Eng. 37 (2011) 302–308.

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