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Effects of microphytobenthos Cylindrotheca closterium on the fate of di-n-butyl phthalate in an aquatic microcosm
Marine Pollution Bulletin 140 (2019) 101–106
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Marine Pollution Bulletin journal homepage: www.elsevier.com/locate/marpolbul
Effects of microphytobenthos Cylindrotheca closterium on the fate of di-nbutyl phthalate in an aquatic microcosm Fan Zhang, Zheng Ding, Haofei Gong, Jie Chi
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School of Environmental Science and Engineering, Tianjin University, Tianjin 300350, PR China
A R T I C LE I N FO
A B S T R A C T
Keywords: Microphytobenthos Cylindrotheca closterium Di‑n‑butyl phthalate (DBP) Fugacity model Bacterial community structure
Effects of Cylindrotheca closterium, a marine benthic diatom, on the fate of di-n-butyl phthalate (DBP) in a watersediment system were investigated. Model calculation results showed that DBP residue was 38.5% lower in the system with C. closterium than in the system without C. closterium. The net flux from water to sediment increased by 7.3 times in the presence of C. closterium. As a result, the total biodegradation flux of DBP in the system with C. closterium was increased by 25.6%. According to the 16 s rDNA sequencing, the presence of C. closterium decreased the bacterial population as well as bacterial community diversity in sediments. Moreover, the population of C. closterium, capable of efficiently degrading DBP, was much higher than that of the dominant DBPdegrading bacteria, demonstrating that degradation of DBP by C. closterium should be the main reason for the degradation enhancement in sediments.
1. Introduction Phthalic acid esters (PAEs) are used as plasticizers in polyvinyl chloride products and are concerned environmental endocrine-disrupting compounds (Gao et al., 2018). Because of their large usage in personal care and plastic manufacturing products as well as the potential to leach out from these products, PAEs have been found in different environmental media (Huang et al., 2008; He et al., 2015; Gao and Wen, 2016). Because the offshore area receives the runoff and wastewater from inland and coastal human activities, it has become one of the most affected environments. Different contents of PAEs have been detected in coastal environments (Net et al., 2015). Once into costal water bodies, PAEs are transformed by various processes in the environment. Microbial degradation is considered to be the main process affecting the fate of PAEs in environments (Chang et al., 2007; Zhang et al., 2018). Except for fungi and bacteria, microalgae, as photosynthetic autotrophs, play a key role in the degradation of organic pollutants. For example, through photosynthesis, they produce oxygen, supporting the degradation of organic pollutants by heterotrophic bacteria (de Godos et al., 2010); they can also directly degrade organic pollutants, such as hydrocarbons, petroleum, biocides, polyaromatic hydrocarbons, polychlorinated biphenyls and PAEs (Subashchandrabose et al., 2013; Gao and Chi, 2015). The benthic microalgae grow within the upper several millimeters of illuminated sediments. In the coastal ecosystem, biomass of the
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planktonic microalgae in the overlying waters is often lower than that of benthic microalgae (MacIntyre and Cullen, 1996; Pinckney, 2018). As the main primary producer, marine benthic microalgae are an important component in nutrient cycles (Ask et al., 2016). Yamamoto et al. (2008) observed that Nitzschia sp. could increase the redox potential as well as decomposition of organic matter in sediments as a result of algal oxygenation. The algal assemblages on sediment surface can also develop benthic biofilms, a matrix of cells, sediments and extracellular polymeric substances (EPS), which create a complex microhabitat and act to stabilize sediments (Brouwer et al., 2003; Xiao and Zheng, 2016). In addition, benthic microalgae were found to degrade organic pollutants efficiently. For example, Gao and Chi (2015) reported that di-n-butyl phthalate (DBP) was degraded more quickly by marine benthic microalga Cylindrotheca closterium than by marine planktonic microalgae Dunaliella salina and Chaetoceros muelleri. The fate of organic pollutants in the coastal ecosystem is controlled by a series of processes such as distribution, transport and transformation between and within different media (e.g. water and sediment). Therefore, understanding of these processes is crucial to evaluate the role of marine benthic microalgae in affecting the fate of organic pollutants in coastal environment. However, to our knowledge, there are no relevant reports so far. C. closterium is a widespread marine benthic microalga (Moreno-Garrido et al., 2003). DBP was the predominant PAE in the environment. In this study, DBP was selected as the typical PAE. Experiments in microcosms with or without C. closterium were
Corresponding author. E-mail address: [email protected] (J. Chi).
https://doi.org/10.1016/j.marpolbul.2019.01.033 Received 26 November 2018; Received in revised form 16 January 2019; Accepted 16 January 2019 0025-326X/ © 2019 Elsevier Ltd. All rights reserved.
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The cylinders were replenished with artificial seawater to a total of about 800 mL. DBP-containing (2.06 mg/L on average) artificial seawater was slowly delivered into the cylinder at a rate of approximately 14.5 mL/h and the overlying water on the top of the sediment was replaced about once per 54 h. The initial DBP concentrations in water and sediment phase were measured just before inflow water passed through the cylinders. The sampling time was at 0, 24, 48, 72, 96, 120, 144, 168 and 192 h for water, 0, 24, 72, 120 and 192 h for sediment, respectively. The experiments were triplicated. Water samples were collected from a depth of 5 cm. The sediment sample was collected each time by taking out an aluminum foil cup previously imbedded in the sediment, and then was stored at −20 °C after freeze drying for 24 h.
developed to study its effects on the fate of DBP in the aquatic system. The changes of microbial community structures in the sediment were analyzed by 16 s rDNA sequencing. Moreover, a level IV multimedia fugacity model was built and applied to the simulation experiment to explore the mechanism of C. closterium in affecting the fate of DBP in the simulation system. 2. Material and methods 2.1. Materials and chemicals DBP (99% purity) was purchased from Sigma. Analytical grade solvents were obtained from Tianjin Chemical Reagent Factory. Dichloromethane was rectified before the experiment. Surface sediments were collected from the intertidal flats of Yongding New River Estuary, Tianjin, China (39°5′41″N, 117°43′36″E). After being air-dried, the sediments were ground, and then passed through a 2-mm sieve to remove small rocks and debris. The basic properties of the sediment samples were as follows: pH 7.45, total organic content 0.34%, and composition of 48.2% sand, 27.4% silt and 24.4% clay. C. closterium was obtained from the Institute of Oceanology of the Chinese Academy of Sciences. Algal cells in the exponential phase of growth were used in the experiment. Details of the algal cultivation procedures and conditions are shown in the Supplementary material.
2.3. Analytical methods 2.3.1. Analysis of DBP in water and sediments Ten milliliters of each water sample were extracted three times with 3 mL of dichloromethane and each for 10 min. The dichloromethane layers were combined, and then concentrated to 1 mL under a gentle stream of nitrogen gas for GC-FID analysis. Samples of sediments were freeze-dried, ground and passed through an 80-mesh sieve. One gram of each dried sample was sonically extracted with 3 mL dichloromethane. Each sample was extracted three times and each for 15 min. After 8 min of centrifugation at 4000 rpm, the extracts were combined and concentrated to a final volume of 1 mL for GC-FID analysis. The concentrated extracts were analyzed for DBP by an Agilent 6890 N gas chromatograph fitted with a splitless injector, a fused-silica capillary column (HP-5, 0.32 μm × 30 m) and a flame ionization detector. The temperatures of injector and detector were both set at 250 °C. Nitrogen was used as a carrier gas at a flow rate of 45 mL/min. The flow rates of hydrogen and air were 40 and 450 mL/min. Injection volume was 1 μL. DBP was eluted with the following temperature program: 100 °C (1 min) → 30 °C/min (6 min) → 280 °C (5 min). The retention time of DBP was 6.91 min. Average recoveries of DBP in samples of water and sediments were higher than 99.1% and 82.4%, respectively. Relative standard deviations (RSD) were both 9% for samples of water and sediments.
2.2. Simulation experiment Two microcosm flux experiments with and without C. closterium were performed using glass cylinders (diameter 10 cm, height 15 cm) with continuously flowing artificial seawater over the top of the sediment (Fig. 1). Previous study showed that C. closterium mainly distributed on the surface sediments within 0.5 cm and also showed more significantly effect on DBP degradation in the surface sediments (top 0.5 cm) than in the bottom layer of sediment (Li et al., 2015). Accordingly, in this work, 48 g of dry sediment was evenly added on the bottom of each cylinder to obtain 0.5 cm of sediment layer. Artificial seawater (about 150 mL) filtered by 0.45 μm cellulose acetate membrane was slowly delivered into each cylinder. Before starting the experiment, all of the cylinders were transferred in the intelligent illumination incubator under the same conditions as that of algal cultivation for 7 days to restore sediment microbial diversity. For the cylinders with C. closterium, C. closterium was inoculated on the surface of sediment at an initial biomass of 2 × 105 cell/cm2 after a 3-day restoration.
2.3.2. DNA extraction and sequencing DNA extraction was carried out using a TIANamp Bacteria DNA Kit (TianGen DP302) according to the manufacturer's instructions. The V3V4 variable region of 16S rDNA gene was amplified using two bacterial primers (341F, 5′-CCTAYGGGRBGCASCAG-3′; 806R, 5′-GGACTACNN
Fig. 1. Schematic diagram of the microcosm and the transport and transformation processes of DBP. 102
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GGGTATCTAAT-3′). The PCR temperature program was as follows: initial denaturation at 94 °C for 5 min, followed by 35 cycles of 45 s at 94 °C, 40 s at 55 °C, 40 s at 72 °C and final elongation with 7 min at 72 °C. The PCR products were detected by 2% agarose gel electrophoresis and purified using a GeneJET gel extraction kit (Thermo Scientific) according to the manufacturer's instructions. The purified products were used to construct the library with Ion Plus Fragment Library Kit (Thermo Fisher) and sequenced with IonS5™XL at Novogene (Tianjin, China). Subsequently, raw sequences were demultiplexed, quality filtered using Cutadapt (V1.9.1) and compared with species annotation database by UCHIME Algorithm to detect the chimera sequence. The quality filtered sequences were clustered into operational taxonomic units (OTUs) by 97% thresholds using U parse (V7.0.1001). Alpha diversity (the Shannon-Weaver and the Simpson indices) of the sequences were analyzed using QIIME (V1.9.1) and R (V2.15.3). 2.3.3. Other analysis Subsamples were used to analyze physicochemical properties of water (total nitrogen (TN) and pH) and sediment (chlorophyll, total organic carbon (TOC), protein and polysaccharide) at the beginning and the end of the experiments. Chlorophyll was determined by the method of Hagy et al. (2005). TN, pH, and TOC were analyzed according to the method of Nanjing Institute of Soil Science of Chinese Academy of Sciences (1978). Protein and polysaccharide were analyzed according to the method of Frφlund et al. (1995) and Gerhardt et al. (1994), respectively.
Fig. 2. Predicted (lines) vs. measured (squares) concentrations of DBP in water of microcosms with (dashed line, open squares) and without (solid lines, closed squares) C. closterium as a function of time.
polysaccharide and protein in sediments increase by 79% and 15%, respectively, suggesting the formation of an algal mat. DBP concentrations in water are shown in Fig. 2. Once discharged into the microcosm, DBP dispersed into the sediment and water phases immediately. During the first 24 h, DBP concentration increased quickly from 0.113 mg/L to 0.244 and 0.312 mg/L in microcosms with and without C. closterium, respectively. Thereafter, DBP concentration fluctuated between 0.244 and 0.269 mg/L in microcosm with C. closterium except that at 48 h. DBP concentration in microcosm without C. closterium remained relatively stable until the 120th hour, and then increased to 0.489 mg/L and fluctuated between 0.428 and 0.489 mg/L. During the experiment, the average concentration of DBP in water was 28.5% lower in microcosm with C. closterium than in microcosm without C. closterium. DBP concentrations in sediments are shown in Fig. 3. Concentrations of DBP in microcosms with and without C. closterium reached the maxima of 1.76 and 3.56 mg/kg at 24 h, respectively. The reason is that the flow-in direction of DBP into the microcosm was down-in and upout, resulting accumulation of DBP in the bottom water layer during the initial period of the experiment. Adsorption of DBP to sediments is a rapid process, leading to the quick increase of DBP in sediments as observed. Thereafter, gentle mixing of DBP in the overlaying water was regularly carried out at each sampling time of the water phase, which brought down DBP concentration in water. Subsequently, DBP concentration in microcosms with and without C. closterium decreased gradually and remained relatively stable at 0.622–0.784 and 1.062–1.117 mg/kg after 120 h, respectively, mainly as a result of adsorption of DBP to sediment followed by degradation in sediments. During the experiment, the average concentration of DBP in sediments was 39.6% lower in microcosm with C. closterium than in microcosm without C. closterium.
2.4. Model development and parameters Based on the fugacity approach as described by Mackay (2001), a two-compartment (including water and sediments) level IV model was developed. Fig. 1 presents a schematic of the system and fate processes included in the model. Details of the model formulation and parameter definitions are given in the Supplementary Material. The input parameters of physical-chemical properties for DBP and the physical environment are listed in Tables S2 and S3 (Supplementary Material), respectively. The parameters were either obtained from the literatures or determined in our laboratory. 3. Results and discussion 3.1. Algal growth and DBP concentrations During the experiment, the chlorophyll content in sediments with C. closterium increased by 4.6 times (Table 1). Obvious yellow-green algal mat was observed on the surface of sediments. Studies showed that algal mat was comprised of a complex mixture of polymeric compounds (i.e. EPS) largely secreted by marine benthic diatoms on the diatomdominated intertidal mudflat (Pierre et al., 2014; Delattre et al., 2016). EPS are rich in polysaccharides, proteins, proteoglycans, lipids and many other compounds, of which carbohydrates are the major component, providing nutrients and energy for bacteria and fungi (Middelburg et al., 2000; Sun et al., 2018). Proteins play a key role in the attachment of a microbe to a surface and also contribute to the binding strength within the developing EPS matrix (Lubarsky et al., 2010; Ding et al., 2015). As shown in Table 1, contents of Table 1 Physicochemical characteristics of water and sediment phases. Treatment
Background At the end
C NC
pH of water 7.45 7.26 6.87
TN of water (mg/L) 1.10 0.12 0.35
Chlorophyll in sediments (μg/g)
TOC in sediments (%)
2.6 6.8 1.7
0.34 0.48 0.42
103
Protein in sediments (mg/ kg)
Polysaccharide in sediments (mg/ kg)
69.80 65.48 57.05
44.24 73.13 40.86
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respectively. Calculation of the sequence similarity (97%) showed that a large portion (51%) of the obtained sequences from sediments with C. closterium was assigned to Bacillariophyta. In order to confirm the species of Bacillariophyta, 16S rDNA sequence of C. closterium was determined. Based on the sequence similarity analysis (99%), the species of Bacillariophyta was identified as C. closterium. After removing the sequences belonging to C. closterium, 25,061 of taxon reads were remained. In addition, the bacterial sequences from sediments with and without C. closterium were clustered into 291 and 330 OTUs, respectively. The Shannon-Weaver and the Simpson indices were lower in sediments with C. closterium (3.55 and 0.70) than in sediments without C. closterium (5.42 and 0.95). The results indicated that the presence of C. closterium decreased the bacterial population as well as bacterial community diversity. As shown in Table 2, the dominant bacterial phyla in sediments with or without C. closterium include Bacteroidetes, Proteobacteria, Firmicutes, Saccharibacteria and Actinobacteria. The five dominant phyla accounted for > 88% of the sequences. Among them, Proteobacteria and Firmicutes were considered to play a key role in PAE degradation (Liang et al., 2008; Gao and Wen, 2016). At the genus level, significant effect of C. closterium on bacterial community structure was observed (Table 3). Taxon read number of the dominant genera was lower in sediments with C. closterium (11468) than in sediments without C. closterium (26162). The relative abundances of the dominant genera were obviously different between the sediments with and without C. closterium. For example, both the taxon read numbers and relative abundances of Salinimicrobium, Phaeobacter, Methylophaga, Bacillus, Caminicella, Pseudomonas and Nocardioides were lower in sediments with C. closterium than in sediments without C. closterium, whilst the other four dominant genera (i.e. Hoeflea, Oceanicaulis, Cohaesibacter and Aestuariibacter) were higher. Among the dominant genera, total taxon read numbers of aerobic and anaerobic bacteria were 3198 and 8270 in sediments with C. closterium, and 4231 and 21,931 in sediments without C. closterium, respectively. It can be seen that the population of anaerobic bacteria was decreased by 62% by C. closterium, which was more significant than aerobic bacteria (24%). The possible reason is that C. closterium inhibited the growth of anaerobic bacteria by increasing dissolved oxygen content in sediments. Obvious increase of redox potential in surface sediments with C. closterium was observed in this work (data not shown). A similar phenomenon was also reported in previous studies (Yamamoto et al., 2008; Denis et al., 2009). Pseudomonas and Bacillus, the dominant genera in this work, have been shown to be able to use DBP as their sole carbon source (Gao and Wen, 2016). The presence of C. closterium decreased the populations of Pseudomonas and Bacillus by 55% and 72%, respectively. There have been no studies reporting the rest of the dominant genera capable of degrading PAEs so far. Algae have been reported to have the ability to degrade PAEs. For example, the half-lives of DBP were 1.7, 8.2 and 8.5 d by C. closterium, D. salina and C. muelleri (Gao and Chi, 2015), 13.8 d by Chlorella vulgaris (Chi et al., 2006), 2.2 d by Chlorella pyrenoidosa (Yan et al., 1995), and 5.1 d by Scenedesmus obliquus (Sun, 2003), respectively. Generally, aerobic degradation of PAEs is much higher than anaerobic degradation (Staples et al., 1997; Gao and Wen, 2016). The abilities of degrading DBP by algae are comparable to those under aerobic condition (half-lives ranging from1.60 to 10.0 d) (Staples et al., 1997; Gao and Wen, 2016). In this work, the taxon read number of C. closterium was much higher than those of the dominant DBP-degrading bacteria (i.e. Pseudomonas and Bacillus). Therefore, it can be concluded that degradation of DBP by C. closterium should be the main reason leading to the degradation enhancement of DBP in sediments.
Fig. 3. Predicted (lines) vs. measured (squares) concentrations of DBP in sediments of microcosms with (dashed line, open squares) and without (solid lines, closed squares) C. closterium as a function of time.
3.2. Model application Model predicted concentrations of DBP are shown in Figs. 2 and 3. It can be seen that the model results fitted well with the experimental data of microcosms with and without C. closterium except those before 120 h in sediments. The difference between measured and calculated concentrations of DBP in sediments before 120 h is mainly because it was assumed in the model that DBP was evenly distributed in the water phase once it was discharged into the microcosm. However, mixing of DBP in the overlaying water was regularly carried out at each sampling time of the water phase. Consequently, accumulation of DBP in the bottom water layer at the initial period resulted in the measured concentrations of DBP in sediments higher than predicted. The steady-state mass balance of DBP in the microcosms with and without C. closterium is shown in Fig. 4. In the microcosm without C. closterium, 40.6% of loaded DBP was removed by advection. Biodegradation of DBP in water and sediment phase was 54.6% and 4.85%, respectively. 80.2% of retained DBP was in the water phase, the rest was in sediment phase. In the microcosm with C. closterium, 25.5% of loaded DBP was removed by advection, 34.3% and 40.4% were by biodegradation in the water and sediment phases, respectively. 81.1% of retained DBP was in the water phase, the rest was in sediment phase. It can be seen that the presence of C. closterium increased the biodegradation flux of DBP in the microcosm by 25.6% and decreased the total residue of DBP in the microcosm by 38.5%. At the same time, advective outflow decreased. The model results showed that fugacity of DBP in all microcosms was in the order of water phase > sediment phase (Fig. 4), indicating that the direction of DBP flux in the microcosms was from water to sediment. In the presence of C. closterium, the net flux from water to sediment increased by 7.3 times. Meanwhile, the presence of C. closterium enhanced the degradation rate constant of DBP in the sediment phase by 75% (0.0112 1/h with C. closterium and 0.0064 1/h without C. closterium), but showed no effect on the degradation rate constant of DBP in the water phase. Therefore, the driving force for the transport enhancement of DBP from water to sediment should be more rapid degradation of DBP in sediments. 3.3. Bacterial community diversity and structure in sediments
4. Conclusions According to the sequencing of 16 s rDNA, the filtered sequences were used for OTUs analysis. In total, 51,160 and 48,844 of taxon reads were obtained from sediments with and without C. closterium,
In this work, a level IV fugacity model was used to study the effects of marine benthic diatom C. closterium on the fate of DBP in a water104
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Fig. 4. Steady-state mass balance of DBP in microcosms with (a) and without (b) C. closterium. Table 2 Taxon read numbers and relative abundances of the top five phyla. Phylum
Bacteroidetes Proteobacteria Firmicutes Saccharibacteria Actinobacteria
Table 3 Taxon read numbers and relative abundances of the dominant generaa.
Taxon read number (relative abundance) Without C. closterium
With C. closterium
23,286 (47.7%) 11,585 (23.7%) 5533 (11.3%) 2047 (4.2%) 718 (1.4%)
13,382 (53.4%) 7649 (30.5%) 1541 (6.1%) 225 (0.9%) 362 (1.4%)
Genus
Phylum
Taxon read number (relative abundance) Without C. closterium
Salinimicrobium (anaerobic) Phaeobacter(anaerobic) Methylophaga(aerobic) Bacillus(anaerobic) Caminicella(anaerobic) Pseudomonas(aerobic) Nocardioides(aerobic) Hoeflea(aerobic) Oceanicaulis(aerobic) Cohaesibacter(anaerobic) Aestuariibacter (aerobic)
sediment system. The model calculation results showed that the presence of C. closterium obviously reduced the accumulation of DBP in the experiment system and enhanced the degradation, which are due to the enhanced biodegradation of DBP in sediments. Furthermore, 16s rDNA sequence analysis indicated that the presence of C. closterium decreased the bacterial population as well as bacterial community diversity in sediments. The population of C. closterium, capable of efficiently degrading DBP, was much higher than that of the dominant DBP-degrading bacteria, suggesting that degradation of DBP by C. closterium should be the main reason for the degradation enhancement in sediments. Algae have been reported to degrade a wide range of organic pollutants. Therefore, it appeared that benthic microalgae (e.g. C. closterium) had a potential for the remediation of coastal water bodies impacted by organic pollutants. So far, investigation on the degradation of organic pollutants by benthic microalgae has been rarely reported. Further research on this area is needed.
a b
With C. closterium
Bacteroidetes
14,997 (30.7%)
6420 (25.6%)
Proteobacteria Proteobacteria Firmicutes Firmicutes Proteobacteria Actinobacteria Proteobacteria Proteobacteria Proteobacteria Proteobacteria
4749 (9.7%) 3033 (6.2%) 1504 (3.1%) 681 (1.4%) 675 (1.4%) 523 (1.1%) < 1% < 1% –b –
898 (3.6%) 662 (2.6%) 415 (1.7%) < 1% 303 (1.2%) < 1% 618 (2.5%) 1253 (5.0%) 537 (2.1%) 362 (1.4%)
Genera of all samples with > 1% abundance are listed. –, non-detected.
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Marine Pollution Bulletin journal homepage: www.elsevier.com/locate/marpolbul
Effects of microphytobenthos Cylindrotheca closterium on the fate of di-nbutyl phthalate in an aquatic microcosm Fan Zhang, Zheng Ding, Haofei Gong, Jie Chi
T
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School of Environmental Science and Engineering, Tianjin University, Tianjin 300350, PR China
A R T I C LE I N FO
A B S T R A C T
Keywords: Microphytobenthos Cylindrotheca closterium Di‑n‑butyl phthalate (DBP) Fugacity model Bacterial community structure
Effects of Cylindrotheca closterium, a marine benthic diatom, on the fate of di-n-butyl phthalate (DBP) in a watersediment system were investigated. Model calculation results showed that DBP residue was 38.5% lower in the system with C. closterium than in the system without C. closterium. The net flux from water to sediment increased by 7.3 times in the presence of C. closterium. As a result, the total biodegradation flux of DBP in the system with C. closterium was increased by 25.6%. According to the 16 s rDNA sequencing, the presence of C. closterium decreased the bacterial population as well as bacterial community diversity in sediments. Moreover, the population of C. closterium, capable of efficiently degrading DBP, was much higher than that of the dominant DBPdegrading bacteria, demonstrating that degradation of DBP by C. closterium should be the main reason for the degradation enhancement in sediments.
1. Introduction Phthalic acid esters (PAEs) are used as plasticizers in polyvinyl chloride products and are concerned environmental endocrine-disrupting compounds (Gao et al., 2018). Because of their large usage in personal care and plastic manufacturing products as well as the potential to leach out from these products, PAEs have been found in different environmental media (Huang et al., 2008; He et al., 2015; Gao and Wen, 2016). Because the offshore area receives the runoff and wastewater from inland and coastal human activities, it has become one of the most affected environments. Different contents of PAEs have been detected in coastal environments (Net et al., 2015). Once into costal water bodies, PAEs are transformed by various processes in the environment. Microbial degradation is considered to be the main process affecting the fate of PAEs in environments (Chang et al., 2007; Zhang et al., 2018). Except for fungi and bacteria, microalgae, as photosynthetic autotrophs, play a key role in the degradation of organic pollutants. For example, through photosynthesis, they produce oxygen, supporting the degradation of organic pollutants by heterotrophic bacteria (de Godos et al., 2010); they can also directly degrade organic pollutants, such as hydrocarbons, petroleum, biocides, polyaromatic hydrocarbons, polychlorinated biphenyls and PAEs (Subashchandrabose et al., 2013; Gao and Chi, 2015). The benthic microalgae grow within the upper several millimeters of illuminated sediments. In the coastal ecosystem, biomass of the
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planktonic microalgae in the overlying waters is often lower than that of benthic microalgae (MacIntyre and Cullen, 1996; Pinckney, 2018). As the main primary producer, marine benthic microalgae are an important component in nutrient cycles (Ask et al., 2016). Yamamoto et al. (2008) observed that Nitzschia sp. could increase the redox potential as well as decomposition of organic matter in sediments as a result of algal oxygenation. The algal assemblages on sediment surface can also develop benthic biofilms, a matrix of cells, sediments and extracellular polymeric substances (EPS), which create a complex microhabitat and act to stabilize sediments (Brouwer et al., 2003; Xiao and Zheng, 2016). In addition, benthic microalgae were found to degrade organic pollutants efficiently. For example, Gao and Chi (2015) reported that di-n-butyl phthalate (DBP) was degraded more quickly by marine benthic microalga Cylindrotheca closterium than by marine planktonic microalgae Dunaliella salina and Chaetoceros muelleri. The fate of organic pollutants in the coastal ecosystem is controlled by a series of processes such as distribution, transport and transformation between and within different media (e.g. water and sediment). Therefore, understanding of these processes is crucial to evaluate the role of marine benthic microalgae in affecting the fate of organic pollutants in coastal environment. However, to our knowledge, there are no relevant reports so far. C. closterium is a widespread marine benthic microalga (Moreno-Garrido et al., 2003). DBP was the predominant PAE in the environment. In this study, DBP was selected as the typical PAE. Experiments in microcosms with or without C. closterium were
Corresponding author. E-mail address: [email protected] (J. Chi).
https://doi.org/10.1016/j.marpolbul.2019.01.033 Received 26 November 2018; Received in revised form 16 January 2019; Accepted 16 January 2019 0025-326X/ © 2019 Elsevier Ltd. All rights reserved.
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The cylinders were replenished with artificial seawater to a total of about 800 mL. DBP-containing (2.06 mg/L on average) artificial seawater was slowly delivered into the cylinder at a rate of approximately 14.5 mL/h and the overlying water on the top of the sediment was replaced about once per 54 h. The initial DBP concentrations in water and sediment phase were measured just before inflow water passed through the cylinders. The sampling time was at 0, 24, 48, 72, 96, 120, 144, 168 and 192 h for water, 0, 24, 72, 120 and 192 h for sediment, respectively. The experiments were triplicated. Water samples were collected from a depth of 5 cm. The sediment sample was collected each time by taking out an aluminum foil cup previously imbedded in the sediment, and then was stored at −20 °C after freeze drying for 24 h.
developed to study its effects on the fate of DBP in the aquatic system. The changes of microbial community structures in the sediment were analyzed by 16 s rDNA sequencing. Moreover, a level IV multimedia fugacity model was built and applied to the simulation experiment to explore the mechanism of C. closterium in affecting the fate of DBP in the simulation system. 2. Material and methods 2.1. Materials and chemicals DBP (99% purity) was purchased from Sigma. Analytical grade solvents were obtained from Tianjin Chemical Reagent Factory. Dichloromethane was rectified before the experiment. Surface sediments were collected from the intertidal flats of Yongding New River Estuary, Tianjin, China (39°5′41″N, 117°43′36″E). After being air-dried, the sediments were ground, and then passed through a 2-mm sieve to remove small rocks and debris. The basic properties of the sediment samples were as follows: pH 7.45, total organic content 0.34%, and composition of 48.2% sand, 27.4% silt and 24.4% clay. C. closterium was obtained from the Institute of Oceanology of the Chinese Academy of Sciences. Algal cells in the exponential phase of growth were used in the experiment. Details of the algal cultivation procedures and conditions are shown in the Supplementary material.
2.3. Analytical methods 2.3.1. Analysis of DBP in water and sediments Ten milliliters of each water sample were extracted three times with 3 mL of dichloromethane and each for 10 min. The dichloromethane layers were combined, and then concentrated to 1 mL under a gentle stream of nitrogen gas for GC-FID analysis. Samples of sediments were freeze-dried, ground and passed through an 80-mesh sieve. One gram of each dried sample was sonically extracted with 3 mL dichloromethane. Each sample was extracted three times and each for 15 min. After 8 min of centrifugation at 4000 rpm, the extracts were combined and concentrated to a final volume of 1 mL for GC-FID analysis. The concentrated extracts were analyzed for DBP by an Agilent 6890 N gas chromatograph fitted with a splitless injector, a fused-silica capillary column (HP-5, 0.32 μm × 30 m) and a flame ionization detector. The temperatures of injector and detector were both set at 250 °C. Nitrogen was used as a carrier gas at a flow rate of 45 mL/min. The flow rates of hydrogen and air were 40 and 450 mL/min. Injection volume was 1 μL. DBP was eluted with the following temperature program: 100 °C (1 min) → 30 °C/min (6 min) → 280 °C (5 min). The retention time of DBP was 6.91 min. Average recoveries of DBP in samples of water and sediments were higher than 99.1% and 82.4%, respectively. Relative standard deviations (RSD) were both 9% for samples of water and sediments.
2.2. Simulation experiment Two microcosm flux experiments with and without C. closterium were performed using glass cylinders (diameter 10 cm, height 15 cm) with continuously flowing artificial seawater over the top of the sediment (Fig. 1). Previous study showed that C. closterium mainly distributed on the surface sediments within 0.5 cm and also showed more significantly effect on DBP degradation in the surface sediments (top 0.5 cm) than in the bottom layer of sediment (Li et al., 2015). Accordingly, in this work, 48 g of dry sediment was evenly added on the bottom of each cylinder to obtain 0.5 cm of sediment layer. Artificial seawater (about 150 mL) filtered by 0.45 μm cellulose acetate membrane was slowly delivered into each cylinder. Before starting the experiment, all of the cylinders were transferred in the intelligent illumination incubator under the same conditions as that of algal cultivation for 7 days to restore sediment microbial diversity. For the cylinders with C. closterium, C. closterium was inoculated on the surface of sediment at an initial biomass of 2 × 105 cell/cm2 after a 3-day restoration.
2.3.2. DNA extraction and sequencing DNA extraction was carried out using a TIANamp Bacteria DNA Kit (TianGen DP302) according to the manufacturer's instructions. The V3V4 variable region of 16S rDNA gene was amplified using two bacterial primers (341F, 5′-CCTAYGGGRBGCASCAG-3′; 806R, 5′-GGACTACNN
Fig. 1. Schematic diagram of the microcosm and the transport and transformation processes of DBP. 102
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GGGTATCTAAT-3′). The PCR temperature program was as follows: initial denaturation at 94 °C for 5 min, followed by 35 cycles of 45 s at 94 °C, 40 s at 55 °C, 40 s at 72 °C and final elongation with 7 min at 72 °C. The PCR products were detected by 2% agarose gel electrophoresis and purified using a GeneJET gel extraction kit (Thermo Scientific) according to the manufacturer's instructions. The purified products were used to construct the library with Ion Plus Fragment Library Kit (Thermo Fisher) and sequenced with IonS5™XL at Novogene (Tianjin, China). Subsequently, raw sequences were demultiplexed, quality filtered using Cutadapt (V1.9.1) and compared with species annotation database by UCHIME Algorithm to detect the chimera sequence. The quality filtered sequences were clustered into operational taxonomic units (OTUs) by 97% thresholds using U parse (V7.0.1001). Alpha diversity (the Shannon-Weaver and the Simpson indices) of the sequences were analyzed using QIIME (V1.9.1) and R (V2.15.3). 2.3.3. Other analysis Subsamples were used to analyze physicochemical properties of water (total nitrogen (TN) and pH) and sediment (chlorophyll, total organic carbon (TOC), protein and polysaccharide) at the beginning and the end of the experiments. Chlorophyll was determined by the method of Hagy et al. (2005). TN, pH, and TOC were analyzed according to the method of Nanjing Institute of Soil Science of Chinese Academy of Sciences (1978). Protein and polysaccharide were analyzed according to the method of Frφlund et al. (1995) and Gerhardt et al. (1994), respectively.
Fig. 2. Predicted (lines) vs. measured (squares) concentrations of DBP in water of microcosms with (dashed line, open squares) and without (solid lines, closed squares) C. closterium as a function of time.
polysaccharide and protein in sediments increase by 79% and 15%, respectively, suggesting the formation of an algal mat. DBP concentrations in water are shown in Fig. 2. Once discharged into the microcosm, DBP dispersed into the sediment and water phases immediately. During the first 24 h, DBP concentration increased quickly from 0.113 mg/L to 0.244 and 0.312 mg/L in microcosms with and without C. closterium, respectively. Thereafter, DBP concentration fluctuated between 0.244 and 0.269 mg/L in microcosm with C. closterium except that at 48 h. DBP concentration in microcosm without C. closterium remained relatively stable until the 120th hour, and then increased to 0.489 mg/L and fluctuated between 0.428 and 0.489 mg/L. During the experiment, the average concentration of DBP in water was 28.5% lower in microcosm with C. closterium than in microcosm without C. closterium. DBP concentrations in sediments are shown in Fig. 3. Concentrations of DBP in microcosms with and without C. closterium reached the maxima of 1.76 and 3.56 mg/kg at 24 h, respectively. The reason is that the flow-in direction of DBP into the microcosm was down-in and upout, resulting accumulation of DBP in the bottom water layer during the initial period of the experiment. Adsorption of DBP to sediments is a rapid process, leading to the quick increase of DBP in sediments as observed. Thereafter, gentle mixing of DBP in the overlaying water was regularly carried out at each sampling time of the water phase, which brought down DBP concentration in water. Subsequently, DBP concentration in microcosms with and without C. closterium decreased gradually and remained relatively stable at 0.622–0.784 and 1.062–1.117 mg/kg after 120 h, respectively, mainly as a result of adsorption of DBP to sediment followed by degradation in sediments. During the experiment, the average concentration of DBP in sediments was 39.6% lower in microcosm with C. closterium than in microcosm without C. closterium.
2.4. Model development and parameters Based on the fugacity approach as described by Mackay (2001), a two-compartment (including water and sediments) level IV model was developed. Fig. 1 presents a schematic of the system and fate processes included in the model. Details of the model formulation and parameter definitions are given in the Supplementary Material. The input parameters of physical-chemical properties for DBP and the physical environment are listed in Tables S2 and S3 (Supplementary Material), respectively. The parameters were either obtained from the literatures or determined in our laboratory. 3. Results and discussion 3.1. Algal growth and DBP concentrations During the experiment, the chlorophyll content in sediments with C. closterium increased by 4.6 times (Table 1). Obvious yellow-green algal mat was observed on the surface of sediments. Studies showed that algal mat was comprised of a complex mixture of polymeric compounds (i.e. EPS) largely secreted by marine benthic diatoms on the diatomdominated intertidal mudflat (Pierre et al., 2014; Delattre et al., 2016). EPS are rich in polysaccharides, proteins, proteoglycans, lipids and many other compounds, of which carbohydrates are the major component, providing nutrients and energy for bacteria and fungi (Middelburg et al., 2000; Sun et al., 2018). Proteins play a key role in the attachment of a microbe to a surface and also contribute to the binding strength within the developing EPS matrix (Lubarsky et al., 2010; Ding et al., 2015). As shown in Table 1, contents of Table 1 Physicochemical characteristics of water and sediment phases. Treatment
Background At the end
C NC
pH of water 7.45 7.26 6.87
TN of water (mg/L) 1.10 0.12 0.35
Chlorophyll in sediments (μg/g)
TOC in sediments (%)
2.6 6.8 1.7
0.34 0.48 0.42
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Protein in sediments (mg/ kg)
Polysaccharide in sediments (mg/ kg)
69.80 65.48 57.05
44.24 73.13 40.86
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respectively. Calculation of the sequence similarity (97%) showed that a large portion (51%) of the obtained sequences from sediments with C. closterium was assigned to Bacillariophyta. In order to confirm the species of Bacillariophyta, 16S rDNA sequence of C. closterium was determined. Based on the sequence similarity analysis (99%), the species of Bacillariophyta was identified as C. closterium. After removing the sequences belonging to C. closterium, 25,061 of taxon reads were remained. In addition, the bacterial sequences from sediments with and without C. closterium were clustered into 291 and 330 OTUs, respectively. The Shannon-Weaver and the Simpson indices were lower in sediments with C. closterium (3.55 and 0.70) than in sediments without C. closterium (5.42 and 0.95). The results indicated that the presence of C. closterium decreased the bacterial population as well as bacterial community diversity. As shown in Table 2, the dominant bacterial phyla in sediments with or without C. closterium include Bacteroidetes, Proteobacteria, Firmicutes, Saccharibacteria and Actinobacteria. The five dominant phyla accounted for > 88% of the sequences. Among them, Proteobacteria and Firmicutes were considered to play a key role in PAE degradation (Liang et al., 2008; Gao and Wen, 2016). At the genus level, significant effect of C. closterium on bacterial community structure was observed (Table 3). Taxon read number of the dominant genera was lower in sediments with C. closterium (11468) than in sediments without C. closterium (26162). The relative abundances of the dominant genera were obviously different between the sediments with and without C. closterium. For example, both the taxon read numbers and relative abundances of Salinimicrobium, Phaeobacter, Methylophaga, Bacillus, Caminicella, Pseudomonas and Nocardioides were lower in sediments with C. closterium than in sediments without C. closterium, whilst the other four dominant genera (i.e. Hoeflea, Oceanicaulis, Cohaesibacter and Aestuariibacter) were higher. Among the dominant genera, total taxon read numbers of aerobic and anaerobic bacteria were 3198 and 8270 in sediments with C. closterium, and 4231 and 21,931 in sediments without C. closterium, respectively. It can be seen that the population of anaerobic bacteria was decreased by 62% by C. closterium, which was more significant than aerobic bacteria (24%). The possible reason is that C. closterium inhibited the growth of anaerobic bacteria by increasing dissolved oxygen content in sediments. Obvious increase of redox potential in surface sediments with C. closterium was observed in this work (data not shown). A similar phenomenon was also reported in previous studies (Yamamoto et al., 2008; Denis et al., 2009). Pseudomonas and Bacillus, the dominant genera in this work, have been shown to be able to use DBP as their sole carbon source (Gao and Wen, 2016). The presence of C. closterium decreased the populations of Pseudomonas and Bacillus by 55% and 72%, respectively. There have been no studies reporting the rest of the dominant genera capable of degrading PAEs so far. Algae have been reported to have the ability to degrade PAEs. For example, the half-lives of DBP were 1.7, 8.2 and 8.5 d by C. closterium, D. salina and C. muelleri (Gao and Chi, 2015), 13.8 d by Chlorella vulgaris (Chi et al., 2006), 2.2 d by Chlorella pyrenoidosa (Yan et al., 1995), and 5.1 d by Scenedesmus obliquus (Sun, 2003), respectively. Generally, aerobic degradation of PAEs is much higher than anaerobic degradation (Staples et al., 1997; Gao and Wen, 2016). The abilities of degrading DBP by algae are comparable to those under aerobic condition (half-lives ranging from1.60 to 10.0 d) (Staples et al., 1997; Gao and Wen, 2016). In this work, the taxon read number of C. closterium was much higher than those of the dominant DBP-degrading bacteria (i.e. Pseudomonas and Bacillus). Therefore, it can be concluded that degradation of DBP by C. closterium should be the main reason leading to the degradation enhancement of DBP in sediments.
Fig. 3. Predicted (lines) vs. measured (squares) concentrations of DBP in sediments of microcosms with (dashed line, open squares) and without (solid lines, closed squares) C. closterium as a function of time.
3.2. Model application Model predicted concentrations of DBP are shown in Figs. 2 and 3. It can be seen that the model results fitted well with the experimental data of microcosms with and without C. closterium except those before 120 h in sediments. The difference between measured and calculated concentrations of DBP in sediments before 120 h is mainly because it was assumed in the model that DBP was evenly distributed in the water phase once it was discharged into the microcosm. However, mixing of DBP in the overlaying water was regularly carried out at each sampling time of the water phase. Consequently, accumulation of DBP in the bottom water layer at the initial period resulted in the measured concentrations of DBP in sediments higher than predicted. The steady-state mass balance of DBP in the microcosms with and without C. closterium is shown in Fig. 4. In the microcosm without C. closterium, 40.6% of loaded DBP was removed by advection. Biodegradation of DBP in water and sediment phase was 54.6% and 4.85%, respectively. 80.2% of retained DBP was in the water phase, the rest was in sediment phase. In the microcosm with C. closterium, 25.5% of loaded DBP was removed by advection, 34.3% and 40.4% were by biodegradation in the water and sediment phases, respectively. 81.1% of retained DBP was in the water phase, the rest was in sediment phase. It can be seen that the presence of C. closterium increased the biodegradation flux of DBP in the microcosm by 25.6% and decreased the total residue of DBP in the microcosm by 38.5%. At the same time, advective outflow decreased. The model results showed that fugacity of DBP in all microcosms was in the order of water phase > sediment phase (Fig. 4), indicating that the direction of DBP flux in the microcosms was from water to sediment. In the presence of C. closterium, the net flux from water to sediment increased by 7.3 times. Meanwhile, the presence of C. closterium enhanced the degradation rate constant of DBP in the sediment phase by 75% (0.0112 1/h with C. closterium and 0.0064 1/h without C. closterium), but showed no effect on the degradation rate constant of DBP in the water phase. Therefore, the driving force for the transport enhancement of DBP from water to sediment should be more rapid degradation of DBP in sediments. 3.3. Bacterial community diversity and structure in sediments
4. Conclusions According to the sequencing of 16 s rDNA, the filtered sequences were used for OTUs analysis. In total, 51,160 and 48,844 of taxon reads were obtained from sediments with and without C. closterium,
In this work, a level IV fugacity model was used to study the effects of marine benthic diatom C. closterium on the fate of DBP in a water104
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Fig. 4. Steady-state mass balance of DBP in microcosms with (a) and without (b) C. closterium. Table 2 Taxon read numbers and relative abundances of the top five phyla. Phylum
Bacteroidetes Proteobacteria Firmicutes Saccharibacteria Actinobacteria
Table 3 Taxon read numbers and relative abundances of the dominant generaa.
Taxon read number (relative abundance) Without C. closterium
With C. closterium
23,286 (47.7%) 11,585 (23.7%) 5533 (11.3%) 2047 (4.2%) 718 (1.4%)
13,382 (53.4%) 7649 (30.5%) 1541 (6.1%) 225 (0.9%) 362 (1.4%)
Genus
Phylum
Taxon read number (relative abundance) Without C. closterium
Salinimicrobium (anaerobic) Phaeobacter(anaerobic) Methylophaga(aerobic) Bacillus(anaerobic) Caminicella(anaerobic) Pseudomonas(aerobic) Nocardioides(aerobic) Hoeflea(aerobic) Oceanicaulis(aerobic) Cohaesibacter(anaerobic) Aestuariibacter (aerobic)
sediment system. The model calculation results showed that the presence of C. closterium obviously reduced the accumulation of DBP in the experiment system and enhanced the degradation, which are due to the enhanced biodegradation of DBP in sediments. Furthermore, 16s rDNA sequence analysis indicated that the presence of C. closterium decreased the bacterial population as well as bacterial community diversity in sediments. The population of C. closterium, capable of efficiently degrading DBP, was much higher than that of the dominant DBP-degrading bacteria, suggesting that degradation of DBP by C. closterium should be the main reason for the degradation enhancement in sediments. Algae have been reported to degrade a wide range of organic pollutants. Therefore, it appeared that benthic microalgae (e.g. C. closterium) had a potential for the remediation of coastal water bodies impacted by organic pollutants. So far, investigation on the degradation of organic pollutants by benthic microalgae has been rarely reported. Further research on this area is needed.
a b
With C. closterium
Bacteroidetes
14,997 (30.7%)
6420 (25.6%)
Proteobacteria Proteobacteria Firmicutes Firmicutes Proteobacteria Actinobacteria Proteobacteria Proteobacteria Proteobacteria Proteobacteria
4749 (9.7%) 3033 (6.2%) 1504 (3.1%) 681 (1.4%) 675 (1.4%) 523 (1.1%) < 1% < 1% –b –
898 (3.6%) 662 (2.6%) 415 (1.7%) < 1% 303 (1.2%) < 1% 618 (2.5%) 1253 (5.0%) 537 (2.1%) 362 (1.4%)
Genera of all samples with > 1% abundance are listed. –, non-detected.
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