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Impact of different environmental particles on degradation of dibutyl phthalate in coastal sediments with and without Cylindrotheca closterium
Environmental Pollution 261 (2020) 114228
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Impact of different environmental particles on degradation of dibutyl phthalate in coastal sediments with and without Cylindrotheca closterium Fan Zhang, Dongxu Zhao, Jie Chi* School of Environmental Science and Engineering, Tianjin University, Tianjin, 300350, PR China
a r t i c l e i n f o
a b s t r a c t
Article history: Received 29 August 2019 Received in revised form 3 February 2020 Accepted 16 February 2020 Available online 19 February 2020
This study investigated the impact of different environmental particles at different concentrations (0.2% and 2%, w/w) on biodegradation of dibutyl phthalate (DBP) in sediments with and without Cylindrotheca closterium, a marine benthic diatom. The particles included biochar pyrolyzed at 400 C, multi-walled carbon nanotube (MWNT), nanoscale zero-valent iron (nZVI) and polyethylene microplastic. In treatments without C. closterium, inhibition effect of the particles on degradation percentage of DBP (up to 15.7% decrement except 1.7% increment for 0.2% nZVI) increased with the increase of particle sorption ability to DBP and particle concentration in general. The results of 16s rDNA sequencing showed that C. closterium was probably the most abundant DBP-degrader, accounting for 20.0e49.3% of the total taxon read numbers. In treatments with C. closterium, inoculation of C. closterium increased the degradation percentage of DBP in all treatments with particle addition by 0.0e11.3%, which increased with the increase of chlorophyll a content in general but decreased with the increase of particle concentration from 0.2% to 2%. The increment was the highest for treatment with 0.2% nZVI addition due to its highest promotion effect on algal growth. In contrast, the increment was the lowest for treatments with MWNT addition due to its strong sorption to DBP and strong inhibition on the growth of C. closterium. Our findings suggested that the environmental particles could influence bioavailability of DBP by sorption and biomass of C. closterium, and thus degradation of DBP in sediments. © 2020 Elsevier Ltd. All rights reserved.
Keywords: Environmental particles Marine benthic diatom Dibutyl phthalate Coastal sediments Degradation
1. Introduction With the development of science and technology, different kinds of particles are released into the environment during manufacture process and usage of chemicals and industrial materials. For example, microplastics and engineering nanoparticles (e.g. carbon nanotubes and zero-valent iron) are ubiquitous in soils, sediments, rivers, lakes and oceans (Horton and Dixon, 2018; Goswami et al., 2017). In addition, biochars pyrolyzed by decomposition of biomass under limited oxygen conditions are widely applied as a soil amendment (Zhu et al., 2017). Due to small size, these environmental particles usually possess large surface area and high sorption capacity for hydrophobic organic pollutants (e.g. phthalates, polycyclic aromatic hydrocarbons and polychlorinated
* Corresponding author. School of Environmental Science and Engineering, Tianjin University, Tianjin, 300350, China. E-mail address: [email protected] (J. Chi). https://doi.org/10.1016/j.envpol.2020.114228 0269-7491/© 2020 Elsevier Ltd. All rights reserved.
biphenyls), thus resulting in accumulation of the pollutants in the environment (Glomstad et al., 2016; Beckingham and Ghosh, 2017; Jiang, 2018). For example, carbon nanotube at a concentration of 5 mg/g significantly increased the residual amount of polycyclic aromatic hydrocarbons in soils (Li et al., 2013); adding biochar to soils obviously retained polycyclic aromatic hydrocarbons et al., 2019). Besides, environmental particles may be (Rombola ingested by organisms and cause disruption to their physical ingestion, metabolism and reproductive performance (Alimba and Faggio, 2019). Pollutants sorbed on the particles might be further released in the organisms after ingestion, enhancing their toxic susceptibility (Anbumani and Kakkar, 2018). It has been reported that coastal sediments are the major sink for environmental particles, especially near the areas of high population and industrial density (Frias et al., 2016; Goswami et al., 2017). Microphytobenthos grow on the surface of sediments. They are one of the most important primary producers in coastal area and have important ecological and environmental significance. Microphytobenthos can release oxygen and soluble substances into
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environment (Ask et al., 2016), thus influencing bacterial community structure and degradation of organic pollutants (Subashchandrabose et al., 2013). Algal growth could also be impacted by environmental particles (Lambreva et al., 2015; Prata et al., 2019). Therefore, coexistence of environmental particles and microphytobenthos can impact degradation and persistence of hydrophobic organic compounds in coastal sediments. Phthalic acid esters (PAEs) are typical hydrophobic organic pollutants in coastal sediments (Gao et al., 2018). Cylindrotheca closterium is a benthic diatom commonly occurring in the marine sediments. Our previous studies found that C. closterium could degrade dibutyl phthalate (DBP) directly (Gao and Chi, 2015; Chi et al., 2019). In this work, four environmental particles were selected, including biochar pyrolyzed at 400 C (BC400), multiwalled carbon nanotube (MWNT), polyethylene microplastic (PE) and nanoscale zero-valent iron (nZVI). The aims are to investigate degradation of DBP in sediments when the particles and C. closterium singly existed or coexisted. Sorption ability of the particles for DBP was measured. The changes of C. closterium biomass and bacterial community structure in sediments were determined. Mechanisms of particle effects on biodegradation of DBP in sediments with or without C. closterium and potential contribution of each experimental compartment (particles, bacteria and the alga) to degradation of DBP were discussed. 2. Material and methods 2.1. Materials In this work, MWNT (>98 wt % purity), nZVI (99.9% purity) and PE (99% purity) were obtained from Chengdu Organic Chemicals Co. Ltd., Shanghai Macklin Biochemical Co. Ltd. and Zhonglian Plastic Technology Co. Ltd., respectively. Biochar was produced by pyrolyzing wheat straw chips at in a muffle furnace at 400 C under oxygen-limited condition. The charred solid was then sieved through a 0.178-mm sieve and named as BC400. The basic properties of the particles are given in Table S1. Surface sediments were sampled from the intertidal flats of the Bohai Bay, Tianjin, China. The sediments were artificially spiked according to the procedures of Li et al. (2015). Briefly, after the sediments were air-dried and passed through a 2-mm sieve, a portion (about 1/6) was spiked with DBP (99% purity; Sigma). The other portion (about 5/6) was flooded with artificial seawater and placed in the dark for one week to recovery bacterial activity. Then the two portions were thoroughly mixed to obtain a final concentration of 8.78 mg DBP/kg. Basic properties of the spiked sediments are listed in Table S1 and S2. C. closterium was purchased from the Institute of Oceanology of the Chinese Academy of Sciences. Procedures and conditions of the algal cultivation are given in the Supplementary material. Before use, algal cells in the exponential growth period were centrifuged at 3000 g for 5 min. The supernatant was removed carefully. The algal cells left were resuspended artificial seawater. 2.2. Experiment design 2.2.1. Degradation experiment After the spiked sediments were prepared, degradation experiment was carried out immediately in a series of 200 mL glass beakers. Twenty five gram of the spiked sediments (dry weight) was added into each of the glass beakers. Two different amounts of the particles (BC400, MWNT, PE and nZVI) were then added to the beakers and thoroughly mixed with the sediments, obtaining two particle concentrations (0.2 and 2.0%, w/w). Beakers containing sterilized (0.2% w/w HgCl2) or unsterilized sediments only were
prepared as control. In half of the beakers, C. closterium was evenly implanted on the surface of sediments. The initial density was 2 105 cell/cm2. One hundred milliliter of artificial seawater was slowly added into each beaker. All beakers were placed in an intelligent illumination incubator with a light intensity of 3200 Lux. The temperature was maintained at 25 C. Light:dark ratio was 16:8. The experiments were performed in triplicate and lasted for 8 d. At the end, sediment samples were collected for the analysis of DBP concentration, chlorophyll a content and bacterial community structure. More details about the experimental setup are shown in Fig S1. 2.2.2. Sorption experiment Sorption isotherms of DBP on the environmental particles and sediment sample were obtained using a batch equilibration technique. Briefly, an appropriate amount of the environmental particles and sediment sample was added into glass vials with Teflonlined screw caps (Angilent). Then, 40 mL of artificial seawater containing 0.4e10 mg/L DBP and 200 mg/L HgCl2 (for the inhibition of biodegradation) was added into the vials. Particle-to-water ratios were 5 mg per 40 mL for PE, nZVI, BC400 and sediment, and 1 mg per 160 mL for MWNT, respectively. The vials were shaken in a rotary shaker (120 rpm) at 25 C for 48 h, which was sufficient to reach sorption equilibrium according to preliminary kinetic experiment (Fig. S2). In addition, the sediment organic carbon was removed by the oxidation of H2O2 according to the method of Zhao et al. (2004). Procedure for obtaining sorption isotherm of DBP on the H2O2-treated sediment sample was the same as that used for the sediment sample before H2O2 treatment. After the sorption equilibrium, all vials were centrifuged at 3000 g for 30 min. The supernatant was sampled for the analysis of DBP by high performance liquid chromatography (HPLC). The amount of DBP sorbed on particles was calculated by the mass difference of DBP in solution because sorption onto the vial walls and biodegradation were negligible. All of the adsorption experiments were performed in triplicate. Details of chromatographic conditions are presented in the Supplementary material. 2.3. Analytical methods Concentration of DBP in the sediment samples were analyzed using ultrasonic extraction with rectified dichloromethane and then determined by gas chromatograph with flame ionization detector (GC-FID) according to the method of Li et al. (2015). Details of extraction procedures and chromatographic conditions are presented in the Supplementary material. The bacterial community structures in sediments were analyzed using 16s rDNA sequencing. DNA extraction, purification and sequencing procedures are given in the Supplementary Material. Concentration of chlorophyll a was measured according to the method of Li et al. (2015). 2.4. Statistical analysis Comparisons of treatment effects and means were obtained by variance analysis (ANOVA) and Duncan’s test at p 0.05, respectively. Statistical analyses were carried out using SPSS (V23.0, USA). 3. Results 3.1. Algal growth At the end of the experiment, contents of chlorophyll a in sediments inoculated with C. closterium are shown in Fig. 1. Compared with the initial value (1.36 mg/g), chlorophyll a content in all
F. Zhang et al. / Environmental Pollution 261 (2020) 114228
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lowest for nZVI (1.30). Moreover, when the equilibrium concentration (Ce) is in the range of 0.5e5% aqueous solubility of DBP (Cs), sorption coefficient (Kd, L/g) of DBP on sediment sample was about two orders of magnitude lower than that on MWNT, 1.2e2.2 times lower than those on BC400 and PE, and comparable to that on nZVI. Freundlich 1/n value of nZVI was higher than the other three particles. In addition, H2O2 treatment slightly decreased the logKf value for sediment. 3.3. DBP concentrations in sediments
Fig. 1. The contents of chlorophyll a in sediments at the end of the experiment. Sediment represents the control in the absence of particles. BC400, PE, nZVI and MWNT represent treatments with biochar pyrolyzed at 400 C, polyethylene microplastic, nanoscale zero-valent iron and multi-walled carbon nanotube, respectively. 0.2% and 2% represent particle addition concentrations, respectively.
treatments increased by 2.8e3.7 times except in treatments with MWNT addition. Compared to the control without particles, addition of PE and nZVI enhanced chlorophyll a content by 13.0e15.9% and 6.7e20.7%, respectively, while addition of MWNT reduced chlorophyll a content by 86.4e98.9%. There was no obvious change (p > 0.05) of chlorophyll a content in treatment with BC400 addition. Moreover, chlorophyll a content decreased (p < 0.05) with the increase of nZVI and MWNT concentrations, but had no obvious change (p > 0.05) with the increase of BC400 and PE concentrations. 3.2. DBP sorption by sediments and environmental particles As shown in Fig. S3, sorption isotherms of DBP on the four environmental particles and the sediment sample were well fitted with the Freundlich model (R2 0.96). The parameters related to the isotherm linearity (1/n) and sorption ability (logKf) are listed in Table 1. The logKf value was the highest for MWNT (3.65) and the Table 1 Freundlich isotherm parameters of DBP sorption on sediments and different particles a. Particle
b
1/n
logKf
R2
Kd (L/g) Ce ¼ 0.005 Cs
Sediment Sediment-t BC400 MWNTs PE nZVI
0.44 0.51 0.50 0.55 0.59 0.76
1.98 1.67 2.21 3.65 2.05 1.30
0.96 0.95 0.99 0.99 0.99 0.98
9.2 6.1 20.0 683.1 20.4 7.4
c
Ce ¼ 0.05 Cs 2.5 2.0 6.3 243.0 8.0 4.2
a The Freundlich parameters were calculated using the logarithmic form of the equation qe ¼ Kf*C1/n e , where qe is the amount sorbed per unit weight of sorbent, mg/ g; Ce is the equilibrium concentration, mg/L; Kf [(mg/g)/(mg/L)1/n] is the Freundlich capacity coefficient, and 1/n is the Freundlich nonlinearity index; R2 is regression coefficient. b Sediment and Sediment-t represent sediment samples before and after H2O2 oxidation. BC400, PE, nZVI and MWNT represent biochar pyrolyzed at 400 C, polyethylene microplastic, nanoscale zero-valent iron and multi-walled carbon nanotube, respectively. c Cs is the aqueous solubility of DBP at 25 C (13.0 mg/L).
Concentration of DBP in sterilized sediments showed no obvious change (p > 0.05) during the experiment (data not shown). Therefore, it can be considered that biotic processes played a major role in the dissipation of DBP. At the end of the experiment, Tables 3 and 2 represent residual concentrations and degradation percentages of DBP in sediments, respectively. In treatments without C. closterium, addition of different particles reduced degradation percentage of DBP to different degrees. When the particle concentration was lower (0.2% w/w), degradation percentage of DBP decreased in treatments with MWNT addition (4.4% decrement) and BC400 addition (1.2% decrement), but increased in treatment with nZVI addition (1.7% increment) and had no obvious changes (p > 0.05) in treatments with PE addition. When the particle concentration was higher (2% w/w), the decrement (1.3e15.7%) of degradation percentage of DBP was increased. In treatments with C. closterium, particle effects were normally more significant. For example, in treatments with 0.2% particle, the decrement was 11.4% in the presence of MWNT, and the increment was 4.1% in the presence of nZVI. In treatments with 2% particle, the decrements (5.3e24.8%) of degradation percentage of DBP were more significantly enhanced and even higher than those in treatments with 2% particle but without C. closterium. Compared with treatments without C. closterium, inoculation of C. closterium increased degradation percentage of DBP in all treatments with particle addition (1.9e11.3% increment) except treatment with 2% MWNT addition (no obvious change, p > 0.05). The increment was the highest for treatment with 0.2% nZVI addition and the lowest for treatment with 2% MWNT addition. 3.4. Bacterial community structure in sediments The results of 16s rDNA sequencing showed that the relative abundances of C. closterium accounted for 20.0e49.3% of the total taxon read numbers in sediment samples (Table S4). After the data of C. closterium was removed, data of the remaining bacterial operative taxonomic units (OTUs) are shown in Table 3. It can be seen that inoculation of C. closterium decreased bacterial OTU number by 1.3e15.5% in all treatments except in treatments with PE addition (increased by 0.6e14.1%) or treatment without particles addition (increased by 2.8%). The decrement was the highest for treatments with MWNT addition (by 12.8e15.5%). The change trends of the Simpson index and the Shannon index were similar to those of OTU number (Table 3). The results of principal coordinates analysis (PCoA) are given in Fig. 2. The first principal component (PC1) and the second principal component (PC2) explained 75.73% and 12.51% of the data variability, respectively. The results showed that the points of treatments with and without C. closterium were separated along PC1, but little apart along PC2 except the points of treatments with MWNT but without C. closterium. As shown in Fig. 3, Proteobacteria, Bacteroidetes and Firmicutes were the top three abundant bacterial phyla in all treatments and accounted for 94.2e97.2% of bacterial sequences in total. In general, inoculation of C. closterium enhanced the relative abundance of
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Fig. 2. Principal coordinates analysis (PCoA) in sediments with (C-) and without (N-) C. closterium. Sediment represents the control in the absence of particles. BC400, PE, nZVI and MWNT represent treatments with biochar pyrolyzed at 400 C, polyethylene microplastic, nanoscale zero-valent iron and multi-walled carbon nanotube, respectively. 0.2% and 2% represent particle addition concentrations, respectively.
Fig. 3. Relative abundances of the top 5 bacterial phyla in sediment samples of different treatments. N and C represent the treatments without and with C. closterium, respectively. Sediment represents the control in the absence of particles. BC400, PE, nZVI and MWNT represent treatments with biochar pyrolyzed at 400 C, polyethylene microplastic, nanoscale zero-valent iron and multi-walled carbon nanotube, respectively. 0.2% and 2% represent particle addition concentrations, respectively.
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Proteobacteria (6.2e16.9% increment), but reduced the relative abundance of Bacteroidetes (0.0e10.8% decrement) and Firmicutes (3.9e9.4% decrement). For treatment without particle, the changes were relatively small. Similarly, addition of the particles increased the relative abundance of Proteobacteria, but decreased the relative abundances of Bacteroidetes and Firmicutes in general. Variation of the relative abundances of the three phyla was more significant with the increase of particle concentrations. The relative abundances of the top 20 bacterial genera belonged to five phyla (Fig. 4). Among them, 11, 4 and 3 genera belonged to Proteobacteria, Firmicutes and Bacteroidetes, respectively. The other two genera (Haloplasma and Rhodococcus) belonged to Tenericutes and Actinobacteria, respectively. It can be noted from Table S5 that most of the dominant genera belonging to Proteobacteria were aerobic bacteria, while all dominant genera belonging to Firmicutes were anaerobic bacteria. Among the three dominant genera belonging to Bacteroidetes, one (Salinimicrobium) was an anaerobic bacterium. Its relative abundance was much higher than those of the other two aerobic bacteria. Moreover, inoculation of C. closterium enhanced the total relative abundances of dominant aerobic genera by 92e122% in treatments with 0.2% particle addition and by 15e79% in treatments with 2% particle addition. Meanwhile, the total relative abundances of dominant anaerobic genera were reduced by 1e21% in treatments with 0.2% particle addition and by 2e12% in treatments with 2% particle addition. 4. Discussion 4.1. Influence of environmental particles on the growth of C. closterium The results of this work indicated that chlorophyll a content in sediments was affected by different kinds and different
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concentrations of environmental particles. Addition of PE was found to increase chlorophyll a content. This may be because microplastics could provide surface for algal attachment (Casabianca et al., 2019) and promote algal growth (Canniff and Hoang, 2018; Casabianca et al., 2019). The reason for nZVI to enhance chlorophyll a content in sediments is different from that for PE. Iron is an essential element for algae. After nZVI enters into the environment, it could release Fe2þ for the growth of algae (Kadar et al., 2012; Guan et al., 2015). However, when nZVI concentration was higher (e.g. 2% in this work), promotion effect on algal growth by nZVI obviously decreased because of its potential negative impact on microorganisms. For example, nZVI attached to the surface of microorganisms may lead to increased membrane permeability and decreased cell mobility and nutrient flow be tween the cell’s exterior and interior compartments (Sevc u et al., 2011). Excess of free intracellular iron could lead to oxidative stress (Adeleye et al., 2016). In contrast, addition of MWNT reduced the chlorophyll a content. At the beginning, a small amount of yellow-green algae mat was observed on the sediment surface and then the algal mat gradually disappeared. The higher the MWNT concentration added, the more obvious the phenomenon. This is because the presence of MWNT adversely impacts bacterial communities which play an important role in nutrient cycling (e.g. phosphorus and carbon) (Rodrigues et al., 2013). In addition, MWNT had harmful impacts on microalgae such as growth, photosynthetic activities, oxidative stress and intracellular glutathione, which was increased with increasing MWNT concentration (Thakkar et al., 2016). 4.2. Influence of environmental particles on bacterial community structure in sediments The results of 16s rDNA sequencing showed that the points of
Fig. 4. Relative abundances of the top 20 bacterial genera in sediment samples of different treatments. N and C represent the treatments without and with C. closterium, respectively. Sediment represents the control in the absence of particles. BC400, PE, nZVI and MWNT represent treatments with biochar pyrolyzed at 400 C, polyethylene microplastic, nanoscale zero-valent iron and multi-walled carbon nanotube, respectively. 0.2% and 2% represent particle addition concentrations, respectively.
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treatments without and with C. closterium were apart from each other along PC1 (Fig. 2), indicating that inoculation of C. closterium obviously influenced bacterial community structure in sediments because PC1 contains most of the information of the original data. In the submerged sediments, the contents of oxygen are usually low due to limited supply of atmospheric oxygen. C. closterium could release oxygen into sediments and then modify the oxygen condition of surface sediments (Li et al., 2015). Similar phenomenon was also observed by Yamamoto et al. (2008). In this work, enhancement of the total relative abundances of dominant aerobic bacteria by C. closterium was mainly due to the increase in relative abundances of the dominant genera belonging to Proteobacteria. At the same time, reduction of the total relative abundances of dominant anaerobic bacteria was a result of the decrease in relative abundances of the dominant genera belonging to Firmicutes. Therefore, the most possible reason for the changes in bacterial community structure between the treatments without and with C. closterium was increased sediment oxygenation by C. closterium, which is consistent with the previous studies (Yamamoto et al., 2008; Li et al., 2015). The PCoA results showed that the points of treatments with different particle addition clustered together along PC2 except the points of treatments with MWNT but without C. closterium (Fig. 2). Therefore, effect of particle addition on the bacterial community structure was much smaller than that of C. closterium. Previous study found that even at very low concentrations (250e500 mg/g soil), the bacterial community was affected by the presence of carbon nanotubes because of their negative impacts on soil nutrient cycling (Rodrigues et al., 2013). In addition, it was found that when MWNT concentrations added in soil reached 500 mg/kg (0.05%, w/w), microbial activity was significantly inhibited (He et al., 2015). In this work, MWNT concentrations added was much larger than this dose. 4.3. Sorption behaviors of DBP on sediments and environmental particles Sorption behaviors of DBP on sediment samples and the four particles could be described well by the Freundlich model (Table 1) and were consistent with those of previous studies (Wang et al., 2010; Sun et al., 2012; Kleinteich et al., 2018). It can be seen that sorption abilities of DBP on the particles were obviously different. Among them, sorption of DBP on MWNT was the strongest (logKf ¼ 3.65) because of its large specific surface area and high surface hydrophobicity. Moreover, the sorption of DBP on MWNT was much higher than that on sediment sample, indicating that addition of MWNT may increase the sorption affinity of DBP to the sediment. In this work, concentrations of DBP in sediment porewater were low (no more than 0.32 mg/L; data not shown). Correspondingly, at low aqueous concentrations (0.5e5% Cs), sorption of DBP on PE and BC400 was only several times higher than that on the sediment (Table 1). Therefore, addition of 0.2% BC400 and PE had negligible impact on sorption affinity of DBP to the sediment, and the impact increased with the increase of particle concentration from 0.2% to 2%. For nZVI, its effect of on the sorption affinity of DBP to the sediment was small because sorption of DBP on nZVI was similar to that on the sediment at low aqueous concentrations (0.5e5% Cs). It is known that DBP is a hydrophobic (logKow ¼ 4.45) compound and preferred to be adsorbed to a hydrophobic environment, such as organic matter (Yang et al., 2013) and clay surfaces (Liu et al., 2013). It was reported that soil adsorption capacity of DBP was positively correlated to soil organic matter content (Wu et al., 2018) and tend to adsorb to the organic matter associated with solids (e.g. soil and sediment) (Staples et al., 1997). However, it can be seen
from Table 1 that sediment organic matter had a minor role in DBP sorption. This might be due to low TOC content of the sediment (0.34%). 4.4. Influence of environmental particles on degradation of DBP in sediments In treatments without C. closterium, addition of 0.2% particle had only small impact on DBP degradation in sediments (from 1.7% increment to 4.4% decrement). With the increase of particle concentration from 0.2% to 2%, degradation percentage of DBP was more obviously inhibited by the particles added (1.3%e15.7% decrement). In general, particle effect on degradation percentage of DBP increased with the increase of particle sorption ability and particle concentration. It is known that particles with small size normally had relatively strong sorption ability for organic pollutants due to their large surface area, hence influencing degradation of the pollutants (Khorram et al., 2016). Stronger sorption to particles leads to lower bioavailability (Glomstad et al., 2016; Zhang et al., 2016; Liu et al., 2019). For example, the decrement of degradation percentage of DBP was the highest in the presence of MWNT due to its strong sorption ability for DBP. Many studies have reported that carbon nanotube amendments can largely decrease the bioavailability of hydrophobic organic pollutants by serving as a strong sorbent, thus resulting in reduced degradation of the pollutants (Xia et al., 2010). In addition, inhibition of MWNT on microbial activity also contributed to the decrease of DBP degradation as mentioned in Section 4.2. For nZVI, its inhibitory effect on DBP degradation was small in treatment with 2% particle addition. This is mainly because sorption affinity of DBP to nZVI was comparable to that of sediment sample. However, degradation percentage of DBP was more obviously inhibited in treatment with 2% PE addition (11.5% decrement) than in treatment with 2% BC400 addition (4.1% decrement) although sorption affinity of DBP to PE was similar to that of BC400. In addition to reducing bioavailability of organic pollutants, particles effect on degradation of the pollutants may occur by changing indigenous sediment microbial community (Shrestha et al., 2013). Among the dominant genera in this work, Bacillus, Rhodococcus and Pseudomonas have been reported to be capable of degrading PAEs (Gao and Wen, 2016; Yang et al., 2018). It can be seen from Table S4 that relative abundance of the dominant PAE-degrading bacteria was about 36% higher in treatment with 2% BC400 addition than in treatment with 2% PE addition. Previous studies reported that biochar contains a range of nutrients (e.g. inorganic salts) and can bring benefits for microbial growth (Joseph et al., 2013). Therefore, higher microbial activity in treatment with 2% BC400 addition led to faster degradation of DBP as compared to treatment with 2% PE addition. It has been reported that C. closterium could degrade DBP directly (Gao and Chi, 2015; Gao and Chi, 2016). Therefore, in treatments with C. closterium, DBP could be degraded not only by bacteria but also by the alga. As shown in Table S4, in treatment with sediments only, relative abundance of the dominant PAEdegrading bacteria was 6.6% in total, which was reduced by 35% in the presence of C. closterium. Meanwhile, relative abundance of C. closterium accounted for 21.6% of the total taxon read numbers including both C. closterium and bacteria. This is probably due to nutrient competition between the alga and bacteria under nutrient-limited conditions, which inhibited bacteria growth (Gao and Chi, 2016). Moreover, as shown in Table 2, the presence of C. closterium increased degradation percentage of DBP (8.9% increment). Contribution of C. closterium to DBP degradation should be much higher than the increment (i.e. 8.9%). Therefore, it can be considered that C. closterium played a key role in degradation of DBP in the alga-bacteria system, which led to the degradation
F. Zhang et al. / Environmental Pollution 261 (2020) 114228 Table 2 Degradation percentages of DBP in sediments with and without particles. Sample
DBP degradation percentages (%) Without C. closterium
Sediment 0.2% BC400 0.2% MWNT 0.2% PE 0.2% nZVI 2% BC400 2% MWNT 2% PE 2% nZVI
76.6 75.4 72.2 76.8 78.3 72.5 60.9 65.1 75.3
± ± ± ± ± ± ± ± ±
0.6B* 0.9C* 0.7D* 2.2B* 0.9A* 2.7D* 0.1F 1.7E* 2.4C*
With C. closterium 85.5 84.8 74.1 85.4 89.6 77.3 60.7 70.9 80.2
± ± ± ± ± ± ± ± ±
2.8b* 2.9b* 0.8e* 1.9b* 4.5a* 4.1d* 0.1g 1.0f* 0.1c*
Sediment represents the control in the absence of particles. BC400, PE, nZVI and MWNT represent treatments with biochar pyrolyzed at 400 C, polyethylene microplastic, nanoscale zero-valent iron and multi-walled carbon nanotube, respectively. 0.2% and 2% represent particle addition concentrations, respectively. Letters AeF and aeg indicate significant difference (p < 0.05) among different particle addition treatments without and with C. closterium, respectively. Symbol * indicates significant difference (p < 0.05) between treatments with and without C. closterium.
enhancement. A similar phenomenon was found in our previous study (Gao and Chi, 2016). Moreover, addition of 0.2% particles could influence DBP degradation mainly by promoting or inhibiting the growth of C. closterium in some cases in this work. For example, addition of 0.2% nZVI had a small effect on DBP degradation in treatment without C. closterium (1.7% increment) but a larger effect in treatment with C. closterium (4.1% increment) by increasing the biomass of C. closterium. In contrast, the decrement in treatment with 0.2% MWNT addition was 4.4% in the absence of C. closterium but 11.4% in the presence of C. closterium, mainly due to its significant inhibition on the growth of C. closterium. Besides, addition of BC400 had no obvious (p > 0.05) effects on both the biomass of C. closterium and degradation percentage of DBP, while addition of PE enhanced the biomass of C. closterium by 13% but showed no obvious (p > 0.05) impact on DBP degradation. This might be partly due to the differences in their properties (Table S1). Clearly, deeper investigation is needed to understand the mechanisms. However, with the increase of particle concentration from 0.2% to 2%, particle addition on sorption affinity of DBP to the sediment played a more significant role in its effect on DBP degradation. Compared with treatments without C. closterium, inoculation of
7
C. closterium increased degradation percentage of DBP in all treatments with particle addition (1.9e11.3% increment) except treatment with 2% MWNT addition (no effect, p > 0.05). The increment increased with the increase of chlorophyll a content in general. For example, the increment was the highest for treatment with 0.2% nZVI addition (11.3%) due to its highest promotion effect on algal growth. In contrast, the increment was the lowest for treatments with MWNT addition (no more than 1.9%). This is consistent with the strong sorption (indicated as logKf) of MWNT to DBP and its strong inhibition on the growth of C. closterium (indicated as chlorophyll a content). The increment decreased with the increase of particle concentration from 0.2% to 2% (from no effect to 5.8% increment for treatments with 2% addition). Because the biomass of C. closterium did not change much with the increase of particle concentration from 0.2% to 2%, the particles achieved more significant effect on DBP degradation by inhibiting its bioavailability. For example, degradation percentage of DBP declined with the increase of BC400 and PE concentrations added although chlorophyll a content had no obvious change (p > 0.05). Based on the above discussion, it can be considered that the particles could influence bioavailability of DBP by sorption and biomass of C. closterium, and thus degradation of DBP in sediments. As more and more environmental particles are being spread and accumulated in the environment due to their growing use, higher particle concentration might negatively impact on the ecological environment, meanwhile reduce biodegradation rate of DBP. Recently, we found that C. closterium could influence the net flux of DBP from water to sediment by changing DBP degradation in sediments, and then the total biodegradation flux in the watersediment system (Zhang et al., 2019). Hence, the reduced biodegradation rate by environmental particles probably leads to higher degradation half-life of the pollutant and thus enhancing its persistence in the coastal environment. Besides environmental particles, many factors can also influence algal growth, such as nutrient condition in the overlying water, sediment composition, and so on. Clearly, more research is needed to evaluate the environmental risk of the pollutant in the presence of marine microphytobenthos. 5. Conclusions The results in this work showed that effects of environmental
Table 3 Bacterial diversity indices, read numbers and OTUs of different sediment samples. Sample
Taxon read number
OTUs
Shannon index
Simpson index
Without C. closterium
Sediment 0.2% BC400 0.2% PE 0.2% nZVI 0.2% MWNT 2% BC400 2% PE 2% nZVI 2% MWNT
65232 64821 64804 64108 62793 63364 63775 64122 63042
532 534 498 529 586 546 517 531 567
5.354 5.370 5.298 5.304 5.936 5.540 5.333 5.394 5.534
0.923 0.923 0.922 0.920 0.954 0.939 0.924 0.933 0.950
With C. closterium
Sediment 0.2% BC400 0.2% PE 0.2% nZVI 0.2% MWNT 2% BC400 2% PE 2% nZVI 2% MWNT
52686 48028 47494 50692 42921 46827 52330 53384 36969
547 527 501 503 511 501 590 507 479
5.300 5.047 4.965 5.189 4.702 4.964 5.571 5.299 3.899
0.919 0.895 0.892 0.913 0.850 0.886 0.933 0.928 0.747
Sediment represents the control in the absence of particles. BC400, PE, nZVI and MWNT represent treatments with biochar pyrolyzed at 400 C, polyethylene microplastic, nanoscale zero-valent iron and multi-walled carbon nanotube, respectively. 0.2% and 2% represent particle addition concentrations, respectively.
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F. Zhang et al. / Environmental Pollution 261 (2020) 114228
particles on the biomass of C. closterium (indicated as chlorophyll content) were obviously different among the particles, which are, inhibition by multi-walled carbon nanotube (86.4e98.9%), stimulation by polyethylene microplastic (13.0e15.9%) and nanoscale zero-valent iron (6.7e20.7%) and no effect by biochar pyrolyzed at 400 C. Besides, the results of 16s rDNA sequencing showed that C. closterium was probably the most abundant DBP-degrader, accounting for 20.0e49.3% of the total taxon read numbers. Its biomass was the most important factor influencing bacterial community structure in sediments. In treatments without C. closterium, degradation percentage of DBP decreased with the increase of particle sorption ability to DBP and particle concentration in general. In treatments with C. closterium, inoculation of C. closterium increased the degradation percentage of DBP in all treatments with particle addition, which increased with the increase of chlorophyll a content in general but decreased with the increase of particle concentration from 0.2% to 2%. Our findings suggested that the environmental particles could influence bioavailability of DBP by sorption and biomass of C. closterium, and thus degradation of DBP in sediments. Therefore, impact of the environmental particles should be considered in the environmental risk assessment of the pollutant. Declaration of competing interest We declare that we do not have any commercial or associative interest that represents a conflict of interest in connection with the work submitted. Appendix A. Supplementary data Supplementary data related to this article can be found at https://doi.org/10.1016/j.envpol.2020.114228. References Adeleye, A.S., Stevenson, L.M., Su, Y., Nisbet, R.M., Zhang, Y., Keller, A.A., 2016. Influence of phytoplankton on fate and effects of modified zerovalent iron nanoparticles. Environ. Sci. Technol. 50, 5597e5605. Alimba, C.G., Faggio, C., 2019. Microplastics in the marine environment: current trends in environmental pollution and mechanisms of toxicological profile. Environ. Toxicol. Pharmacol. 68, 61e74. Anbumani, S., Kakkar, P., 2018. Ecotoxicological effects of microplastics on biota: a review. Environ. Sci. Pollut. Res. 25, 14373e14396. Ask, J., Rowe, O., Brugel, S., Stromgren, M., Bystrom, P., Andersson, A., 2016. Importance of coastal primary production in the northern Baltic Sea. Ambio 45 (6), 635e648. Beckingham, B., Ghosh, U., 2017. Differential bioavailability of polychlorinated biphenyls associated with environmental particles: microplastic in comparison to wood, coal and biochar. Environ. Pollut. 220, 150e158. Canniff, P.M., Hoang, T.C., 2018. Microplastic ingestion by Daphnia magna and its enhancement on algal growth. Sci. Total Environ. 633, 500e507. Casabianca, S., Capellacci, S., Giacobbe, M.G., Dell’Aversano, C., Tartaglione, L., Varriale, F., Narizzano, R., Risso, F., Moretto, P., Dagnino, A., Bertolotto, R., Barbone, E., Ungaro, N., Penna, A., 2019. Plastic-associated harmful microalgal assemblages in marine environment. Environ. Pollut. 244, 617e626. Chi, J., Li, Y., Gao, J., 2019. Interaction between three marine microalgae and two phthalate acid esters. Ecotoxicol. Environ. Saf. 170, 407e411. Frias, J.P.G.L., Gago, J., Otero, V., Sobral, P., 2016. Microplastics in coastal sediments from southern Portuguese shelf waters. Mar. Environ. Res. 114, 24e30. Gao, D., Li, Z., Wang, H., Liang, H., 2018. An overview of phthalate acid ester pollution in China over the last decade: environmental occurrence and human exposure. Sci. Total Environ. 645, 1400e1409. Gao, D.W., Wen, Z.D., 2016. Phthalate esters in the environment: a critical review of their occurrence, biodegradation and removal during wastewater treatment processes. Sci. Total Environ. 541, 986e1001. Gao, J., Chi, J., 2015. Biodegradation of phthalate acid esters by different marine microalgal species. Mar. Pollut. Bull. 99 (1e2), 70e75. Gao, J., Chi, J., 2016. A new approach to biodegrading diethyl phthalate in the benthic sea-floor diatom-bacteria system. J. Saf. Environ. 16 (4), 311e314 (in Chinese). Glomstad, B., Altin, D., Sørensen, L., Liu, J., Jenssen, B.M., Booth, A.M., 2016. Carbon nanotube properties influence adsorption of phenanthrene and subsequent bioavailability and toxicity to Pseudokirchneriella subcapitata. Environ. Sci.
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Shrestha, B., Acosta-Martinez, V., Cox, S.B., Green, M.J., Li, S., Can An evaluation of the impact of multiwalled carbon nanotubes on soil microbial community structure and functioning. J. Hazard Mater. 261, 188e197. Staples, C.A., Peterson, D.R., Parkerton, T.F., Adams, W.J., 1997. The environmental fate of phthalate esters: a literature review. Chemosphere 35 (4), 667e749. Subashchandrabose, S.R., Ramakrishnan, B., Megharaj, M., Venkateswarlu, K., Naidu, R., 2013. Mixotrophic cyanobacteria and microalgae as distinctive biological agents for organic pollutant degradation. Environ. Int. 51, 59e72. Sun, K., Jin, J., Keiluweit, M., Kleber, M., Wang, Z., Pan, Z., Xing, B., 2012. Polar and aliphatic domains regulate sorption of phthalic acid esters (PAEs) to biochars. Bioresour. Technol. 118, 120e127. Thakkar, M., Mitra, S., Wei, L., 2016. Effect on growth, photosynthesis, and oxidative stress of single walled carbon nanotubes exposure to marine alga Dunaliella tertiolecta. J. 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Impact of different environmental particles on degradation of dibutyl phthalate in coastal sediments with and without Cylindrotheca closterium Fan Zhang, Dongxu Zhao, Jie Chi* School of Environmental Science and Engineering, Tianjin University, Tianjin, 300350, PR China
a r t i c l e i n f o
a b s t r a c t
Article history: Received 29 August 2019 Received in revised form 3 February 2020 Accepted 16 February 2020 Available online 19 February 2020
This study investigated the impact of different environmental particles at different concentrations (0.2% and 2%, w/w) on biodegradation of dibutyl phthalate (DBP) in sediments with and without Cylindrotheca closterium, a marine benthic diatom. The particles included biochar pyrolyzed at 400 C, multi-walled carbon nanotube (MWNT), nanoscale zero-valent iron (nZVI) and polyethylene microplastic. In treatments without C. closterium, inhibition effect of the particles on degradation percentage of DBP (up to 15.7% decrement except 1.7% increment for 0.2% nZVI) increased with the increase of particle sorption ability to DBP and particle concentration in general. The results of 16s rDNA sequencing showed that C. closterium was probably the most abundant DBP-degrader, accounting for 20.0e49.3% of the total taxon read numbers. In treatments with C. closterium, inoculation of C. closterium increased the degradation percentage of DBP in all treatments with particle addition by 0.0e11.3%, which increased with the increase of chlorophyll a content in general but decreased with the increase of particle concentration from 0.2% to 2%. The increment was the highest for treatment with 0.2% nZVI addition due to its highest promotion effect on algal growth. In contrast, the increment was the lowest for treatments with MWNT addition due to its strong sorption to DBP and strong inhibition on the growth of C. closterium. Our findings suggested that the environmental particles could influence bioavailability of DBP by sorption and biomass of C. closterium, and thus degradation of DBP in sediments. © 2020 Elsevier Ltd. All rights reserved.
Keywords: Environmental particles Marine benthic diatom Dibutyl phthalate Coastal sediments Degradation
1. Introduction With the development of science and technology, different kinds of particles are released into the environment during manufacture process and usage of chemicals and industrial materials. For example, microplastics and engineering nanoparticles (e.g. carbon nanotubes and zero-valent iron) are ubiquitous in soils, sediments, rivers, lakes and oceans (Horton and Dixon, 2018; Goswami et al., 2017). In addition, biochars pyrolyzed by decomposition of biomass under limited oxygen conditions are widely applied as a soil amendment (Zhu et al., 2017). Due to small size, these environmental particles usually possess large surface area and high sorption capacity for hydrophobic organic pollutants (e.g. phthalates, polycyclic aromatic hydrocarbons and polychlorinated
* Corresponding author. School of Environmental Science and Engineering, Tianjin University, Tianjin, 300350, China. E-mail address: [email protected] (J. Chi). https://doi.org/10.1016/j.envpol.2020.114228 0269-7491/© 2020 Elsevier Ltd. All rights reserved.
biphenyls), thus resulting in accumulation of the pollutants in the environment (Glomstad et al., 2016; Beckingham and Ghosh, 2017; Jiang, 2018). For example, carbon nanotube at a concentration of 5 mg/g significantly increased the residual amount of polycyclic aromatic hydrocarbons in soils (Li et al., 2013); adding biochar to soils obviously retained polycyclic aromatic hydrocarbons et al., 2019). Besides, environmental particles may be (Rombola ingested by organisms and cause disruption to their physical ingestion, metabolism and reproductive performance (Alimba and Faggio, 2019). Pollutants sorbed on the particles might be further released in the organisms after ingestion, enhancing their toxic susceptibility (Anbumani and Kakkar, 2018). It has been reported that coastal sediments are the major sink for environmental particles, especially near the areas of high population and industrial density (Frias et al., 2016; Goswami et al., 2017). Microphytobenthos grow on the surface of sediments. They are one of the most important primary producers in coastal area and have important ecological and environmental significance. Microphytobenthos can release oxygen and soluble substances into
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F. Zhang et al. / Environmental Pollution 261 (2020) 114228
environment (Ask et al., 2016), thus influencing bacterial community structure and degradation of organic pollutants (Subashchandrabose et al., 2013). Algal growth could also be impacted by environmental particles (Lambreva et al., 2015; Prata et al., 2019). Therefore, coexistence of environmental particles and microphytobenthos can impact degradation and persistence of hydrophobic organic compounds in coastal sediments. Phthalic acid esters (PAEs) are typical hydrophobic organic pollutants in coastal sediments (Gao et al., 2018). Cylindrotheca closterium is a benthic diatom commonly occurring in the marine sediments. Our previous studies found that C. closterium could degrade dibutyl phthalate (DBP) directly (Gao and Chi, 2015; Chi et al., 2019). In this work, four environmental particles were selected, including biochar pyrolyzed at 400 C (BC400), multiwalled carbon nanotube (MWNT), polyethylene microplastic (PE) and nanoscale zero-valent iron (nZVI). The aims are to investigate degradation of DBP in sediments when the particles and C. closterium singly existed or coexisted. Sorption ability of the particles for DBP was measured. The changes of C. closterium biomass and bacterial community structure in sediments were determined. Mechanisms of particle effects on biodegradation of DBP in sediments with or without C. closterium and potential contribution of each experimental compartment (particles, bacteria and the alga) to degradation of DBP were discussed. 2. Material and methods 2.1. Materials In this work, MWNT (>98 wt % purity), nZVI (99.9% purity) and PE (99% purity) were obtained from Chengdu Organic Chemicals Co. Ltd., Shanghai Macklin Biochemical Co. Ltd. and Zhonglian Plastic Technology Co. Ltd., respectively. Biochar was produced by pyrolyzing wheat straw chips at in a muffle furnace at 400 C under oxygen-limited condition. The charred solid was then sieved through a 0.178-mm sieve and named as BC400. The basic properties of the particles are given in Table S1. Surface sediments were sampled from the intertidal flats of the Bohai Bay, Tianjin, China. The sediments were artificially spiked according to the procedures of Li et al. (2015). Briefly, after the sediments were air-dried and passed through a 2-mm sieve, a portion (about 1/6) was spiked with DBP (99% purity; Sigma). The other portion (about 5/6) was flooded with artificial seawater and placed in the dark for one week to recovery bacterial activity. Then the two portions were thoroughly mixed to obtain a final concentration of 8.78 mg DBP/kg. Basic properties of the spiked sediments are listed in Table S1 and S2. C. closterium was purchased from the Institute of Oceanology of the Chinese Academy of Sciences. Procedures and conditions of the algal cultivation are given in the Supplementary material. Before use, algal cells in the exponential growth period were centrifuged at 3000 g for 5 min. The supernatant was removed carefully. The algal cells left were resuspended artificial seawater. 2.2. Experiment design 2.2.1. Degradation experiment After the spiked sediments were prepared, degradation experiment was carried out immediately in a series of 200 mL glass beakers. Twenty five gram of the spiked sediments (dry weight) was added into each of the glass beakers. Two different amounts of the particles (BC400, MWNT, PE and nZVI) were then added to the beakers and thoroughly mixed with the sediments, obtaining two particle concentrations (0.2 and 2.0%, w/w). Beakers containing sterilized (0.2% w/w HgCl2) or unsterilized sediments only were
prepared as control. In half of the beakers, C. closterium was evenly implanted on the surface of sediments. The initial density was 2 105 cell/cm2. One hundred milliliter of artificial seawater was slowly added into each beaker. All beakers were placed in an intelligent illumination incubator with a light intensity of 3200 Lux. The temperature was maintained at 25 C. Light:dark ratio was 16:8. The experiments were performed in triplicate and lasted for 8 d. At the end, sediment samples were collected for the analysis of DBP concentration, chlorophyll a content and bacterial community structure. More details about the experimental setup are shown in Fig S1. 2.2.2. Sorption experiment Sorption isotherms of DBP on the environmental particles and sediment sample were obtained using a batch equilibration technique. Briefly, an appropriate amount of the environmental particles and sediment sample was added into glass vials with Teflonlined screw caps (Angilent). Then, 40 mL of artificial seawater containing 0.4e10 mg/L DBP and 200 mg/L HgCl2 (for the inhibition of biodegradation) was added into the vials. Particle-to-water ratios were 5 mg per 40 mL for PE, nZVI, BC400 and sediment, and 1 mg per 160 mL for MWNT, respectively. The vials were shaken in a rotary shaker (120 rpm) at 25 C for 48 h, which was sufficient to reach sorption equilibrium according to preliminary kinetic experiment (Fig. S2). In addition, the sediment organic carbon was removed by the oxidation of H2O2 according to the method of Zhao et al. (2004). Procedure for obtaining sorption isotherm of DBP on the H2O2-treated sediment sample was the same as that used for the sediment sample before H2O2 treatment. After the sorption equilibrium, all vials were centrifuged at 3000 g for 30 min. The supernatant was sampled for the analysis of DBP by high performance liquid chromatography (HPLC). The amount of DBP sorbed on particles was calculated by the mass difference of DBP in solution because sorption onto the vial walls and biodegradation were negligible. All of the adsorption experiments were performed in triplicate. Details of chromatographic conditions are presented in the Supplementary material. 2.3. Analytical methods Concentration of DBP in the sediment samples were analyzed using ultrasonic extraction with rectified dichloromethane and then determined by gas chromatograph with flame ionization detector (GC-FID) according to the method of Li et al. (2015). Details of extraction procedures and chromatographic conditions are presented in the Supplementary material. The bacterial community structures in sediments were analyzed using 16s rDNA sequencing. DNA extraction, purification and sequencing procedures are given in the Supplementary Material. Concentration of chlorophyll a was measured according to the method of Li et al. (2015). 2.4. Statistical analysis Comparisons of treatment effects and means were obtained by variance analysis (ANOVA) and Duncan’s test at p 0.05, respectively. Statistical analyses were carried out using SPSS (V23.0, USA). 3. Results 3.1. Algal growth At the end of the experiment, contents of chlorophyll a in sediments inoculated with C. closterium are shown in Fig. 1. Compared with the initial value (1.36 mg/g), chlorophyll a content in all
F. Zhang et al. / Environmental Pollution 261 (2020) 114228
3
lowest for nZVI (1.30). Moreover, when the equilibrium concentration (Ce) is in the range of 0.5e5% aqueous solubility of DBP (Cs), sorption coefficient (Kd, L/g) of DBP on sediment sample was about two orders of magnitude lower than that on MWNT, 1.2e2.2 times lower than those on BC400 and PE, and comparable to that on nZVI. Freundlich 1/n value of nZVI was higher than the other three particles. In addition, H2O2 treatment slightly decreased the logKf value for sediment. 3.3. DBP concentrations in sediments
Fig. 1. The contents of chlorophyll a in sediments at the end of the experiment. Sediment represents the control in the absence of particles. BC400, PE, nZVI and MWNT represent treatments with biochar pyrolyzed at 400 C, polyethylene microplastic, nanoscale zero-valent iron and multi-walled carbon nanotube, respectively. 0.2% and 2% represent particle addition concentrations, respectively.
treatments increased by 2.8e3.7 times except in treatments with MWNT addition. Compared to the control without particles, addition of PE and nZVI enhanced chlorophyll a content by 13.0e15.9% and 6.7e20.7%, respectively, while addition of MWNT reduced chlorophyll a content by 86.4e98.9%. There was no obvious change (p > 0.05) of chlorophyll a content in treatment with BC400 addition. Moreover, chlorophyll a content decreased (p < 0.05) with the increase of nZVI and MWNT concentrations, but had no obvious change (p > 0.05) with the increase of BC400 and PE concentrations. 3.2. DBP sorption by sediments and environmental particles As shown in Fig. S3, sorption isotherms of DBP on the four environmental particles and the sediment sample were well fitted with the Freundlich model (R2 0.96). The parameters related to the isotherm linearity (1/n) and sorption ability (logKf) are listed in Table 1. The logKf value was the highest for MWNT (3.65) and the Table 1 Freundlich isotherm parameters of DBP sorption on sediments and different particles a. Particle
b
1/n
logKf
R2
Kd (L/g) Ce ¼ 0.005 Cs
Sediment Sediment-t BC400 MWNTs PE nZVI
0.44 0.51 0.50 0.55 0.59 0.76
1.98 1.67 2.21 3.65 2.05 1.30
0.96 0.95 0.99 0.99 0.99 0.98
9.2 6.1 20.0 683.1 20.4 7.4
c
Ce ¼ 0.05 Cs 2.5 2.0 6.3 243.0 8.0 4.2
a The Freundlich parameters were calculated using the logarithmic form of the equation qe ¼ Kf*C1/n e , where qe is the amount sorbed per unit weight of sorbent, mg/ g; Ce is the equilibrium concentration, mg/L; Kf [(mg/g)/(mg/L)1/n] is the Freundlich capacity coefficient, and 1/n is the Freundlich nonlinearity index; R2 is regression coefficient. b Sediment and Sediment-t represent sediment samples before and after H2O2 oxidation. BC400, PE, nZVI and MWNT represent biochar pyrolyzed at 400 C, polyethylene microplastic, nanoscale zero-valent iron and multi-walled carbon nanotube, respectively. c Cs is the aqueous solubility of DBP at 25 C (13.0 mg/L).
Concentration of DBP in sterilized sediments showed no obvious change (p > 0.05) during the experiment (data not shown). Therefore, it can be considered that biotic processes played a major role in the dissipation of DBP. At the end of the experiment, Tables 3 and 2 represent residual concentrations and degradation percentages of DBP in sediments, respectively. In treatments without C. closterium, addition of different particles reduced degradation percentage of DBP to different degrees. When the particle concentration was lower (0.2% w/w), degradation percentage of DBP decreased in treatments with MWNT addition (4.4% decrement) and BC400 addition (1.2% decrement), but increased in treatment with nZVI addition (1.7% increment) and had no obvious changes (p > 0.05) in treatments with PE addition. When the particle concentration was higher (2% w/w), the decrement (1.3e15.7%) of degradation percentage of DBP was increased. In treatments with C. closterium, particle effects were normally more significant. For example, in treatments with 0.2% particle, the decrement was 11.4% in the presence of MWNT, and the increment was 4.1% in the presence of nZVI. In treatments with 2% particle, the decrements (5.3e24.8%) of degradation percentage of DBP were more significantly enhanced and even higher than those in treatments with 2% particle but without C. closterium. Compared with treatments without C. closterium, inoculation of C. closterium increased degradation percentage of DBP in all treatments with particle addition (1.9e11.3% increment) except treatment with 2% MWNT addition (no obvious change, p > 0.05). The increment was the highest for treatment with 0.2% nZVI addition and the lowest for treatment with 2% MWNT addition. 3.4. Bacterial community structure in sediments The results of 16s rDNA sequencing showed that the relative abundances of C. closterium accounted for 20.0e49.3% of the total taxon read numbers in sediment samples (Table S4). After the data of C. closterium was removed, data of the remaining bacterial operative taxonomic units (OTUs) are shown in Table 3. It can be seen that inoculation of C. closterium decreased bacterial OTU number by 1.3e15.5% in all treatments except in treatments with PE addition (increased by 0.6e14.1%) or treatment without particles addition (increased by 2.8%). The decrement was the highest for treatments with MWNT addition (by 12.8e15.5%). The change trends of the Simpson index and the Shannon index were similar to those of OTU number (Table 3). The results of principal coordinates analysis (PCoA) are given in Fig. 2. The first principal component (PC1) and the second principal component (PC2) explained 75.73% and 12.51% of the data variability, respectively. The results showed that the points of treatments with and without C. closterium were separated along PC1, but little apart along PC2 except the points of treatments with MWNT but without C. closterium. As shown in Fig. 3, Proteobacteria, Bacteroidetes and Firmicutes were the top three abundant bacterial phyla in all treatments and accounted for 94.2e97.2% of bacterial sequences in total. In general, inoculation of C. closterium enhanced the relative abundance of
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Fig. 2. Principal coordinates analysis (PCoA) in sediments with (C-) and without (N-) C. closterium. Sediment represents the control in the absence of particles. BC400, PE, nZVI and MWNT represent treatments with biochar pyrolyzed at 400 C, polyethylene microplastic, nanoscale zero-valent iron and multi-walled carbon nanotube, respectively. 0.2% and 2% represent particle addition concentrations, respectively.
Fig. 3. Relative abundances of the top 5 bacterial phyla in sediment samples of different treatments. N and C represent the treatments without and with C. closterium, respectively. Sediment represents the control in the absence of particles. BC400, PE, nZVI and MWNT represent treatments with biochar pyrolyzed at 400 C, polyethylene microplastic, nanoscale zero-valent iron and multi-walled carbon nanotube, respectively. 0.2% and 2% represent particle addition concentrations, respectively.
F. Zhang et al. / Environmental Pollution 261 (2020) 114228
Proteobacteria (6.2e16.9% increment), but reduced the relative abundance of Bacteroidetes (0.0e10.8% decrement) and Firmicutes (3.9e9.4% decrement). For treatment without particle, the changes were relatively small. Similarly, addition of the particles increased the relative abundance of Proteobacteria, but decreased the relative abundances of Bacteroidetes and Firmicutes in general. Variation of the relative abundances of the three phyla was more significant with the increase of particle concentrations. The relative abundances of the top 20 bacterial genera belonged to five phyla (Fig. 4). Among them, 11, 4 and 3 genera belonged to Proteobacteria, Firmicutes and Bacteroidetes, respectively. The other two genera (Haloplasma and Rhodococcus) belonged to Tenericutes and Actinobacteria, respectively. It can be noted from Table S5 that most of the dominant genera belonging to Proteobacteria were aerobic bacteria, while all dominant genera belonging to Firmicutes were anaerobic bacteria. Among the three dominant genera belonging to Bacteroidetes, one (Salinimicrobium) was an anaerobic bacterium. Its relative abundance was much higher than those of the other two aerobic bacteria. Moreover, inoculation of C. closterium enhanced the total relative abundances of dominant aerobic genera by 92e122% in treatments with 0.2% particle addition and by 15e79% in treatments with 2% particle addition. Meanwhile, the total relative abundances of dominant anaerobic genera were reduced by 1e21% in treatments with 0.2% particle addition and by 2e12% in treatments with 2% particle addition. 4. Discussion 4.1. Influence of environmental particles on the growth of C. closterium The results of this work indicated that chlorophyll a content in sediments was affected by different kinds and different
5
concentrations of environmental particles. Addition of PE was found to increase chlorophyll a content. This may be because microplastics could provide surface for algal attachment (Casabianca et al., 2019) and promote algal growth (Canniff and Hoang, 2018; Casabianca et al., 2019). The reason for nZVI to enhance chlorophyll a content in sediments is different from that for PE. Iron is an essential element for algae. After nZVI enters into the environment, it could release Fe2þ for the growth of algae (Kadar et al., 2012; Guan et al., 2015). However, when nZVI concentration was higher (e.g. 2% in this work), promotion effect on algal growth by nZVI obviously decreased because of its potential negative impact on microorganisms. For example, nZVI attached to the surface of microorganisms may lead to increased membrane permeability and decreased cell mobility and nutrient flow be tween the cell’s exterior and interior compartments (Sevc u et al., 2011). Excess of free intracellular iron could lead to oxidative stress (Adeleye et al., 2016). In contrast, addition of MWNT reduced the chlorophyll a content. At the beginning, a small amount of yellow-green algae mat was observed on the sediment surface and then the algal mat gradually disappeared. The higher the MWNT concentration added, the more obvious the phenomenon. This is because the presence of MWNT adversely impacts bacterial communities which play an important role in nutrient cycling (e.g. phosphorus and carbon) (Rodrigues et al., 2013). In addition, MWNT had harmful impacts on microalgae such as growth, photosynthetic activities, oxidative stress and intracellular glutathione, which was increased with increasing MWNT concentration (Thakkar et al., 2016). 4.2. Influence of environmental particles on bacterial community structure in sediments The results of 16s rDNA sequencing showed that the points of
Fig. 4. Relative abundances of the top 20 bacterial genera in sediment samples of different treatments. N and C represent the treatments without and with C. closterium, respectively. Sediment represents the control in the absence of particles. BC400, PE, nZVI and MWNT represent treatments with biochar pyrolyzed at 400 C, polyethylene microplastic, nanoscale zero-valent iron and multi-walled carbon nanotube, respectively. 0.2% and 2% represent particle addition concentrations, respectively.
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treatments without and with C. closterium were apart from each other along PC1 (Fig. 2), indicating that inoculation of C. closterium obviously influenced bacterial community structure in sediments because PC1 contains most of the information of the original data. In the submerged sediments, the contents of oxygen are usually low due to limited supply of atmospheric oxygen. C. closterium could release oxygen into sediments and then modify the oxygen condition of surface sediments (Li et al., 2015). Similar phenomenon was also observed by Yamamoto et al. (2008). In this work, enhancement of the total relative abundances of dominant aerobic bacteria by C. closterium was mainly due to the increase in relative abundances of the dominant genera belonging to Proteobacteria. At the same time, reduction of the total relative abundances of dominant anaerobic bacteria was a result of the decrease in relative abundances of the dominant genera belonging to Firmicutes. Therefore, the most possible reason for the changes in bacterial community structure between the treatments without and with C. closterium was increased sediment oxygenation by C. closterium, which is consistent with the previous studies (Yamamoto et al., 2008; Li et al., 2015). The PCoA results showed that the points of treatments with different particle addition clustered together along PC2 except the points of treatments with MWNT but without C. closterium (Fig. 2). Therefore, effect of particle addition on the bacterial community structure was much smaller than that of C. closterium. Previous study found that even at very low concentrations (250e500 mg/g soil), the bacterial community was affected by the presence of carbon nanotubes because of their negative impacts on soil nutrient cycling (Rodrigues et al., 2013). In addition, it was found that when MWNT concentrations added in soil reached 500 mg/kg (0.05%, w/w), microbial activity was significantly inhibited (He et al., 2015). In this work, MWNT concentrations added was much larger than this dose. 4.3. Sorption behaviors of DBP on sediments and environmental particles Sorption behaviors of DBP on sediment samples and the four particles could be described well by the Freundlich model (Table 1) and were consistent with those of previous studies (Wang et al., 2010; Sun et al., 2012; Kleinteich et al., 2018). It can be seen that sorption abilities of DBP on the particles were obviously different. Among them, sorption of DBP on MWNT was the strongest (logKf ¼ 3.65) because of its large specific surface area and high surface hydrophobicity. Moreover, the sorption of DBP on MWNT was much higher than that on sediment sample, indicating that addition of MWNT may increase the sorption affinity of DBP to the sediment. In this work, concentrations of DBP in sediment porewater were low (no more than 0.32 mg/L; data not shown). Correspondingly, at low aqueous concentrations (0.5e5% Cs), sorption of DBP on PE and BC400 was only several times higher than that on the sediment (Table 1). Therefore, addition of 0.2% BC400 and PE had negligible impact on sorption affinity of DBP to the sediment, and the impact increased with the increase of particle concentration from 0.2% to 2%. For nZVI, its effect of on the sorption affinity of DBP to the sediment was small because sorption of DBP on nZVI was similar to that on the sediment at low aqueous concentrations (0.5e5% Cs). It is known that DBP is a hydrophobic (logKow ¼ 4.45) compound and preferred to be adsorbed to a hydrophobic environment, such as organic matter (Yang et al., 2013) and clay surfaces (Liu et al., 2013). It was reported that soil adsorption capacity of DBP was positively correlated to soil organic matter content (Wu et al., 2018) and tend to adsorb to the organic matter associated with solids (e.g. soil and sediment) (Staples et al., 1997). However, it can be seen
from Table 1 that sediment organic matter had a minor role in DBP sorption. This might be due to low TOC content of the sediment (0.34%). 4.4. Influence of environmental particles on degradation of DBP in sediments In treatments without C. closterium, addition of 0.2% particle had only small impact on DBP degradation in sediments (from 1.7% increment to 4.4% decrement). With the increase of particle concentration from 0.2% to 2%, degradation percentage of DBP was more obviously inhibited by the particles added (1.3%e15.7% decrement). In general, particle effect on degradation percentage of DBP increased with the increase of particle sorption ability and particle concentration. It is known that particles with small size normally had relatively strong sorption ability for organic pollutants due to their large surface area, hence influencing degradation of the pollutants (Khorram et al., 2016). Stronger sorption to particles leads to lower bioavailability (Glomstad et al., 2016; Zhang et al., 2016; Liu et al., 2019). For example, the decrement of degradation percentage of DBP was the highest in the presence of MWNT due to its strong sorption ability for DBP. Many studies have reported that carbon nanotube amendments can largely decrease the bioavailability of hydrophobic organic pollutants by serving as a strong sorbent, thus resulting in reduced degradation of the pollutants (Xia et al., 2010). In addition, inhibition of MWNT on microbial activity also contributed to the decrease of DBP degradation as mentioned in Section 4.2. For nZVI, its inhibitory effect on DBP degradation was small in treatment with 2% particle addition. This is mainly because sorption affinity of DBP to nZVI was comparable to that of sediment sample. However, degradation percentage of DBP was more obviously inhibited in treatment with 2% PE addition (11.5% decrement) than in treatment with 2% BC400 addition (4.1% decrement) although sorption affinity of DBP to PE was similar to that of BC400. In addition to reducing bioavailability of organic pollutants, particles effect on degradation of the pollutants may occur by changing indigenous sediment microbial community (Shrestha et al., 2013). Among the dominant genera in this work, Bacillus, Rhodococcus and Pseudomonas have been reported to be capable of degrading PAEs (Gao and Wen, 2016; Yang et al., 2018). It can be seen from Table S4 that relative abundance of the dominant PAE-degrading bacteria was about 36% higher in treatment with 2% BC400 addition than in treatment with 2% PE addition. Previous studies reported that biochar contains a range of nutrients (e.g. inorganic salts) and can bring benefits for microbial growth (Joseph et al., 2013). Therefore, higher microbial activity in treatment with 2% BC400 addition led to faster degradation of DBP as compared to treatment with 2% PE addition. It has been reported that C. closterium could degrade DBP directly (Gao and Chi, 2015; Gao and Chi, 2016). Therefore, in treatments with C. closterium, DBP could be degraded not only by bacteria but also by the alga. As shown in Table S4, in treatment with sediments only, relative abundance of the dominant PAEdegrading bacteria was 6.6% in total, which was reduced by 35% in the presence of C. closterium. Meanwhile, relative abundance of C. closterium accounted for 21.6% of the total taxon read numbers including both C. closterium and bacteria. This is probably due to nutrient competition between the alga and bacteria under nutrient-limited conditions, which inhibited bacteria growth (Gao and Chi, 2016). Moreover, as shown in Table 2, the presence of C. closterium increased degradation percentage of DBP (8.9% increment). Contribution of C. closterium to DBP degradation should be much higher than the increment (i.e. 8.9%). Therefore, it can be considered that C. closterium played a key role in degradation of DBP in the alga-bacteria system, which led to the degradation
F. Zhang et al. / Environmental Pollution 261 (2020) 114228 Table 2 Degradation percentages of DBP in sediments with and without particles. Sample
DBP degradation percentages (%) Without C. closterium
Sediment 0.2% BC400 0.2% MWNT 0.2% PE 0.2% nZVI 2% BC400 2% MWNT 2% PE 2% nZVI
76.6 75.4 72.2 76.8 78.3 72.5 60.9 65.1 75.3
± ± ± ± ± ± ± ± ±
0.6B* 0.9C* 0.7D* 2.2B* 0.9A* 2.7D* 0.1F 1.7E* 2.4C*
With C. closterium 85.5 84.8 74.1 85.4 89.6 77.3 60.7 70.9 80.2
± ± ± ± ± ± ± ± ±
2.8b* 2.9b* 0.8e* 1.9b* 4.5a* 4.1d* 0.1g 1.0f* 0.1c*
Sediment represents the control in the absence of particles. BC400, PE, nZVI and MWNT represent treatments with biochar pyrolyzed at 400 C, polyethylene microplastic, nanoscale zero-valent iron and multi-walled carbon nanotube, respectively. 0.2% and 2% represent particle addition concentrations, respectively. Letters AeF and aeg indicate significant difference (p < 0.05) among different particle addition treatments without and with C. closterium, respectively. Symbol * indicates significant difference (p < 0.05) between treatments with and without C. closterium.
enhancement. A similar phenomenon was found in our previous study (Gao and Chi, 2016). Moreover, addition of 0.2% particles could influence DBP degradation mainly by promoting or inhibiting the growth of C. closterium in some cases in this work. For example, addition of 0.2% nZVI had a small effect on DBP degradation in treatment without C. closterium (1.7% increment) but a larger effect in treatment with C. closterium (4.1% increment) by increasing the biomass of C. closterium. In contrast, the decrement in treatment with 0.2% MWNT addition was 4.4% in the absence of C. closterium but 11.4% in the presence of C. closterium, mainly due to its significant inhibition on the growth of C. closterium. Besides, addition of BC400 had no obvious (p > 0.05) effects on both the biomass of C. closterium and degradation percentage of DBP, while addition of PE enhanced the biomass of C. closterium by 13% but showed no obvious (p > 0.05) impact on DBP degradation. This might be partly due to the differences in their properties (Table S1). Clearly, deeper investigation is needed to understand the mechanisms. However, with the increase of particle concentration from 0.2% to 2%, particle addition on sorption affinity of DBP to the sediment played a more significant role in its effect on DBP degradation. Compared with treatments without C. closterium, inoculation of
7
C. closterium increased degradation percentage of DBP in all treatments with particle addition (1.9e11.3% increment) except treatment with 2% MWNT addition (no effect, p > 0.05). The increment increased with the increase of chlorophyll a content in general. For example, the increment was the highest for treatment with 0.2% nZVI addition (11.3%) due to its highest promotion effect on algal growth. In contrast, the increment was the lowest for treatments with MWNT addition (no more than 1.9%). This is consistent with the strong sorption (indicated as logKf) of MWNT to DBP and its strong inhibition on the growth of C. closterium (indicated as chlorophyll a content). The increment decreased with the increase of particle concentration from 0.2% to 2% (from no effect to 5.8% increment for treatments with 2% addition). Because the biomass of C. closterium did not change much with the increase of particle concentration from 0.2% to 2%, the particles achieved more significant effect on DBP degradation by inhibiting its bioavailability. For example, degradation percentage of DBP declined with the increase of BC400 and PE concentrations added although chlorophyll a content had no obvious change (p > 0.05). Based on the above discussion, it can be considered that the particles could influence bioavailability of DBP by sorption and biomass of C. closterium, and thus degradation of DBP in sediments. As more and more environmental particles are being spread and accumulated in the environment due to their growing use, higher particle concentration might negatively impact on the ecological environment, meanwhile reduce biodegradation rate of DBP. Recently, we found that C. closterium could influence the net flux of DBP from water to sediment by changing DBP degradation in sediments, and then the total biodegradation flux in the watersediment system (Zhang et al., 2019). Hence, the reduced biodegradation rate by environmental particles probably leads to higher degradation half-life of the pollutant and thus enhancing its persistence in the coastal environment. Besides environmental particles, many factors can also influence algal growth, such as nutrient condition in the overlying water, sediment composition, and so on. Clearly, more research is needed to evaluate the environmental risk of the pollutant in the presence of marine microphytobenthos. 5. Conclusions The results in this work showed that effects of environmental
Table 3 Bacterial diversity indices, read numbers and OTUs of different sediment samples. Sample
Taxon read number
OTUs
Shannon index
Simpson index
Without C. closterium
Sediment 0.2% BC400 0.2% PE 0.2% nZVI 0.2% MWNT 2% BC400 2% PE 2% nZVI 2% MWNT
65232 64821 64804 64108 62793 63364 63775 64122 63042
532 534 498 529 586 546 517 531 567
5.354 5.370 5.298 5.304 5.936 5.540 5.333 5.394 5.534
0.923 0.923 0.922 0.920 0.954 0.939 0.924 0.933 0.950
With C. closterium
Sediment 0.2% BC400 0.2% PE 0.2% nZVI 0.2% MWNT 2% BC400 2% PE 2% nZVI 2% MWNT
52686 48028 47494 50692 42921 46827 52330 53384 36969
547 527 501 503 511 501 590 507 479
5.300 5.047 4.965 5.189 4.702 4.964 5.571 5.299 3.899
0.919 0.895 0.892 0.913 0.850 0.886 0.933 0.928 0.747
Sediment represents the control in the absence of particles. BC400, PE, nZVI and MWNT represent treatments with biochar pyrolyzed at 400 C, polyethylene microplastic, nanoscale zero-valent iron and multi-walled carbon nanotube, respectively. 0.2% and 2% represent particle addition concentrations, respectively.
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particles on the biomass of C. closterium (indicated as chlorophyll content) were obviously different among the particles, which are, inhibition by multi-walled carbon nanotube (86.4e98.9%), stimulation by polyethylene microplastic (13.0e15.9%) and nanoscale zero-valent iron (6.7e20.7%) and no effect by biochar pyrolyzed at 400 C. Besides, the results of 16s rDNA sequencing showed that C. closterium was probably the most abundant DBP-degrader, accounting for 20.0e49.3% of the total taxon read numbers. Its biomass was the most important factor influencing bacterial community structure in sediments. In treatments without C. closterium, degradation percentage of DBP decreased with the increase of particle sorption ability to DBP and particle concentration in general. In treatments with C. closterium, inoculation of C. closterium increased the degradation percentage of DBP in all treatments with particle addition, which increased with the increase of chlorophyll a content in general but decreased with the increase of particle concentration from 0.2% to 2%. Our findings suggested that the environmental particles could influence bioavailability of DBP by sorption and biomass of C. closterium, and thus degradation of DBP in sediments. Therefore, impact of the environmental particles should be considered in the environmental risk assessment of the pollutant. Declaration of competing interest We declare that we do not have any commercial or associative interest that represents a conflict of interest in connection with the work submitted. Appendix A. Supplementary data Supplementary data related to this article can be found at https://doi.org/10.1016/j.envpol.2020.114228. References Adeleye, A.S., Stevenson, L.M., Su, Y., Nisbet, R.M., Zhang, Y., Keller, A.A., 2016. Influence of phytoplankton on fate and effects of modified zerovalent iron nanoparticles. Environ. Sci. Technol. 50, 5597e5605. Alimba, C.G., Faggio, C., 2019. Microplastics in the marine environment: current trends in environmental pollution and mechanisms of toxicological profile. Environ. Toxicol. Pharmacol. 68, 61e74. Anbumani, S., Kakkar, P., 2018. Ecotoxicological effects of microplastics on biota: a review. Environ. Sci. Pollut. Res. 25, 14373e14396. Ask, J., Rowe, O., Brugel, S., Stromgren, M., Bystrom, P., Andersson, A., 2016. Importance of coastal primary production in the northern Baltic Sea. Ambio 45 (6), 635e648. Beckingham, B., Ghosh, U., 2017. Differential bioavailability of polychlorinated biphenyls associated with environmental particles: microplastic in comparison to wood, coal and biochar. Environ. Pollut. 220, 150e158. Canniff, P.M., Hoang, T.C., 2018. Microplastic ingestion by Daphnia magna and its enhancement on algal growth. Sci. Total Environ. 633, 500e507. Casabianca, S., Capellacci, S., Giacobbe, M.G., Dell’Aversano, C., Tartaglione, L., Varriale, F., Narizzano, R., Risso, F., Moretto, P., Dagnino, A., Bertolotto, R., Barbone, E., Ungaro, N., Penna, A., 2019. Plastic-associated harmful microalgal assemblages in marine environment. Environ. Pollut. 244, 617e626. Chi, J., Li, Y., Gao, J., 2019. Interaction between three marine microalgae and two phthalate acid esters. Ecotoxicol. Environ. Saf. 170, 407e411. Frias, J.P.G.L., Gago, J., Otero, V., Sobral, P., 2016. Microplastics in coastal sediments from southern Portuguese shelf waters. Mar. Environ. Res. 114, 24e30. Gao, D., Li, Z., Wang, H., Liang, H., 2018. An overview of phthalate acid ester pollution in China over the last decade: environmental occurrence and human exposure. Sci. Total Environ. 645, 1400e1409. Gao, D.W., Wen, Z.D., 2016. Phthalate esters in the environment: a critical review of their occurrence, biodegradation and removal during wastewater treatment processes. Sci. Total Environ. 541, 986e1001. Gao, J., Chi, J., 2015. Biodegradation of phthalate acid esters by different marine microalgal species. Mar. Pollut. Bull. 99 (1e2), 70e75. Gao, J., Chi, J., 2016. A new approach to biodegrading diethyl phthalate in the benthic sea-floor diatom-bacteria system. J. Saf. Environ. 16 (4), 311e314 (in Chinese). Glomstad, B., Altin, D., Sørensen, L., Liu, J., Jenssen, B.M., Booth, A.M., 2016. Carbon nanotube properties influence adsorption of phenanthrene and subsequent bioavailability and toxicity to Pseudokirchneriella subcapitata. Environ. Sci.
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