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Dissolved organic matter as a terminal electron acceptor in the microbial oxidation of steroid estrogen
Environmental Pollution 218 (2016) 26e33
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Dissolved organic matter as a terminal electron acceptor in the microbial oxidation of steroid estrogen Lipeng Gu, Bin Huang*, Zhixiang Xu, Xiaodong Ma, Xuejun Pan Faculty of Environmental Science and Engineering, Kunming University of Science and Technology, Kunming, Yunnan 650500, PR China
a r t i c l e i n f o
a b s t r a c t
Article history: Received 6 April 2016 Received in revised form 13 July 2016 Accepted 9 August 2016
Steroid estrogen in natural waters may be biodegraded by quinone-reducing bacteria, dissolved organic matter (DOM) may serve as a terminal electron acceptor in this process. The influence of temperature, pH, dissolved oxygen and light illumination on the reduction efficiency of anthraquinone-2-disulfonate (AQS) was investigated using 17b-estradiol (E2) as the target species. The optimum reduction conditions were found to be in the dark under anaerobic conditions at pH 8.0 and 30 C. Quinone-reducing bacteria can use the quinone structure of DOM components as a terminal electron acceptor coupling with microbial growth to promote biodegradation. Compared with other DOM models, AQS best stimulated E2 biodegradation and the mediating effect was improved as the AQS concentration increased from 0 to 0.5 mM. However, further increase had an inhibiting effect. Natural DOM containing lake humic acid (LHA) and lake fulvic acid (LFA) had a very important accelerating effect on the degradation of E2, the action mechanism of which was consistent with that defined using DOM models. The natural DOM contained more aromatic compounds, demonstrating their greater electron-accepting capacity and generally more effective support for microorganism growth and E2 oxidation than Aldrich humic acid (HA). These results provide a more comprehensive understanding of microbial degradation of steroid estrogens in anaerobic environments and confirm DOM as an important terminal electron acceptor in pollutant transformation. © 2016 Elsevier Ltd. All rights reserved.
Handling Editor: Maria Cristina Fossi Keywords: Quinone Bacteria Dissolved organic matter Electron acceptor 17b-estradiol Biodegradation
1. Introduction Natural and synthetic steroid estrogens (SEs) such as estrone (E1), 17b-estradiol (E2), estriol (E3) and 17a-ethynylestradiol (EE2) have been recognized as potential endocrine disruptors (Huang et al., 2013). Their incomplete removal in sewage treatment plants (STPs) and indeed direct discharge have led to their being widely detected in aquatic environments (Huang et al., 2014). Although those compounds are normally found in water at only ng per litre levels, field and laboratory studies have demonstrated that they can still alter normal hormone functions and the physiological status of wildlife (Liu et al., 2011, 2012; Huang et al., 2015). Thus, the fate and behavior of SEs in natural water environments have generated extensive concern. Among the various fates of environmental SEs, biodegradation has been identified as one of the predominant removal mechanisms from both natural water and sediment. However, it has been
* Corresponding author. E-mail address: [email protected] (B. Huang). http://dx.doi.org/10.1016/j.envpol.2016.08.028 0269-7491/© 2016 Elsevier Ltd. All rights reserved.
shown that SE biodegradation by heterotrophic bacteria is very slow. SEs have been reported to have a biodegradation half-life of 20e40 days under aerobic conditions (Clouzot et al., 2008), and longer could be expected under anaerobic conditions (Ying et al., 2003). According to previous studies in our laboratory, even when SEs are continuously discharged into surface water from STPs, their accumulation is not very obvious (Huang et al., 2013). Therefore, it is speculated that some active substances in natural aquatic environments enhance the biodegradation of SEs. Many studies have shown that the fast degradation of pollutants in natural waters can usually be attributed to the accelerating effects of dissolved organic matter (DOM). Quinone groups in DOM can promote microbes to degrade heavy metals (Reijonen et al., 2016) and organic pollutants (Guha et al., 2001). For instance, such groups have recently been reported to play an active role in the de-colorization of azo dyes (Meng et al., 2014) and in the reduction of Cr(VI) (Guo et al., 2012; Huang et al., 2016) under anaerobic conditions. But those studies have been restricted to considering organic acids as electron donors. There are no published reports regarding DOM as a terminal electron acceptor acting
L. Gu et al. / Environmental Pollution 218 (2016) 26e33
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to accelerate SEs’ biodegradation by quinone-reducing bacteria in aquatic environments. That was the phenomenon explored in this study. Quinone-reducing bacteria have attracted much interest for their respiration diversity, their wide distribution in diverse environments and their potential application in pollution bioremediation through biodegradation. Recently, extensive studies have assessed the catalytic effect of redox mediators on the reduction of heavy metals and organic contaminants by quinone-reducing bacteria isolated from terrestrial or freshwater sources (Cao et al., 2013; Pearce et al., 2006). Recent studies in our own laboratory have demonstrated for the first time their ability to degrade E2 under alkaline conditions. Both DOM and DOM models were considered as terminal electron acceptors in that work. The testing of DOM derived from natural water was a new attempt to mediate E2 biodegradation, and no such results have been published so far. The aim of this study was to test the mediating mechanism of four models of DOM and two actual samples of DOM on the microbial degradation of E2. The influences of temperature, pH, dissolved oxygen and light illumination on the reduction efficiency of the DOM models were assessed using sodium formate as an electron donor. Also investigated were the ability of quinone-reducing bacteria to accelerate DOM reduction and E2 biodegradation, the coupling of E2 as the electron donor and DOM as a terminal electron acceptor to support microbial growth, as well as the effects of DOM concentration on the anaerobic microbial oxidation of E2.
5 g L1 of yeast extract and 10 g L1 of NaCl at 30 C on a rotary shaker and agitated at 160 rpm for 24 h. Subsequently, the strains were separated from the medium by centrifugation at 10,000 rpm for 10 min. The harvested biomass was washed three times with buffer solution before being employed to the experiments. Standard anaerobic culturing techniques were used throughout the study. The base medium was modified from that recommended by Lovley and Phillips, 1988. It contained (in mg L1): NaCl, 1000; NH4Cl, 800; KH2PO4, 500; K2HPO4, 600; MgCl2, 200; CaCl2$2H2O, 50; and yeast extract, 10.
2. Materials and methods
2.4. E2 degradation with model and natural DOM
2.1. Reagents and chemicals
In order to explore the mediating mechanisms of the four DOM models (AQS, AQC, LQ and JQ) and the natural DOM (LHA, LFA and HA) in the microbial degradation of E2, batch experiments were performed assessing: (1) AQS reduction coupling with the oxidation of E2 to support microbial growth; (2) Any effects of the DOM models and their concentrations on the microbial degradation of E2; (3) DOM-mediated microbial degradation of E2; (4) The reduction of DOM by quinone-reducing bacteria with E2 as the electron donor; and (5) The effects of DOM concentration on E2 biodegradation. The experiments were conducted in rubber-sealed 100 mL serum bottles. Prior to use, all of the materials, including the serum bottles, the sealing rubber and the solutions were sterilized in an autoclave at 121 C for 25 min. The DOM model experiments were performed in the dark with 100 mL of MSM that contained different concentrations of AQS, AQC, LQ and JQ (from 0 to 2 mmol L1), different concentrations of E2 (0.5, 1, 1.5 and 2 mg L1), and 0.1 g L1 of quinone-reducing bacteria under anaerobic conditions at pH 8.0 and 30 C for 120 h. The natural DOM experiments were performed with different concentrations of LHA, LFA and HA (from 0 to 8 mgC L1), 1 mg L1 of E2, and 0.1 g L1 of quinone-reducing bacteria under the same conditions.
Anthraquinone-2-sodium sulfonate (AQS), anthraquinone-2carboxylate (AQC), 5-hydroxy-1,4-naphthoquinone (juglone, JQ), 2-hydroxy-1,4-naphthoquinone (lawsone, LQ), E2 and Aldrich humic acid (HA) were supplied by Sigma-Aldrich. The E2 used was high performance liquid chromatograph grade. Its key physicochemical properties are shown in Table 1 (Lai et al., 2000). The natural DOM was extracted from sediment collected from ErHai Lake in Yunnan Province of China. It was sieved using a 4.0 mm sieve and rinsed with deionized water. Lake humic acid (LHA) and lake fulvic acid (LFA) were extracted using the alkali-acid method of the International Humic Substances Society. The HA, LHA and LFA stock solutions were stored in polyethylene containers at 4 C in the dark and used within 3 weeks. All of the other reagents are analytical grade if not otherwise mentioned. 2.2. Bacteria and culturing conditions The quinone-reducing bacteria were also enriched and isolated from the ErHai Lake sediment. The enriched mineral salts medium (MSM) was supplemented with AQS (1 mM) as an electron acceptor and sodium formate (5 mM) as a substrate. To isolate quinonereducing bacteria from the enriched MSM, the enrich MSM was diluted serially and incubated on agar plates containing AQS (1 mM) and sodium formate (5 mM). Selected well-developed colonies were streaked three times with new agar and then preserved for further study. The quinone-reducing bacteria were inoculated into a medium, which contained 10 g L1 of tryptones,
2.3. Reducing characteristics of the AQS Serum bottles containing 100 mL MSM, 0.5 mM AQS, 5 mM sodium formate and 0.1 g L1 of quinone-reducing bacteria were agitated on a rotary shaker at 160 rpm for 120 h to determine the conditions delivering the best AQS reduction efficiency. The effects of temperature, pH, dissolved oxygen and illumination conditions were assessed. The temperatures tested were 0, 5, 10, 15, 25, 30, 35 and 45 C conducted at pH 7.0, and the pH values were 4.0, 5.0, 6.0, 7.0, 8.0, 9.0, 10.0 and 11.0 conducted at 30 C. The dissolved oxygen levels were 0, 0.5, 1.0, 2.0 and 3.0 mg L1 and agitation under natural and ultraviolet light and in the dark were tested. Sterile controls were tested under the same conditions. They were prepared by autoclaving at 121 C for 25 min. Each treatment was set up in triplicate.
2.5. Analytical methods and data analysis AQS absorbance was measured at 336 nm using a UVeVis spectrophotometer, and the reduction efficiency of the AQS was then calculated using the change in AQS absorbance as follows.
The reduction efficiency of AQSð%Þ ¼ ðA0 At Þ=A0 100%
(1)
Table 1 Selected physiochemical properties of the 17b-estradiol. Estrogen
Molecular formula
Molecular weight (g$mol1)
Water solubility at 20 C (mg L1)
logKow
pka
17b-estradiol (E2)
C18H24O2
272.4
3.9e13.3
3.1e4.0
10.5e10.7
28
L. Gu et al. / Environmental Pollution 218 (2016) 26e33
where At is the absorbance of AQS after the biodegradation experiment and A0 is the initial absorbance. The amount of reduced DOM (DOMred) was determined by adding Fe(III) (Lovley et al., 1996). In this process, DOMred was quickly oxidized to DOM by Fe(III), at the same time, Fe(III) was reduced to Fe(II). As Fe(II) was continuously being produced during the interaction of DOMred and Fe(III), the Fe(II) concentration was recorded at selected time points using the 1,10 phenanthroline chemical method (Amonette and Templeton, 1998). The bacterial growth was monitored by measuring the protein concentration with Bradford's method using bovine serum albumin as the standard (Bradford, 1976). The concentration of E2 was quantified using a high performance liquid chromatograph (Agilent Technologies 1260) equipped with a Waters symmetry-C18 reversed-phase column (5 mm, 4.6 250 mm) and a fluorescence detector. The E2 was eluted by the isocratic mixture mobile phase at a rate of 1 mL min1 and detected at the excitation and emission wavelengths of 283 and 345 nm. The mobile phase consisted of acetonitrile and ultrapure water in a ratio of 60:40 containing 0.1% trifluoroacetic acid. The detection limit for E2 was determined to be 0.01 mg L1, and the relative standard deviation for all samples was within 5%. All of the experiments were conducted in triplicate. Statistical analysis was performed using version 20.0 of SPSS for Windows. All of the data were expressed as the mean ± the standard error of the mean. 3. Results and discussion 3.1. AQS reduction Temperature strongly influences microbial growth and thus the reduction efficiency of DOM. If the temperature exceeds the
adaptation range of the microbial enzymes involved, it can induce changes in the charge of the microorganisms' plasma membranes (Stoffels et al., 2008). In these experiments the reduction efficiency of AQS increased with a rise in temperature between 5 and 30 C (Fig. 1-A). Higher temperatures then reduced the reduction efficiency. Fig. 1-A shows that the reduction efficiency of AQS after 120 h was not much different at 25 or 30 C, though the reduction efficiency at 30 C was better than at 25 C between 5 and 60 h. So the optimum temperature was 30 C. Quinone-reducing bacteria at that temperature are at their most effective. The hydrogen ion concentration in MSM greatly influences the bacterial growth, since pH limits enzyme activity (Bhalla et al., 2013; Fenner and Freeman, 2011). In order to determine the optimal pH at which quinone-reducing bacteria reduce AQS, different pH conditions (4.0, 5.0, 6.0, 7.0, 8.0, 9.0, 10.0 and 11.0) were investigated. The relationship between AQS reduction efficiency and pH is shown in Fig. 1-B. The rate of AQS reduction increased quickly when the pH of the MSM was increased from 5.0 to 8.0. A high AQS reduction rate was observed at pH 7.0 or 8.0, but it decreased again at pH values of 9.0 or more. Quinone-reducing bacteria prefer alkaline and neutral environments. Within the pH range of 7.0e9.0 the bacteria exhibited better efficiency in AQS reduction than in more acidic conditions. One reason might be that ErHai Lake is an alkaline lake. However, at pH 10.0 or pH < 5.0, there was no detectable AQS reduction (AH2QS) in the MSM. The quinone-reducing bacteria had apparently been irreversible poisoned by Hþ or OH in overly acidic or alkaline conditions. Previous research has shown that dissolved oxygen is very important to microbial metabolism of organic contaminants in an aquatic environment (Larcher and Yargeau, 2013; Tran et al., 2013). The reduction efficiency of AQS at various dissolved oxygen concentrations is shown in Fig. 1-C. After 10 h there was no significant difference in AQS reduction efficiency among the various dissolved
Fig. 1. Reduction characteristics of AQS (0.5 mM) under different conditions. (A): Temperature (conducted at pH 7.0); (B): pH value (conducted at 30 C); (C): oxygen availability; (D): light conditions. Error bars represent standard deviation of the mean (n ¼ 3).
L. Gu et al. / Environmental Pollution 218 (2016) 26e33
29
oxygen conditions tested (0.5e3.0 mg L1). But in the range of 20e80 h it was dramatically enhanced at lower dissolved oxygen concentrations, and anaerobic was the best condition. Subsequently, after 120 h, the effect of dissolved oxygen on AQS reduction efficiency was unremarkable. One likely explanation is that the AH2QS produced (through reduction of AQS) was oxidized by dissolved oxygen in the MSM. Accumulated AH2QS began to appear in the reaction system only when dissolved oxygen was near to being exhausted (Li et al., 2012; Semblante et al., 2015). Therefore, quinone-reducing bacteria are more effective in anaerobic conditions. The influence of light on AQS reduction efficiency is shown in Fig. 1-D. The reduction under ultraviolet light was only 6.51% after 5 h and there were almost no subsequent changes, probably because the quinone-reducing bacteria had been killed by the ultraviolet light after 5 h. The AQS reduction in the dark was more complete than under natural light. But the efficiency was almost the same in the dark and under natural light between 0 and 10 h. The data show, however, that the reduction efficiency in the dark gradually gained an advantage from 10 to 80 h. Two major reasons suggest themselves. Quinone-reducing bacteria isolated from ErHai Lake sediment have never experienced natural light exposure. And at the same time, light is likely to affect microbial enzyme metabolism and inhibit the growth of many microorganisms (Sivonen, 1990). The quinone-reducing bacteria tested seem to have more applicability in dark condition.
3.2. Microbial oxidation of E2 with DOM models 3.2.1. AQS reduction with E2 as the electron donor Previous research has shown that electron transfer is influenced by the quinone structure in model DOM, such as AQS, AQC and anthraquinone-2-6-sulfonate (AQDS) (Wu et al., 2014). Recent studies have demonstrated that the mechanism of quinone structure involved pollutant transformation was along with the cycles between the oxidation of carbonyl structure and the reduction of hydroxyl quinone (Lovley et al., 1996). Applying the optimum conditions previously determined, all of the following experiments were carried out in the dark under anaerobic conditions at pH 8.0 and 30 C. The reduction of model DOM (take AQS as examples) by quinone-reducing bacteria was investigated with E2 as the electron donor. Fig. 2-A shows that almost no AH2QS could be detected when only AQS and E2 were added to the MSM without adding quinone-reducing bacteria. The addition of quinone-reducing bacteria after 20 h produced a small amount of AH2QS even without E2 as an electron donor. This is probably because the medium contained trace amounts of yeast extract, which could enhance quinone-reducing bacteria growth from 0 to 10 h (Fig. 2-B), resulting in some AQS reduction. However, experimental results show that adding E2 as electron donor significantly improved the reduction efficiency. It eventually reached about 46% after 120 h (Fig. 2-A). Fig. 2-B indicates that after 120 h the microbial growth with E2 was 3.01 times as much as without it. This shows that quinone-reducing bacteria can use E2 as an electron donor to enhance the reducing efficiency of AQS and their own growth.
3.2.2. Effects of DOM models on E2 biodegradation A number of kinetic models have been applied to describe the biodegradation of organic pollutants, with pseudo-first-order kinetics frequently used as a convenient way to describe the biodegradation progress of xenobiotic compounds (Nyholm et al., 1996). The biodegradation of E2 was assumed at the outset to follow pseudo-first-order kinetics.
Fig. 2. AQS reduction coupling with the oxidation of E2 to support microbial growth. (A): The reduction efficiency of AQS; (B): microbial growth (OD value). The experiments were performed in the dark under anaerobic conditions at pH 8.0 and 30 C for 120 h. The initial concentration of E2 was 1 mg L1. The error bars represent the standard deviation of the mean (n ¼ 3).
ln C=C0 ¼ kt þ b
(2)
here C is the concentration of E2 after a period of biodegradation; C0 is the initial concentration of E2 in the MSM; k (h1) is the degradation rate constant; t (h) is the degradation time; and b is a fitting parameter. When b ¼ 0, Eq. (2) can be expressed as a first-order kinetic equation.
ln C=C0 ¼ k0 $t
(3)
The half-life (T1/2) of E2 is then:
T1=2 ¼ ln 2=k0
(4)
Although some studies have shown that the presence of AQDS can improve the reduction performance of Shewanella with organic contaminants (Guha et al., 2001) and Cr(VI) (Reijonen et al., 2016; Huang et al., 2016), no information has previously been published about the effects of quinone compounds on E2 biodegradation. Fig. 3 demonstrates that AQS best stimulated the biodegradation of E2 at AQS concentrations from 0 to 0.5 mM and that the mediation effect improved with increasing AQS concentration over that range.
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L. Gu et al. / Environmental Pollution 218 (2016) 26e33
Fig. 3. Effect of DOM models and their concentrations on the microbial degradation of E2. The experiments were performed in the dark under anaerobic conditions at pH 8.0 and 30 C for 120 h. The initial concentration of E2 was 1.0 mg L1. The error bars represent the standard deviation of the mean (n ¼ 3).
Adding 0.5 mM AQS increased the k value by almost 50% (from 0.00197 h1 to 0.00295 h1). Adding 0.2 mM AQC resulted in a 34% increase. However, greater AQS and AQC concentrations demonstrated no further enhancement of E2 biodegradation. Improved E2 biodegradation was also observed with the addition of LQ and JQ at concentrations up to 0.1 mM. When the concentration of these DOM models was greater, inhibited E2 biodegradation was observed. The inhibition might be attributed to their toxicity to bacterial cells (Lown, 1983). The constants of the relevant kinetic equations are presented in Table 2. The results show that increasing E2 concentration would give rise to slower biodegradation, whether or not the optimum concentration of the DOM model is present. With the E2 concentration increased from 0.5 to 2 mg L1 in MSM without a model DOM, the half-life of E2 is enhanced from approximately 290 to 436 h. However, adding the optimum concentration of a DOM model can reduce the half-life significantly. With 0.5 mM AQS it becomes 218e288 h. With 0.2 mM AQC it was 239e318 h 0.1 mM
Fig. 4. DOM-mediated microbial degradation of E2. The experiments were performed in the dark under anaerobic conditions at pH 8.0 and 30 C for 120 h. The concentration of LHA, LFA and HA was 3.0 mgC L1. The initial concentration of E2 was 1.0 mg L1. Error bars represent standard deviation of the mean (n ¼ 3).
LQ reduced it to 270e358 h. And 0.1 mM JQ yielded a half-life of 274e371 h (Table 2). These results further illustrate the extent to which quinones can improve the biodegradation of E2. 3.3. DOM-mediated microbial degradation of E2 in an alkaline environment Previous studies have reported that DOM models can act electron shuttle mediators to remove toxic heavy metals (Wang et al., 2011) and organic contaminants (Meng et al., 2014). However, no report of actual DOM mediating biodegradation of E2 has been published so far. The results presented in Fig. 4 indicate that LHA, LFA or HA alone existed in the MSM had a positive effect on the biodegradation of E2 in a sterile environment at the beginning. The main reason is perhaps that DOM has plenty of hydrophobic functional groups that could interact with steroid endocrine disruptors (Zhu et al., 2012). After 120 h, the microbial degradation of
Table 2 Rate of E2 biodegradation by quinone-reducing bacteria with DOM models. DOM models
Concentration of DOM model (mmol L1)
Initial concentration of E2 (mg L1)
Rate constant (h1)
T1/2 (h)
R2
None
0
0.5 1.0 1.5 2.0
0.00239 0.00213 0.00186 0.00159
290.02 325.42 372.66 435.94
0.9281 0.9332 0.9447 0.9406
AQS
0.5
0.5 1.0 1.5 2.0
0.00318 0.00292 0.00267 0.00241
217.97 237.38 259.61 287.61
0.9452 0.9687 0.9512 0.9496
AQC
0.2
0.5 1.0 1.5 2.0
0.00290 0.00264 0.00241 0.00218
239.02 262.56 287.61 317.96
0.9367 0.9442 0.9531 0.9621
LQ
0.1
0.5 1.0 1.5 2.0
0.00257 0.00236 0.00215 0.00194
269.71 293.71 322.39 357.29
0.9335 0.9531 0.9504 0.9448
JQ
0.1
0.5 1.0 1.5 2.0
0.00253 0.00231 0.00209 0.00187
273.97 300.06 331.65 370.67
0.9451 0.9578 0.9483 0.9517
L. Gu et al. / Environmental Pollution 218 (2016) 26e33
E2 was about 23% without DOM, but with the addition of LHA it was 38%, with LFA 43% and with HA 30% after 120 h. So LHA and HFA can more effectively promote microbial degradation of E2. These results suggest that the quinone structure of DOM plays a very important
31
role in promoting the degradation of E2 by quinone-reducing bacteria, and that the mechanism is likely to be consistent with that of DOM models. Meanwhile, coupling E2 biodegradation with DOM reduction can accelerate microbial growth. The data of Fig. 4 suggest that the mediating ability of LHA, LFA and HA is in the order LFA > LHA > HA. The amount of reduced DOM (DOMred) was determined by adding Fe(III) and determining the extent of Fe(II) production (Lovley et al., 1996). Fig. 5 demonstrates that DOM can directly restore Fe(III) without the presence of microorganisms, probably because DOM is a mixture of macromolecules (Schmidt et al., 2011) and possesses trace amounts of reducing substances (He et al., 2014). In addition, DOM is very complex and it contains a small amount of beneficial nutrients (Thurman, 1985) which can accelerate the growth of microorganisms. For example, a little of the protein co-substrate in MSM could be utilized by quinone-reducing bacteria to promote the reduction of Fe(III) (Hur et al., 2011). Unfortunately, however, the protein content of DOM is in general much too low to support the growth of quinone-reducing bacteria. But Fig. 5 shows that adding a certain amount of E2 as an electron donor can effectively promote the reduction of DOM. After 120 h, the DOM was significantly reduced in the active treatments (E2 þ DOM þ quinone-reducing bacteria), and the LHAred, LFAred or HAred concentration reached 0.172 mM, 0.211 mM or 0.091 mM, respectively. These results indicate that quinone groups in actual DOM extracts can be terminal electron acceptors coupling with microbial growth to promote the biodegradation of E2. 3.4. DOM concentration The results presented in Fig. 6 and Table 3 indicate that LFA best promoted E2 biodegradation, and that its mediating effect was enhanced as the LFA concentration increased from 0 to 2 mgC L1. Compared to that obtained in the absence of DOM (0.00197 h1), the k value increased by 78% (to 0.00351 h1) when 2 mgC L1 of LFA was present. The LFA was reduced to 0.234 ± 0.001 mM along with the OD growing to 1.025 ± 0.012. Adding 3 mgC L1 of LHA gave a 55% increase and 5 mgC L1 of HA produced a 22% increase in the k value. LHA and HA were reduced to 0.172 ± 0.003 mM and 0.104 ± 0.002 mM respectively, coupled with optical densities as high as 0.781 ± 0.010 for LHA and 0.683 ± 0.010 for HA (Fig. 6 and
Fig. 5. The reduction of DOM by quinone-reducing bacteria with E2 as the electron donor. (A) LHA (3.0 mgC L1); (B) LFA (3.0 mgC L1); (C) HA (3.0 mgC L1). The experiments were performed in the dark under anaerobic conditions at pH 8.0 and 30 C for 120 h. The initial concentration of E2 was 1.0 mg L1. The error bars represent the standard deviation of the mean (n ¼ 3).
Fig. 6. Effects of DOM concentration on E2 biodegradation. The experiments were performed in the dark under anaerobic conditions at pH 8.0 and 30 C for 120 h. The initial concentration of E2 was 1.0 mg L1. The error bars represent the standard deviation of the mean (n ¼ 3).
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L. Gu et al. / Environmental Pollution 218 (2016) 26e33
Table 3 Effects of DOM concentration on DOM reduction and microorganism growth. Initial concentration of DOM (mgC L1)
HA LHAred (mmol L1)
OD
0 0.2 0.5 1 2 3 4 5 8
0 0.023 0.065 0.116 0.144 0.172 0.157 0.124 0.048
0.447 0.449 0.496 0.584 0.703 0.781 0.752 0.646 0.573
FA
± ± ± ± ± ± ± ±
0.001 0.001 0.002 0.001 0.003 0.003 0.001 0.002
± ± ± ± ± ± ± ± ±
0.011 0.012 0.014 0.012 0.013 0.010 0.009 0.016 0.012
PHA
LFAred (mmol L1)
OD
0 0.056 0.137 0.178 0.234 0.211 0.161 0.142 0.097
0.447 0.549 0.737 0.921 1.025 0.937 0.821 0.756 0.663
± ± ± ± ± ± ± ±
0.003 0.002 0.002 0.001 0.005 0.001 0.003 0.003
± ± ± ± ± ± ± ± ±
0.011 0.014 0.011 0.016 0.012 0.022 0.023 0.019 0.014
HAred (mmol L1)
OD
0 0.022 0.046 0.073 0.081 0.091 0.096 0.104 0.066
0.447 0.457 0.478 0.526 0.583 0.594 0.602 0.683 0.501
± ± ± ± ± ± ± ±
0.001 0.001 0.001 0.004 0.002 0.001 0.002 0.003
± ± ± ± ± ± ± ± ±
0.011 0.012 0.018 0.013 0.011 0.014 0.018 0.010 0.014
Note: ± error represents the standard deviation of the mean of 3 replicates.
Table 3). However, the continuously increasing DOM concentration would restrain the growth of microorganisms and the biodegradation process. These results are consistent with the findings of the studies with the DOM models. The better stimulating affect observed with LHA and LFA, which tend to be more aromatic, demonstrates their greater electron-accepting capacity (O'Loughlin, 2008) and generally more effective support for the growth of microorganisms than HA. 4. Conclusion Quinone-reducing bacteria from ErHai Lake sediment gave their best reduction performance with AQS. The optimal pH value was 8.0 and the optimal temperature was at 30 C. Additionally, those quinone-reducing bacteria work best under anaerobic condition and in the dark. Quinone structures in DOM apparently function as terminal electron acceptors and couple with microbial growth to promote the biodegradation of E2. The optimum DOM concentration can reduce the half-life of E2 substantially, however too high DOM concentration can restrain microorganism growth and E2 biodegradation. The more aromatic components of DOM have greater electron-accepting capacity and are generally more effective support for the growth of microorganisms. These results go some way to explaining the observed biodegradation of steroid estrogens in alkaline environments containing DOM. Acknowledgments This project was sponsored by the National Natural Science Foundation of China (grant no. 41401558) and the China Postdoctoral Science Foundation (grant no. 2014T70887). References Amonette, J.E., Templeton, J.C., 1998. Improvements to the quantitative assay of nonrefractory minerals for Fe (II) and total Fe using 1, 10-phenanthroline. Clay. Clay Miner 46 (1), 51e62. Bhalla, A., Bansal, N., Kumar, S., Bischoff, K.M., Sani, R.K., 2013. Improved lignocellulose conversion to biofuels with thermophilic bacteria and thermostable enzymes. Bioresour. Technol. 128, 751e759. Bradford, M.M., 1976. A rapid and sensitive method for the quantitation of microgram quantities of protein utilizing the principle of protein-dye binding. Anal. Biochem. 72, 248e254. Cao, D.M., Xiao, X., Wu, Y.M., Ma, X.B., Wang, M.N., Wu, Y.Y., Du, D.L., 2013. Role of electricity production in the anaerobic decolorization of dye mixture by exoelectrogenic bacterium Shewanella oneidensis MR-1. Bioresour. Technol. 136, 176e181. Clouzot, L., Marrot, B., Doumenq, P., Roche, N., 2008. 17a-ethinylestradiol: an endocrine disrupter of great concern. Analytical methods and removal processes applied to water purification. a review. Environ. Prog. 27, 383e396. Fenner, N., Freeman, C., 2011. Drought-induced carbon loss in peatlands. Nat. Geosci. 4, 895e900. Guha, H., Jayachandran, K., Maurrasse, F., 2001. Kinetics of chromium (VI) reduction by a type strain Shewanella alga under different growth conditions. Environ.
Pollut. 115, 209e218. Guo, J.B., Lian, J., Xu, Z.F., Xi, Z.H., Yang, J.L., Jefferson, W., Liu, C., Li, Z.X., Yue, L., 2012. Reduction of Cr (VI) by Escherichia coli BL21 in the presence of redox mediators. Bioresour. Technol. 123, 713e716. He, X.S., Xi, B.D., Cui, D.Y., Liu, Y., Tan, W.B., Pan, H.W., Li, D., 2014. Influence of chemical and structural evolution of dissolved organic matter on electron transfer capacity during composting. J. Hazard. Mater. 268, 256e263. Huang, B., Wang, B., Ren, D., Jin, W., Liu, J.L., Peng, J.H., Pan, X.J., 2013. Occurrence, removal and bioaccumulation of steroid estrogens in Dianchi lake catchment, China. Environ. Int. 59, 262e273. Huang, B., Li, X.M., Sun, W.W., Ren, D., Li, X., Li, X.N., Liu, Y., Li, Q., Pan, X.J., 2014. Occurrence, removal, and fate of progestogens, androgens, estrogens, and phenols in six sewage treatment plants around Dianchi lake in China. Environ. Sci. Pollut. R. 21 (22), 12898e12908. Huang, B., Sun, W.W., Li, X.M., Liu, J.L., Li, Q., Wang, R.M., Pan, X.J., 2015. Effects and bioaccumulation of 17b-estradiol and 17a-ethynylestradiol following long-term exposure in crucian carp. Ecotoxicol. Environ. Safe. 112, 169e176. Huang, B., Gu, L.P., He, H., Xu, Z.H., Pan, X.J., 2016. Enhanced biotic and abiotic transformation of Cr(VI) by quinone-reducing bacteria/dissolved organic matters/Fe(III) in anaerobic environment. Environ. Sci. Proc. Imp. http://dx.doi.org/ 10.1039/C6EM00229C. Hur, J., Lee, B., Shin, H., 2011. Microbial degradation of dissolved organic matter (DOM) and its influence on phenanthrene-DOM interactions. Chemosphere 85, 1360e1367. Lai, K.M., Johnson, K.L., Scrimshaw, M.D., Lester, J.N., 2000. Binding of waterborne steroid estrogens to solid phases in river and estuarine systems. Environ. Sci. Technol. 34, 3890e3894. Larcher, S., Yargeau, V., 2013. The effect of ozone on the biodegradation of 17aethinylestradiol and sulfamethoxazole by mixed bacterial cultures. Appl. Microbiol. Biotechnol. 97, 2201e2210. Li, Z.T., Nandakumar, R., Madayiputhiya, N., Li, X., 2012. Proteomic analysis of 17bestradiol degradation by Stenotrophomonas maltophilia. Environ. Sci. Technol. 46, 5947e5955. Liu, J.L., Wang, R.M., Huang, B., Lin, C., Wang, Y., Pan, X.J., 2011. Distribution and bioaccumulation of steroidal and phenolic endocrine disrupting chemicals in wild fish species from Dianchi lake, China. Environ. Pollut. 159 (10), 2815e2822. Liu, J.L., Wang, R.M., Huang, B., Lin, C., Zhou, J.L., Pan, X.J., 2012. Biological effects and bioaccumulation of steroidal and phenolic endocrine disrupting chemicals in high-back crucian carp exposed to wastewater treatment plant effluents. Environ. Pollut. 162, 325e331. Lovley, D.R., Coates, J.D., Blunt-Harris, E.L., Phillips, E.J., Woodward, J.C., 1996. Humic substances as electron acceptors for microbial respiration. Nature 382, 445e448. Lovley, D.R., Phillips, E.J., 1988. Novel mode of microbial energy metabolism: organic carbon oxidation coupled to dissimilatory reduction of iron or manganese. Appl. Environ. Microbiol. 54, 1472e1480. Lown, J.W., 1983. The mechanism of action of quinone antibiotics. Mol. Cell. Biochem. 55, 17e40. Meng, X.M., Liu, G.F., Zhou, J.T., Fu, Q.S., 2014. Effects of redox mediators on azo dye decolorization by Shewanella algae under saline conditions. Bioresour. Technol. 151, 63e68. Nyholm, N., Ingerslev, F., Berg, U., Pedersen, J., Frimer-Larsen, H., 1996. Estimation of kinetic rate constants for biodegradation of chemicals in activated sludge wastewater treatment plants using short term batch experiments and mg L-1 range spiked concentrations. Chemosphere 33, 851e864. O'Loughlin, E.J., 2008. Effects of electron transfer mediators on the bioreduction of lepidocrocite (c-FeOOH) by Shewanella putrefaciens CN32. Environ. Sci. Technol. 42, 6876e6882. Pearce, C.I., Christie, R., Boothman, C., von Canstein, H., Guthrie, J.T., Lloyd, J.R., 2006. Reactive azo dye reduction by Shewanella strain J18 143. Biotechnol. Bioeng. 95, 692e703. Reijonen, I., Metzler, M., Hartikainen, H., 2016. Impact of soil pH and organic matter on the chemical bioavailability of vanadium species: the underlying basis for risk assessment. Environ. Pollut. 210, 371e379. Schmidt, M.W., Torn, M.S., Abiven, S., Dittmar, T., Guggenberger, G., Janssens, I.A.,
L. Gu et al. / Environmental Pollution 218 (2016) 26e33 €gel-Knabner, I., Lehmann, J., Manning, D.A., 2011. Persistence of Kleber, M., Ko soil organic matter as an ecosystem property. Nature 478, 49e56. Semblante, G.U., Hai, F.I., Huang, X., Ball, A.S., Price, W.E., Nghiem, L.D., 2015. Hazardous trace organic contaminants in biosolids: impact of conventional wastewater and sludge processing technologies and emerging alternatives. J. Hazard. Mater. 300, 1e17. Sivonen, K., 1990. Effects of light, temperature, nitrate, orthophosphate, and bacteria on growth of and hepatotoxin production by Oscillatoria agardhii strains. Appl. Environ. Microbiol. 56, 2658e2666. Stoffels, E., Sakiyama, Y., Grave, D.B., 2008. Cold atmospheric plasma: charged species and their interactions with cells and tissues. IEEE Trans. Plasma Sci. 36, 1441e1457. Thurman, E.M., 1985. Amount of organic carbon in natural waters[M]. In: Organic Geochemistry of Natural Waters. Springer, Netherlands, pp. 7e65. Tran, N.H., Urase, T., Ngo, H.H., Hu, J., Ong, S.L., 2013. Insight into metabolic and
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cometabolic activities of autotrophic and heterotrophic microorganisms in the biodegradation of emerging trace organic contaminants. Bioresour. Technol. 146, 721e731. Wang, X.J., Liu, G.F., Zhou, J.T., Wang, J., Jin, R.F., Lu, H., 2011. Quinone-mediated reduction of selenite and tellurite by Escherichia coli. Bioresour. Technol. 102, 3268e3271. Wu, Y.D., Liu, T.X., Li, X.M., Li, F.B., 2014. Exogenous electron shuttle-mediated extracellular electron transfer of Shewanella putrefaciens 200: electrochemical parameters and thermodynamics. Environ. Sci. Technol. 48, 9306e9314. Ying, G., Kookana, R.S., Dillon, P., 2003. Sorption and degradation of selected five endocrine disrupting chemicals in aquifer material. Water Res. 37, 3785e3791. Zhu, F.D., Choo, K.H., Chang, H.S., Lee, B., 2012. Interaction of bisphenol A with dissolved organic matter in extractive and adsorptive removal processes. Chemosphere 87, 857e864.
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Environmental Pollution journal homepage: www.elsevier.com/locate/envpol
Dissolved organic matter as a terminal electron acceptor in the microbial oxidation of steroid estrogen Lipeng Gu, Bin Huang*, Zhixiang Xu, Xiaodong Ma, Xuejun Pan Faculty of Environmental Science and Engineering, Kunming University of Science and Technology, Kunming, Yunnan 650500, PR China
a r t i c l e i n f o
a b s t r a c t
Article history: Received 6 April 2016 Received in revised form 13 July 2016 Accepted 9 August 2016
Steroid estrogen in natural waters may be biodegraded by quinone-reducing bacteria, dissolved organic matter (DOM) may serve as a terminal electron acceptor in this process. The influence of temperature, pH, dissolved oxygen and light illumination on the reduction efficiency of anthraquinone-2-disulfonate (AQS) was investigated using 17b-estradiol (E2) as the target species. The optimum reduction conditions were found to be in the dark under anaerobic conditions at pH 8.0 and 30 C. Quinone-reducing bacteria can use the quinone structure of DOM components as a terminal electron acceptor coupling with microbial growth to promote biodegradation. Compared with other DOM models, AQS best stimulated E2 biodegradation and the mediating effect was improved as the AQS concentration increased from 0 to 0.5 mM. However, further increase had an inhibiting effect. Natural DOM containing lake humic acid (LHA) and lake fulvic acid (LFA) had a very important accelerating effect on the degradation of E2, the action mechanism of which was consistent with that defined using DOM models. The natural DOM contained more aromatic compounds, demonstrating their greater electron-accepting capacity and generally more effective support for microorganism growth and E2 oxidation than Aldrich humic acid (HA). These results provide a more comprehensive understanding of microbial degradation of steroid estrogens in anaerobic environments and confirm DOM as an important terminal electron acceptor in pollutant transformation. © 2016 Elsevier Ltd. All rights reserved.
Handling Editor: Maria Cristina Fossi Keywords: Quinone Bacteria Dissolved organic matter Electron acceptor 17b-estradiol Biodegradation
1. Introduction Natural and synthetic steroid estrogens (SEs) such as estrone (E1), 17b-estradiol (E2), estriol (E3) and 17a-ethynylestradiol (EE2) have been recognized as potential endocrine disruptors (Huang et al., 2013). Their incomplete removal in sewage treatment plants (STPs) and indeed direct discharge have led to their being widely detected in aquatic environments (Huang et al., 2014). Although those compounds are normally found in water at only ng per litre levels, field and laboratory studies have demonstrated that they can still alter normal hormone functions and the physiological status of wildlife (Liu et al., 2011, 2012; Huang et al., 2015). Thus, the fate and behavior of SEs in natural water environments have generated extensive concern. Among the various fates of environmental SEs, biodegradation has been identified as one of the predominant removal mechanisms from both natural water and sediment. However, it has been
* Corresponding author. E-mail address: [email protected] (B. Huang). http://dx.doi.org/10.1016/j.envpol.2016.08.028 0269-7491/© 2016 Elsevier Ltd. All rights reserved.
shown that SE biodegradation by heterotrophic bacteria is very slow. SEs have been reported to have a biodegradation half-life of 20e40 days under aerobic conditions (Clouzot et al., 2008), and longer could be expected under anaerobic conditions (Ying et al., 2003). According to previous studies in our laboratory, even when SEs are continuously discharged into surface water from STPs, their accumulation is not very obvious (Huang et al., 2013). Therefore, it is speculated that some active substances in natural aquatic environments enhance the biodegradation of SEs. Many studies have shown that the fast degradation of pollutants in natural waters can usually be attributed to the accelerating effects of dissolved organic matter (DOM). Quinone groups in DOM can promote microbes to degrade heavy metals (Reijonen et al., 2016) and organic pollutants (Guha et al., 2001). For instance, such groups have recently been reported to play an active role in the de-colorization of azo dyes (Meng et al., 2014) and in the reduction of Cr(VI) (Guo et al., 2012; Huang et al., 2016) under anaerobic conditions. But those studies have been restricted to considering organic acids as electron donors. There are no published reports regarding DOM as a terminal electron acceptor acting
L. Gu et al. / Environmental Pollution 218 (2016) 26e33
27
to accelerate SEs’ biodegradation by quinone-reducing bacteria in aquatic environments. That was the phenomenon explored in this study. Quinone-reducing bacteria have attracted much interest for their respiration diversity, their wide distribution in diverse environments and their potential application in pollution bioremediation through biodegradation. Recently, extensive studies have assessed the catalytic effect of redox mediators on the reduction of heavy metals and organic contaminants by quinone-reducing bacteria isolated from terrestrial or freshwater sources (Cao et al., 2013; Pearce et al., 2006). Recent studies in our own laboratory have demonstrated for the first time their ability to degrade E2 under alkaline conditions. Both DOM and DOM models were considered as terminal electron acceptors in that work. The testing of DOM derived from natural water was a new attempt to mediate E2 biodegradation, and no such results have been published so far. The aim of this study was to test the mediating mechanism of four models of DOM and two actual samples of DOM on the microbial degradation of E2. The influences of temperature, pH, dissolved oxygen and light illumination on the reduction efficiency of the DOM models were assessed using sodium formate as an electron donor. Also investigated were the ability of quinone-reducing bacteria to accelerate DOM reduction and E2 biodegradation, the coupling of E2 as the electron donor and DOM as a terminal electron acceptor to support microbial growth, as well as the effects of DOM concentration on the anaerobic microbial oxidation of E2.
5 g L1 of yeast extract and 10 g L1 of NaCl at 30 C on a rotary shaker and agitated at 160 rpm for 24 h. Subsequently, the strains were separated from the medium by centrifugation at 10,000 rpm for 10 min. The harvested biomass was washed three times with buffer solution before being employed to the experiments. Standard anaerobic culturing techniques were used throughout the study. The base medium was modified from that recommended by Lovley and Phillips, 1988. It contained (in mg L1): NaCl, 1000; NH4Cl, 800; KH2PO4, 500; K2HPO4, 600; MgCl2, 200; CaCl2$2H2O, 50; and yeast extract, 10.
2. Materials and methods
2.4. E2 degradation with model and natural DOM
2.1. Reagents and chemicals
In order to explore the mediating mechanisms of the four DOM models (AQS, AQC, LQ and JQ) and the natural DOM (LHA, LFA and HA) in the microbial degradation of E2, batch experiments were performed assessing: (1) AQS reduction coupling with the oxidation of E2 to support microbial growth; (2) Any effects of the DOM models and their concentrations on the microbial degradation of E2; (3) DOM-mediated microbial degradation of E2; (4) The reduction of DOM by quinone-reducing bacteria with E2 as the electron donor; and (5) The effects of DOM concentration on E2 biodegradation. The experiments were conducted in rubber-sealed 100 mL serum bottles. Prior to use, all of the materials, including the serum bottles, the sealing rubber and the solutions were sterilized in an autoclave at 121 C for 25 min. The DOM model experiments were performed in the dark with 100 mL of MSM that contained different concentrations of AQS, AQC, LQ and JQ (from 0 to 2 mmol L1), different concentrations of E2 (0.5, 1, 1.5 and 2 mg L1), and 0.1 g L1 of quinone-reducing bacteria under anaerobic conditions at pH 8.0 and 30 C for 120 h. The natural DOM experiments were performed with different concentrations of LHA, LFA and HA (from 0 to 8 mgC L1), 1 mg L1 of E2, and 0.1 g L1 of quinone-reducing bacteria under the same conditions.
Anthraquinone-2-sodium sulfonate (AQS), anthraquinone-2carboxylate (AQC), 5-hydroxy-1,4-naphthoquinone (juglone, JQ), 2-hydroxy-1,4-naphthoquinone (lawsone, LQ), E2 and Aldrich humic acid (HA) were supplied by Sigma-Aldrich. The E2 used was high performance liquid chromatograph grade. Its key physicochemical properties are shown in Table 1 (Lai et al., 2000). The natural DOM was extracted from sediment collected from ErHai Lake in Yunnan Province of China. It was sieved using a 4.0 mm sieve and rinsed with deionized water. Lake humic acid (LHA) and lake fulvic acid (LFA) were extracted using the alkali-acid method of the International Humic Substances Society. The HA, LHA and LFA stock solutions were stored in polyethylene containers at 4 C in the dark and used within 3 weeks. All of the other reagents are analytical grade if not otherwise mentioned. 2.2. Bacteria and culturing conditions The quinone-reducing bacteria were also enriched and isolated from the ErHai Lake sediment. The enriched mineral salts medium (MSM) was supplemented with AQS (1 mM) as an electron acceptor and sodium formate (5 mM) as a substrate. To isolate quinonereducing bacteria from the enriched MSM, the enrich MSM was diluted serially and incubated on agar plates containing AQS (1 mM) and sodium formate (5 mM). Selected well-developed colonies were streaked three times with new agar and then preserved for further study. The quinone-reducing bacteria were inoculated into a medium, which contained 10 g L1 of tryptones,
2.3. Reducing characteristics of the AQS Serum bottles containing 100 mL MSM, 0.5 mM AQS, 5 mM sodium formate and 0.1 g L1 of quinone-reducing bacteria were agitated on a rotary shaker at 160 rpm for 120 h to determine the conditions delivering the best AQS reduction efficiency. The effects of temperature, pH, dissolved oxygen and illumination conditions were assessed. The temperatures tested were 0, 5, 10, 15, 25, 30, 35 and 45 C conducted at pH 7.0, and the pH values were 4.0, 5.0, 6.0, 7.0, 8.0, 9.0, 10.0 and 11.0 conducted at 30 C. The dissolved oxygen levels were 0, 0.5, 1.0, 2.0 and 3.0 mg L1 and agitation under natural and ultraviolet light and in the dark were tested. Sterile controls were tested under the same conditions. They were prepared by autoclaving at 121 C for 25 min. Each treatment was set up in triplicate.
2.5. Analytical methods and data analysis AQS absorbance was measured at 336 nm using a UVeVis spectrophotometer, and the reduction efficiency of the AQS was then calculated using the change in AQS absorbance as follows.
The reduction efficiency of AQSð%Þ ¼ ðA0 At Þ=A0 100%
(1)
Table 1 Selected physiochemical properties of the 17b-estradiol. Estrogen
Molecular formula
Molecular weight (g$mol1)
Water solubility at 20 C (mg L1)
logKow
pka
17b-estradiol (E2)
C18H24O2
272.4
3.9e13.3
3.1e4.0
10.5e10.7
28
L. Gu et al. / Environmental Pollution 218 (2016) 26e33
where At is the absorbance of AQS after the biodegradation experiment and A0 is the initial absorbance. The amount of reduced DOM (DOMred) was determined by adding Fe(III) (Lovley et al., 1996). In this process, DOMred was quickly oxidized to DOM by Fe(III), at the same time, Fe(III) was reduced to Fe(II). As Fe(II) was continuously being produced during the interaction of DOMred and Fe(III), the Fe(II) concentration was recorded at selected time points using the 1,10 phenanthroline chemical method (Amonette and Templeton, 1998). The bacterial growth was monitored by measuring the protein concentration with Bradford's method using bovine serum albumin as the standard (Bradford, 1976). The concentration of E2 was quantified using a high performance liquid chromatograph (Agilent Technologies 1260) equipped with a Waters symmetry-C18 reversed-phase column (5 mm, 4.6 250 mm) and a fluorescence detector. The E2 was eluted by the isocratic mixture mobile phase at a rate of 1 mL min1 and detected at the excitation and emission wavelengths of 283 and 345 nm. The mobile phase consisted of acetonitrile and ultrapure water in a ratio of 60:40 containing 0.1% trifluoroacetic acid. The detection limit for E2 was determined to be 0.01 mg L1, and the relative standard deviation for all samples was within 5%. All of the experiments were conducted in triplicate. Statistical analysis was performed using version 20.0 of SPSS for Windows. All of the data were expressed as the mean ± the standard error of the mean. 3. Results and discussion 3.1. AQS reduction Temperature strongly influences microbial growth and thus the reduction efficiency of DOM. If the temperature exceeds the
adaptation range of the microbial enzymes involved, it can induce changes in the charge of the microorganisms' plasma membranes (Stoffels et al., 2008). In these experiments the reduction efficiency of AQS increased with a rise in temperature between 5 and 30 C (Fig. 1-A). Higher temperatures then reduced the reduction efficiency. Fig. 1-A shows that the reduction efficiency of AQS after 120 h was not much different at 25 or 30 C, though the reduction efficiency at 30 C was better than at 25 C between 5 and 60 h. So the optimum temperature was 30 C. Quinone-reducing bacteria at that temperature are at their most effective. The hydrogen ion concentration in MSM greatly influences the bacterial growth, since pH limits enzyme activity (Bhalla et al., 2013; Fenner and Freeman, 2011). In order to determine the optimal pH at which quinone-reducing bacteria reduce AQS, different pH conditions (4.0, 5.0, 6.0, 7.0, 8.0, 9.0, 10.0 and 11.0) were investigated. The relationship between AQS reduction efficiency and pH is shown in Fig. 1-B. The rate of AQS reduction increased quickly when the pH of the MSM was increased from 5.0 to 8.0. A high AQS reduction rate was observed at pH 7.0 or 8.0, but it decreased again at pH values of 9.0 or more. Quinone-reducing bacteria prefer alkaline and neutral environments. Within the pH range of 7.0e9.0 the bacteria exhibited better efficiency in AQS reduction than in more acidic conditions. One reason might be that ErHai Lake is an alkaline lake. However, at pH 10.0 or pH < 5.0, there was no detectable AQS reduction (AH2QS) in the MSM. The quinone-reducing bacteria had apparently been irreversible poisoned by Hþ or OH in overly acidic or alkaline conditions. Previous research has shown that dissolved oxygen is very important to microbial metabolism of organic contaminants in an aquatic environment (Larcher and Yargeau, 2013; Tran et al., 2013). The reduction efficiency of AQS at various dissolved oxygen concentrations is shown in Fig. 1-C. After 10 h there was no significant difference in AQS reduction efficiency among the various dissolved
Fig. 1. Reduction characteristics of AQS (0.5 mM) under different conditions. (A): Temperature (conducted at pH 7.0); (B): pH value (conducted at 30 C); (C): oxygen availability; (D): light conditions. Error bars represent standard deviation of the mean (n ¼ 3).
L. Gu et al. / Environmental Pollution 218 (2016) 26e33
29
oxygen conditions tested (0.5e3.0 mg L1). But in the range of 20e80 h it was dramatically enhanced at lower dissolved oxygen concentrations, and anaerobic was the best condition. Subsequently, after 120 h, the effect of dissolved oxygen on AQS reduction efficiency was unremarkable. One likely explanation is that the AH2QS produced (through reduction of AQS) was oxidized by dissolved oxygen in the MSM. Accumulated AH2QS began to appear in the reaction system only when dissolved oxygen was near to being exhausted (Li et al., 2012; Semblante et al., 2015). Therefore, quinone-reducing bacteria are more effective in anaerobic conditions. The influence of light on AQS reduction efficiency is shown in Fig. 1-D. The reduction under ultraviolet light was only 6.51% after 5 h and there were almost no subsequent changes, probably because the quinone-reducing bacteria had been killed by the ultraviolet light after 5 h. The AQS reduction in the dark was more complete than under natural light. But the efficiency was almost the same in the dark and under natural light between 0 and 10 h. The data show, however, that the reduction efficiency in the dark gradually gained an advantage from 10 to 80 h. Two major reasons suggest themselves. Quinone-reducing bacteria isolated from ErHai Lake sediment have never experienced natural light exposure. And at the same time, light is likely to affect microbial enzyme metabolism and inhibit the growth of many microorganisms (Sivonen, 1990). The quinone-reducing bacteria tested seem to have more applicability in dark condition.
3.2. Microbial oxidation of E2 with DOM models 3.2.1. AQS reduction with E2 as the electron donor Previous research has shown that electron transfer is influenced by the quinone structure in model DOM, such as AQS, AQC and anthraquinone-2-6-sulfonate (AQDS) (Wu et al., 2014). Recent studies have demonstrated that the mechanism of quinone structure involved pollutant transformation was along with the cycles between the oxidation of carbonyl structure and the reduction of hydroxyl quinone (Lovley et al., 1996). Applying the optimum conditions previously determined, all of the following experiments were carried out in the dark under anaerobic conditions at pH 8.0 and 30 C. The reduction of model DOM (take AQS as examples) by quinone-reducing bacteria was investigated with E2 as the electron donor. Fig. 2-A shows that almost no AH2QS could be detected when only AQS and E2 were added to the MSM without adding quinone-reducing bacteria. The addition of quinone-reducing bacteria after 20 h produced a small amount of AH2QS even without E2 as an electron donor. This is probably because the medium contained trace amounts of yeast extract, which could enhance quinone-reducing bacteria growth from 0 to 10 h (Fig. 2-B), resulting in some AQS reduction. However, experimental results show that adding E2 as electron donor significantly improved the reduction efficiency. It eventually reached about 46% after 120 h (Fig. 2-A). Fig. 2-B indicates that after 120 h the microbial growth with E2 was 3.01 times as much as without it. This shows that quinone-reducing bacteria can use E2 as an electron donor to enhance the reducing efficiency of AQS and their own growth.
3.2.2. Effects of DOM models on E2 biodegradation A number of kinetic models have been applied to describe the biodegradation of organic pollutants, with pseudo-first-order kinetics frequently used as a convenient way to describe the biodegradation progress of xenobiotic compounds (Nyholm et al., 1996). The biodegradation of E2 was assumed at the outset to follow pseudo-first-order kinetics.
Fig. 2. AQS reduction coupling with the oxidation of E2 to support microbial growth. (A): The reduction efficiency of AQS; (B): microbial growth (OD value). The experiments were performed in the dark under anaerobic conditions at pH 8.0 and 30 C for 120 h. The initial concentration of E2 was 1 mg L1. The error bars represent the standard deviation of the mean (n ¼ 3).
ln C=C0 ¼ kt þ b
(2)
here C is the concentration of E2 after a period of biodegradation; C0 is the initial concentration of E2 in the MSM; k (h1) is the degradation rate constant; t (h) is the degradation time; and b is a fitting parameter. When b ¼ 0, Eq. (2) can be expressed as a first-order kinetic equation.
ln C=C0 ¼ k0 $t
(3)
The half-life (T1/2) of E2 is then:
T1=2 ¼ ln 2=k0
(4)
Although some studies have shown that the presence of AQDS can improve the reduction performance of Shewanella with organic contaminants (Guha et al., 2001) and Cr(VI) (Reijonen et al., 2016; Huang et al., 2016), no information has previously been published about the effects of quinone compounds on E2 biodegradation. Fig. 3 demonstrates that AQS best stimulated the biodegradation of E2 at AQS concentrations from 0 to 0.5 mM and that the mediation effect improved with increasing AQS concentration over that range.
30
L. Gu et al. / Environmental Pollution 218 (2016) 26e33
Fig. 3. Effect of DOM models and their concentrations on the microbial degradation of E2. The experiments were performed in the dark under anaerobic conditions at pH 8.0 and 30 C for 120 h. The initial concentration of E2 was 1.0 mg L1. The error bars represent the standard deviation of the mean (n ¼ 3).
Adding 0.5 mM AQS increased the k value by almost 50% (from 0.00197 h1 to 0.00295 h1). Adding 0.2 mM AQC resulted in a 34% increase. However, greater AQS and AQC concentrations demonstrated no further enhancement of E2 biodegradation. Improved E2 biodegradation was also observed with the addition of LQ and JQ at concentrations up to 0.1 mM. When the concentration of these DOM models was greater, inhibited E2 biodegradation was observed. The inhibition might be attributed to their toxicity to bacterial cells (Lown, 1983). The constants of the relevant kinetic equations are presented in Table 2. The results show that increasing E2 concentration would give rise to slower biodegradation, whether or not the optimum concentration of the DOM model is present. With the E2 concentration increased from 0.5 to 2 mg L1 in MSM without a model DOM, the half-life of E2 is enhanced from approximately 290 to 436 h. However, adding the optimum concentration of a DOM model can reduce the half-life significantly. With 0.5 mM AQS it becomes 218e288 h. With 0.2 mM AQC it was 239e318 h 0.1 mM
Fig. 4. DOM-mediated microbial degradation of E2. The experiments were performed in the dark under anaerobic conditions at pH 8.0 and 30 C for 120 h. The concentration of LHA, LFA and HA was 3.0 mgC L1. The initial concentration of E2 was 1.0 mg L1. Error bars represent standard deviation of the mean (n ¼ 3).
LQ reduced it to 270e358 h. And 0.1 mM JQ yielded a half-life of 274e371 h (Table 2). These results further illustrate the extent to which quinones can improve the biodegradation of E2. 3.3. DOM-mediated microbial degradation of E2 in an alkaline environment Previous studies have reported that DOM models can act electron shuttle mediators to remove toxic heavy metals (Wang et al., 2011) and organic contaminants (Meng et al., 2014). However, no report of actual DOM mediating biodegradation of E2 has been published so far. The results presented in Fig. 4 indicate that LHA, LFA or HA alone existed in the MSM had a positive effect on the biodegradation of E2 in a sterile environment at the beginning. The main reason is perhaps that DOM has plenty of hydrophobic functional groups that could interact with steroid endocrine disruptors (Zhu et al., 2012). After 120 h, the microbial degradation of
Table 2 Rate of E2 biodegradation by quinone-reducing bacteria with DOM models. DOM models
Concentration of DOM model (mmol L1)
Initial concentration of E2 (mg L1)
Rate constant (h1)
T1/2 (h)
R2
None
0
0.5 1.0 1.5 2.0
0.00239 0.00213 0.00186 0.00159
290.02 325.42 372.66 435.94
0.9281 0.9332 0.9447 0.9406
AQS
0.5
0.5 1.0 1.5 2.0
0.00318 0.00292 0.00267 0.00241
217.97 237.38 259.61 287.61
0.9452 0.9687 0.9512 0.9496
AQC
0.2
0.5 1.0 1.5 2.0
0.00290 0.00264 0.00241 0.00218
239.02 262.56 287.61 317.96
0.9367 0.9442 0.9531 0.9621
LQ
0.1
0.5 1.0 1.5 2.0
0.00257 0.00236 0.00215 0.00194
269.71 293.71 322.39 357.29
0.9335 0.9531 0.9504 0.9448
JQ
0.1
0.5 1.0 1.5 2.0
0.00253 0.00231 0.00209 0.00187
273.97 300.06 331.65 370.67
0.9451 0.9578 0.9483 0.9517
L. Gu et al. / Environmental Pollution 218 (2016) 26e33
E2 was about 23% without DOM, but with the addition of LHA it was 38%, with LFA 43% and with HA 30% after 120 h. So LHA and HFA can more effectively promote microbial degradation of E2. These results suggest that the quinone structure of DOM plays a very important
31
role in promoting the degradation of E2 by quinone-reducing bacteria, and that the mechanism is likely to be consistent with that of DOM models. Meanwhile, coupling E2 biodegradation with DOM reduction can accelerate microbial growth. The data of Fig. 4 suggest that the mediating ability of LHA, LFA and HA is in the order LFA > LHA > HA. The amount of reduced DOM (DOMred) was determined by adding Fe(III) and determining the extent of Fe(II) production (Lovley et al., 1996). Fig. 5 demonstrates that DOM can directly restore Fe(III) without the presence of microorganisms, probably because DOM is a mixture of macromolecules (Schmidt et al., 2011) and possesses trace amounts of reducing substances (He et al., 2014). In addition, DOM is very complex and it contains a small amount of beneficial nutrients (Thurman, 1985) which can accelerate the growth of microorganisms. For example, a little of the protein co-substrate in MSM could be utilized by quinone-reducing bacteria to promote the reduction of Fe(III) (Hur et al., 2011). Unfortunately, however, the protein content of DOM is in general much too low to support the growth of quinone-reducing bacteria. But Fig. 5 shows that adding a certain amount of E2 as an electron donor can effectively promote the reduction of DOM. After 120 h, the DOM was significantly reduced in the active treatments (E2 þ DOM þ quinone-reducing bacteria), and the LHAred, LFAred or HAred concentration reached 0.172 mM, 0.211 mM or 0.091 mM, respectively. These results indicate that quinone groups in actual DOM extracts can be terminal electron acceptors coupling with microbial growth to promote the biodegradation of E2. 3.4. DOM concentration The results presented in Fig. 6 and Table 3 indicate that LFA best promoted E2 biodegradation, and that its mediating effect was enhanced as the LFA concentration increased from 0 to 2 mgC L1. Compared to that obtained in the absence of DOM (0.00197 h1), the k value increased by 78% (to 0.00351 h1) when 2 mgC L1 of LFA was present. The LFA was reduced to 0.234 ± 0.001 mM along with the OD growing to 1.025 ± 0.012. Adding 3 mgC L1 of LHA gave a 55% increase and 5 mgC L1 of HA produced a 22% increase in the k value. LHA and HA were reduced to 0.172 ± 0.003 mM and 0.104 ± 0.002 mM respectively, coupled with optical densities as high as 0.781 ± 0.010 for LHA and 0.683 ± 0.010 for HA (Fig. 6 and
Fig. 5. The reduction of DOM by quinone-reducing bacteria with E2 as the electron donor. (A) LHA (3.0 mgC L1); (B) LFA (3.0 mgC L1); (C) HA (3.0 mgC L1). The experiments were performed in the dark under anaerobic conditions at pH 8.0 and 30 C for 120 h. The initial concentration of E2 was 1.0 mg L1. The error bars represent the standard deviation of the mean (n ¼ 3).
Fig. 6. Effects of DOM concentration on E2 biodegradation. The experiments were performed in the dark under anaerobic conditions at pH 8.0 and 30 C for 120 h. The initial concentration of E2 was 1.0 mg L1. The error bars represent the standard deviation of the mean (n ¼ 3).
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L. Gu et al. / Environmental Pollution 218 (2016) 26e33
Table 3 Effects of DOM concentration on DOM reduction and microorganism growth. Initial concentration of DOM (mgC L1)
HA LHAred (mmol L1)
OD
0 0.2 0.5 1 2 3 4 5 8
0 0.023 0.065 0.116 0.144 0.172 0.157 0.124 0.048
0.447 0.449 0.496 0.584 0.703 0.781 0.752 0.646 0.573
FA
± ± ± ± ± ± ± ±
0.001 0.001 0.002 0.001 0.003 0.003 0.001 0.002
± ± ± ± ± ± ± ± ±
0.011 0.012 0.014 0.012 0.013 0.010 0.009 0.016 0.012
PHA
LFAred (mmol L1)
OD
0 0.056 0.137 0.178 0.234 0.211 0.161 0.142 0.097
0.447 0.549 0.737 0.921 1.025 0.937 0.821 0.756 0.663
± ± ± ± ± ± ± ±
0.003 0.002 0.002 0.001 0.005 0.001 0.003 0.003
± ± ± ± ± ± ± ± ±
0.011 0.014 0.011 0.016 0.012 0.022 0.023 0.019 0.014
HAred (mmol L1)
OD
0 0.022 0.046 0.073 0.081 0.091 0.096 0.104 0.066
0.447 0.457 0.478 0.526 0.583 0.594 0.602 0.683 0.501
± ± ± ± ± ± ± ±
0.001 0.001 0.001 0.004 0.002 0.001 0.002 0.003
± ± ± ± ± ± ± ± ±
0.011 0.012 0.018 0.013 0.011 0.014 0.018 0.010 0.014
Note: ± error represents the standard deviation of the mean of 3 replicates.
Table 3). However, the continuously increasing DOM concentration would restrain the growth of microorganisms and the biodegradation process. These results are consistent with the findings of the studies with the DOM models. The better stimulating affect observed with LHA and LFA, which tend to be more aromatic, demonstrates their greater electron-accepting capacity (O'Loughlin, 2008) and generally more effective support for the growth of microorganisms than HA. 4. Conclusion Quinone-reducing bacteria from ErHai Lake sediment gave their best reduction performance with AQS. The optimal pH value was 8.0 and the optimal temperature was at 30 C. Additionally, those quinone-reducing bacteria work best under anaerobic condition and in the dark. Quinone structures in DOM apparently function as terminal electron acceptors and couple with microbial growth to promote the biodegradation of E2. The optimum DOM concentration can reduce the half-life of E2 substantially, however too high DOM concentration can restrain microorganism growth and E2 biodegradation. The more aromatic components of DOM have greater electron-accepting capacity and are generally more effective support for the growth of microorganisms. These results go some way to explaining the observed biodegradation of steroid estrogens in alkaline environments containing DOM. Acknowledgments This project was sponsored by the National Natural Science Foundation of China (grant no. 41401558) and the China Postdoctoral Science Foundation (grant no. 2014T70887). References Amonette, J.E., Templeton, J.C., 1998. Improvements to the quantitative assay of nonrefractory minerals for Fe (II) and total Fe using 1, 10-phenanthroline. Clay. Clay Miner 46 (1), 51e62. Bhalla, A., Bansal, N., Kumar, S., Bischoff, K.M., Sani, R.K., 2013. Improved lignocellulose conversion to biofuels with thermophilic bacteria and thermostable enzymes. Bioresour. Technol. 128, 751e759. Bradford, M.M., 1976. A rapid and sensitive method for the quantitation of microgram quantities of protein utilizing the principle of protein-dye binding. Anal. Biochem. 72, 248e254. Cao, D.M., Xiao, X., Wu, Y.M., Ma, X.B., Wang, M.N., Wu, Y.Y., Du, D.L., 2013. Role of electricity production in the anaerobic decolorization of dye mixture by exoelectrogenic bacterium Shewanella oneidensis MR-1. Bioresour. Technol. 136, 176e181. Clouzot, L., Marrot, B., Doumenq, P., Roche, N., 2008. 17a-ethinylestradiol: an endocrine disrupter of great concern. Analytical methods and removal processes applied to water purification. a review. Environ. Prog. 27, 383e396. Fenner, N., Freeman, C., 2011. Drought-induced carbon loss in peatlands. Nat. Geosci. 4, 895e900. Guha, H., Jayachandran, K., Maurrasse, F., 2001. Kinetics of chromium (VI) reduction by a type strain Shewanella alga under different growth conditions. Environ.
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