Insight into the short- and long-term effects of Cu(II) on denitrifying biogranules

Journal of Hazardous Materials 304 (2016) 448–456

Contents lists available at ScienceDirect

Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat

Insight into the short- and long-term effects of Cu(II) on denitrifying biogranules Hui Chen a,b , Qian-Qian Chen a,b , Xiao-Yan Jiang a,b , Hai-Yan Hu a , Man-Ling Shi a , Ren-Cun Jin a,b,∗ a b

College of Life and Environmental Sciences, Hangzhou Normal University, Hangzhou 310036, China Key Laboratory of Hangzhou City for Ecosystem Protection and Restoration, Hangzhou Normal University, Hangzhou 310036, China

h i g h l i g h t s

g r a p h i c a l

a b s t r a c t

• It is the first time to evaluate the effect of Cu2+ on denitrifying biogranules. • A high level of Cu(II) was investigated during batch assays and continuous tests. • Mechanisms of the effects of Cu2+ on denitrifying biogranules were discussed. • Effects of pre-exposure to Cu2+ and starvation treatments were investigated.

a r t i c l e

i n f o

Article history: Received 3 August 2015 Received in revised form 16 October 2015 Accepted 8 November 2015 Available online 14 November 2015 Keywords: Cu2+ toxicity Denitrifying biogranules Pre-exposure Kinetic models Acclimation

∗ Corresponding author. Fax: +86 571 28865333. E-mail address: [email protected] (R.-C. Jin). http://dx.doi.org/10.1016/j.jhazmat.2015.11.012 0304-3894/© 2015 Elsevier B.V. All rights reserved.

a b s t r a c t This study aimed to investigate the short- and long-term effects of Cu2+ on the activity and performance of denitrifying bacteria. The short-term effects of various concentrations of Cu2+ on the denitrifying bacteria were evaluated using batch assays. The specific denitrifying activity (SDA) decreased from 14.3 ± 2.2 (without Cu2+ ) to 6.1 ± 0.1 mg N h−1 g−1 VSS (100 mgCu2+ L−1 ) when Cu2+ increased from 0 to 100 mg L−1 with an increment of 10 mgCu2+ L−1 . A non-competitive inhibition model was used to calculate the 50% inhibition concentration (IC50 ) of Cu2+ on denitrifying sludge (30.6 ± 2.5 mg L−1 ). Monod and Luong models were applied to investigate the influence of the initial substrate concentration, and the results suggested that the maximum substrate removal rate would be reduced with Cu2+ supplementation. Preexposure to Cu2+ could lead to an 18.2–46.2% decrease in the SDA and decreasing percentage of the SDA increased with both exposure time and concentration. In the continuous-flow test, Cu2+ concentration varied from 1 to 75 mg L−1 ; however, no clear deterioration was observed in the reactor, and the reactor was kept stable, with the total nitrogen removal efficiency and total organic carbon efficiency greater than 89.0 and 85.0%, respectively. The results demonstrated the short-term inhibition of Cu2+ upon denitrification, and no notable adversity was observed during the continuous-flow test after long-term acclimation. © 2015 Elsevier B.V. All rights reserved.

H. Chen et al. / Journal of Hazardous Materials 304 (2016) 448–456

1. Introduction With the high demand for nitrogen removal from wastewater, biological nitrogen removal processes have been widely applied in wastewater treatment procedures worldwide [1–5]. Stable denitrification must be maintained to achieve optimal biological nitrogen removal efficiency because of the important role that denitrification plays in the biological nitrogen removal process. Anaerobic granules are self-immobilized aggregates and have been reported to possess many advantages. Anaerobic sludge has the capability to form self-immobilized granules, provided that the physical and chemical conditions for sludge flocculation are favorable. Previous studies also demonstrated that denitrifying bacteria can create well-sedimented granules [6,7]. Granules have more favorable sludge characteristics than the flocculent sludge, e.g., better settleability, denser structure and greater tolerance to shock loadings, favoring biomass retention and maintaining an acceptable reactor performance [8]. Microorganisms require trace elements, such as heavy metals, to maintain various metabolic activities [9,10]. The effect of heavy metals on microbial growth and activity is concentration dependent [11]. Low concentrations of heavy metals may be beneficial for microbial activities because some of the metals are components of many enzymes or co-enzymes, whereas excessive concentrations can inhibit microbial activities due to their accumulative bio-toxicity [11–13]. Copper is a heavy metal commonly found in wastewater, and some studies have demonstrated its effect on the performance of biological removal processes at concentrations of 0–100 mg L−1 [5,11,13]. Although many studies have tested the effect of copper on nitrification, anaerobic ammonium oxidation (anammox), biological phosphorus removal and other biological processes [11,14–18], few studies have focused on denitrification. The addition of 30 mg L−1 of Cu(II) was reported to lead to a 50% decrease in the bioactivity of granules for fermentative hydrogen production [19]. Ting et al. [18] found that biological nitrogen and phosphorus removal capacity would significantly decrease upon the addition of 1 mg L−1 copper. Li et al. [17] determined that the 50% inhibition concentration of copper for the anammox biomass was 4.2 mg L−1 . Yang et al. [11] revealed the short- and long-term effects of copper on the anammox process, and the 50% inhibition concentration was calculated to be 12.9 mg L−1 by batch assays, and 94% of the microbial activity was lost when 5 mg L−1 copper was added during the continuous-flow test. The effects of Cu2+ inhibition on the denitrifying biogranules were based on batch assays that utilized biomass from a nursing reactor. Little is known about the sensitivity of denitrifying biogranules toward Cu2+ after being starved. However, denitrifying sludge can experience starvation in under-loaded bioreactors or during preservation of the sludge. In this study, the denitrifying biomass was fed with synthetic wastewater containing glucose as the electron donor. The short-term effect of Cu2+ on the denitrifying biogranules was examined by batch assays, the influence of the initial substrate concentration was investigated, and the effects of starvation and pre-exposure were evaluated. Furthermore, the long-term effect of Cu2+ on the denitrifying reactor was observed.

2. Methods and materials 2.1. Experimental setup and inoculums

in a thermostatic room at 30 ± 1 ◦ C and operated with a constant hydraulic retention time (HRT) of 2.4 h. Denitrifying seed granules were harvested from a stably operated laboratory-scale denitrifying UASB reactor. These mature denitrifying granules were kept under favorable conditions, favoring reactor performance with an average settling velocity of 93.2 ± 4.9 m h−1 and a particle diameter of 1.69 mm. 2.2. Synthetic wastewater The feeding materials were synthetic, and glucose and sodium nitrate were chosen as the C and N sources, respectively. The chemical oxygen demand (COD) to nitrate-nitrogen (COD/NO3 − -N) ratio of 6.0 was higher than the theoretical stoichiometric ratio (4.9) for complete denitrification (including bacterial growth) [20], allowing nitrate to be the limiting substrate. The composition of the synthetic wastewater is presented in Table 1. The pH of the synthetic wastewater was maintained in the range of 6.9–7.3 and was adjusted by the addition of 1 mol L−1 sodium hydroxide or hydrochloric acid solutions. 2.3. Analytical methods Influent and effluent samples were taken and analyzed every 2 days. The SS, VSS, NH4 + -N, NO2 − -N and NO3 − -N values were determined according to standard methods [21]. The pH was measured using a pH meter (Delta320, Mettle-Toledo, Switzerland). The Cu2+ content in the sludge was determined according to the method described by Chen et al. [6]. The total organic carbon (TOC) content was measured using a TOC analyzer (TOC-L-CPN, Shimadzu, Japan). 2.4. Specific denitrifying activity (SDA) assays SDA batch assays were performed in serum bottles with a liquid volume of 120 mL; the set of assays is listed in Table 2. The wastewater was prepared as described above, the pH was adjusted as needed (6.9–7.3), and the COD/NO3 − -N ratio was fixed at 6.0. The bottles were flushed with pure argon gas (99.99%) for 15 min and firmly closed with a 4-mm-thick butyl rubber septa to maintain an anaerobic environment. Subsequently, the bottles were placed in a vibrator at 35 ± 1 ◦ C and on a shaker at 180 rpm. The NO3 − -N and COD concentrations were monitored every 30 min, and the SDA was estimated according to a procedure reported by Chen et al. [6]. 2.5. Pre-exposure assays The set of pre-exposure assays is presented in Table 2. The assays were divided into two parts to observe the influence of pre-exposure on the SDA, i.e., the influence of the exposure concentration of Cu2+ at a constant exposure duration and the pre-exposure duration at a constant concentration of Cu2+ . The influence of the substrate during exposure was also investigated. Table 1 Composition of the synthetic wastewater. Composition

Concentration (mg L−1 )

Composition

Concentration (mg L−1 )

Na2 HPO4 NaH2 PO4 NaHCO3 CaCl2

2655 2375 400 0.5

Trace element Ia Trace element IIb Nitrate Glucose

1.25* 1.25* 700 4200

The ingredient of the trace element I (g L−1 ): 5.000 EDTA, 9.14 FeSO4 ·7H2 O. The ingredient of the trace element II (g L−1 ): 15.000 EDTA, 0.430 ZnSO4 ·7H2 O, 0.240CoCl2 ·6H2 O, 0.990 MnCl2 ·4H2 O, 0.250CuSO4 ·5H2 O, 0.220 NaMoO4 ·2H2 O, 0.210 NiCl2 ·6H2 O, 0.014H3 BO4 . a

The experiment was conducted in an upflow anaerobic sludge blanket (UASB) with a working volume of 1 L. The internal diameter of the reactor was 60 mm. The denitrifying reactor was placed

449

b

450

H. Chen et al. / Journal of Hazardous Materials 304 (2016) 448–456

Table 2 Experimental conditions of the short-term tests. Substrate concentration (mg L−1 )

Contents

−

Cu2+ concentration (mg L−1 )

Exposure time (min)

–

720

NO3 -N

COD

Effect of initial substrate

100,200,400,600,800,1000

600,1200,2400,3600,4800,6000

Effect of Cu2+ Role of exposure time

400 400 – 400 –

2400 2400

20 – 0,2,5,7,10,15,20,25,50,75,100 10

2400

5,10,15

Role of exposure concentration

– 30,60,90,120

Hence, assays under each condition were carried out with or without the substrate. The biomass was washed with phosphate buffer solution three times after pre-exposure, and then, the SDA was measured. 2.6. Starvation assays A set of experiments were designed to investigate how starvation could affect the SDA, and 24, 48 and 72 h were chosen as the starvation periods. The starvation assay was processed in the sealed serum bottles and the biomass was washed with phosphate buffer solution three times after starvation, and then, the SDA was measured according to the method described in Section 2.4. 2.7. Kinetic models The modified non-competitive inhibition model (Eq. (1)) was applied to determine the inhibitory characteristics of the Cu2+ concentration on the SDA.



I (%) = 100 ×

1−





1

b

(1)

1 + [Cu] /a

where I (%) is the inhibition response, [Cu] is the concentration of Cu2+ , a is the 50% inhibitory Cu2+ concentration (mg L−1 ) and b is the fitting parameter. The Monod equation is generally accepted as the basis of the activated sludge models used for modeling the majority of biological treatment processes in the wastewater treatment field. q=

qmax S S + KS

(2)

where q is the substrate removal rate (mgN h−1 g−1 VSS), qmax is the maximum substrate removal rate (mgN h−1 g−1 VSS), S is the substrate concentration (mg L−1 ) and KS is the half saturation constant (mg L−1 ). The Luong model [22] can be used to describe biochemical reactions with low to high substrate concentrations, and the model is described below.



r=

rmax S 1 − S/Sm

n

Ks + S

Fig. 1. Short-term effects of Cu2+ on the denitrifying granules. A, relationship between the Cu2+ concentration and SDA; B, plot of the non-competitive inhibition model.

(3)

where r is the substrate removal rate (mgN h−1 g−1 VSS), rmax is the maximum substrate removal rate (mgN h−1 g−1 VSS), S is the concentration of the substrate (mg L−1 ), Sm is the critical inhibitory concentration above which the reaction stops (mg L−1 ), n is an empirical constant and KS is the half saturation coefficient (mg L−1 ). 2.8. Statistical analysis Statistical comparison between variables was conducted using one-way analysis of variance (ANOVA) by SPSS software (SPSS 13.0). A p-value of 0.05 or lower indicates that the difference between the variables under comparison is statistically significant.

3. Results and discussion 3.1. Short-term effects of Cu2+ on the denitrifying biogranules 3.1.1. Effects of the Cu2+ concentration The short-term responses of the denitrifying biomass to Cu2+ stress with a constant initial substrate level were observed (Table 2). Various concentrations of Cu2+ were prepared to confirm the short-term response of the SDA to the Cu2+ dosage. The toxicity of Cu2+ to the denitrifying bacteria was positively related with its dosage, i.e., the SDA decreased markedly with an increasing Cu2+ dosage (Fig. 1A).

H. Chen et al. / Journal of Hazardous Materials 304 (2016) 448–456

451

The modified non-competitive inhibition model was used to describe the characteristics of Cu2+ inhibition on denitrification (Fig. 1B). The regression equation is shown below (Eq. (4)). From the regression, the 50% inhibition concentration of Cu2+ (IC50 ) was calculated to be 30.6 mg L−1 .



I (%) = 100 ×

1−





1

−0.5

1 + [Cu] /30.6

(R2 = 0.8819)

(4)

where I (%) is the inhibition response and [Cu] is the Cu2+ concentration (mg L−1 ). The non-competitive inhibition model has been frequently used in the anammox system to find the relation between the inhibitor concentration and specific anammox activity (SAA). Yang et al. [11] tested the short-term effect of Cu2+ on the anammox biomass, and the IC50 was calculated as 12.9 mg L−1 . However, the noncompetitive inhibition model is seldom used in a denitrifying system. Quantitative studies have been carried out to investigate the effect of Cu2+ on the microorganisms [11,14,17,23,24]. The activity of an acetate-degrading methanogenic enrichment culture was inhibited by 50% at a Cu2+ concentration of 12.5 mg L−1 [23]. Cu2+ caused 50% inhibition of a sulfate-reducing mixed culture at an initial dissolved concentration of 10.5 mg L−1 [24]. Yang et al. [11] reported that the IC50 of Cu2+ toward the anammox biomass was 12.9 mg L−1 . Similar results were achieved, which suggested that the low concentration of Cu2+ would significantly inhibit the SAA. However, few studies have considered the effect of Cu2+ on denitrifiers. Wu et al. [5] found that compared with the control without the addition of copper, SDA decreased by 89% for acetate-acclimated denitrifiers and by only 15% for methanol-acclimated denitrifiers at a copper concentration of 1 mg L−1 , whereas it decreased by 98% for acetate-acclimated denitrifiers and by 46% for methanolacclimated denitrifiers at a copper concentration of 2 mg L−1 . In this study, 95% SDA remained at a Cu2+ dosage of 2 mg L−1 . The great discrepancy may be due to the difference in carbon source (glucose was used in this study) and the initial substrate concentration. Cu2+ is an essential element for the microorganism, although an excess dose of Cu2+ can restrain the metabolism of the cells. Cu2+ and some compounds containing Cu2+ are often used as antimicrobial and antifungal chemicals by chelating with hydrosulfonyl groups to damage the enzymes and proteins [25]. The toxicity of Cu ions toward microorganisms has been well documented [26,27]. Cu ions can interact with intracellular iron-sulfur clusters of proteins and inhibit their function under anoxic conditions [26]. 3.1.2. Effect of the initial substrate concentration The relationship between the SDAs and substrate concentration, both with and without 20 mg L−1 Cu2+ , is shown in Fig. 2. The SDA initially increased and then decreased as the initial substrate concentration increased, both with and without Cu2+ . The substrate levels clearly affect the SAA. A low substrate level is known to result in a low impetus for mass transfer, and thus, the maximum SAA value was difficult to attain, and a high resistance to mass transfer in the granule may have further decreased the SDA. Anammox performance could be suppressed with an exceedingly high initial substrate concentration [11]. The results obtained in this study were consistent with those obtained by Yang et al. [11], i.e., when the initial substrate reached a certain concentration, the SDA started to decline. Without added Cu2+ , the simulation results of the Monod model yielded qmax and KS values of 13.6 mgN h−1 g−1 VSS and 68.9 mg L−1 , respectively (R2 = 0.7636). The addition of Cu2+ led to a smaller qmax of 7.1 mgN h−1 g−1 VSS but a higher KS of 70.5 (R2 = 0.9401). qmax will decline in the presence of noncompetitive inhibitors [28]. The simulation results from the Luong

Fig. 2. Short-term effect of the initial NO3 − concentration on the SDA with/without Cu2+ .

model are presented below (Eq. (5) without Cu2+ and Eq. (6) with Cu2+ ). The equations suggested that rmax , Sm and KS decreased significantly in the presence of 20 mg L−1 Cu2+ (p < 0.05), indicating a lower maximum substrate removal rate (rmax ), i.e., a decrease in the treatment potential of the denitrifying reactor, a decrease in the affinity to the substrate (KS ) and a lower critical inhibitory concentration, above which the reaction stops. Moreover, there were significant differences in the SDAs obtained with or without Cu2+ , which was confirmed by statistical analysis (p < 0.05). In other words, 20 mg L−1 Cu2+ could seriously inhibit the microbial activity. r = 30.7 × r = 7.5 ×

S(1 − S/2113.8) 270.9 + S

S(1 − S/1008.9) 80.9 + S

1.92

0.03







R2 = 0.9147



R2 = 0.8995

(5)

(6)

2.1.3. Role of pre-exposure The response of denitrifying biogranules to pre-exposure to Cu2+ was investigated, and the experimental conditions are presented in Table 2. The effects of both pre-exposure concentration and time were determined, and the results are shown in Fig. 3. In Fig. 3A, the SDA decreased with an increasing Cu2+ concentration, particularly with substrate during pre-exposure, when the pre-exposure time was fixed at 12 h. The normalized denitrifying activity (NDA = SAAinhibited /SAAcontrol × 100%) was below 60% for the biomass with a substrate at when the dosage was over 10 mg L−1 Cu2+ , whereas the corresponding NDA for the biomass without substrate was approximately 80%. In both cases, the NDA tended to decline as the Cu2+ concentration increased during preexposure. Different trends were observed in the pre-exposure time tests. In the test with a substrate during pre-exposure, the SDA decreased as the pre-exposure time increased, whereas the change of the SDA was irregular in the one without a substrate (Fig. 3B). The pre-exposure toxicity was magnified by the addition of substrate during the exposure period. In the study of Carvajal-Arroyo et al. [29], non-metabolic (without substrate) pre-exposure to NO2 − had a rapid detrimental effect, with 74% of the full effect occurring with 30 min. This rapid effect occurs because during active metabolism, the anammox reaction constitutes a continuous sink for NO2 − , preventing its accumulation inside a putative sensitive region of the cells. In this study, however, metabolic pre-exposure (with substrate) to Cu2+ was shown to have a more severe effect than non-metabolic pre-exposure. Cu strongly interacts with organic matter [30,31]. Copper may undergo

452

H. Chen et al. / Journal of Hazardous Materials 304 (2016) 448–456

Fig. 4. Details of the short-term effect of starvation (24, 48 and 72 h) on the SDA.

Fig. 3. Short-term effect of pre-exposure to Cu2+ on the SDA. A, effect of exposure concentration of Cu2+ with exposure time of 12 h with/without substrate; B, exposure time with 10 mgCu2+ L−1 with/without substrate.

biosorption by cell membrane surfaces with proteins and acid groups that serve as binding sites [32–34]. Hence, pre-exposure with a substrate for the biomass increased the chances of accumulating more Cu2+ . Metal inhibition is a physicochemical rather than a biological transport process because sorption plays a dominant role in metal uptake [35]. The rates of sorption and internalization of Cu2+ by nitrifying bacteria are relatively faster than those for Zn2+ , Ni2+ and Cd2+ , and the inhibitory mechanism of Cu2+ is also different and may involve a rapid loss of membrane integrity [35]. Cu2+ can catalyze the production of hydroxyl radicals and press redox cycling activity, resulting in damage to membrane function [35]. In summary, the results of the tests suggested that the SDA would significantly decline after pre-exposure to Cu2+ compared to the control biomass that was not subjected to pre-exposure treatment. Moreover, the existence of a substrate during pre-exposure led to a lower SD compared with the one without a substrate under identical conditions. 3.1.3. Role of starvation A set of experiments was designed to investigate how starvation could affect the SDA, and 24, 48 and 72 h were chosen as the starvation periods. Fig. 4 shows the relationship between SDA/NDA and the starvation period. In Fig. 4, compared with the control, i.e., fresh biomass without starvation, the SDA of the starved biomass declined to 12.6 ± 0.4, 11.8 ± 1.0 and 11.1 ± 0.1 mgN h−1 g−1 VSS after a starvation period of 24, 48 and 72 h, respectively, whereas the SDA for the control was 13.6 ± 1.2 mgN h−1 g−1 VSS. The cor-

responding NDAs were 92.3%, 86.8% and 81.3% after 24-, 48, and 72 h starvation periods, respectively. As previously discussed, Cu2+ had an adverse impact on the denitrifying bacteria and may have worsened the situation when the biomass suffered starvation. The results showed that the NDA achieved was inversely proportional to the duration of starvation suffered by the biomass. Wu et al. [36] confirmed that the quantity of the anammox bacteria would increase over the starvation duration. Similar results were obtained by Carvajal-Arroyo et al. [37] in an anammox system. The IC50 of nitrite in starved- and fresh-resting-cells was 7 and 52 mgN L−1 , respectively. Anammox biomass that suffered from starvation treatment was less resistant to NO2 − inhibition compared with the fresh one. Moreover, a significantly reduced extracellular polymeric substance content, inhibited microbial activity, poor settleability and weakened structural integrity of aerobic granules were observed under the starvation of the substrate and nutrients in an aerobic system [38]. However, in other studies, stimulation was observed after starvation. In a previously reported study, phosphorus removal efficiency was enhanced by a combination of starvation and co-immobilization with microalgae growth-promoting bacteria [39]. Torà et al. [40] announced that a long-term starvation (30 days) was not an insurmountable problem for the implementation of a high-rate partial nitrification activated sludge process for treating industrial high-strength ammonium wastewater. The influence of starvation may depend on various factors, e.g., the type of bacteria, the loading rate of the nursing reactor, and the starvation period. Hence, contradictory results were achieved in this study and previous studies. 3.2. Long-term effects of Cu2+ on the denitrifying reactor 3.2.1. Performance of the denitrifying reactor A continuous-flow test was conducted to investigate the longterm effect of Cu2+ on the denitrifying reactor. The SS and VSS of the reactor were 51.5 and 36.3 g L−1 , respectively, with an SS/VSS ratio of 70.3% (Fig. 5). The denitrifying reactor was stably operated for 8 days before the addition of Cu2+ , and the total nitrogen removal efficiency (NRE) and total organic carbon removal efficiency (ORE) were 93.9 ± 1.7 and 85.2 ± 1.6%, respectively. Approximately 1 mg L−1 Cu2+ was added to the synthetic wastewater on day 0 to start the continuous-flow test, which was run for a total duration of 82 days. Yang et al. [11] obtained an IC50 of 12.9 mg L−1 for Cu2+ in the anammox system. However, when 5 mg L−1 Cu2+ was added in the continuous-flow test, the biomass was strongly inhibited, and 94% of the microbial activity was lost. To prevent the overload of Cu2+ in this study, the concentration of Cu2+ was initially increased by

H. Chen et al. / Journal of Hazardous Materials 304 (2016) 448–456

453

Fig. 5. Performance of the denitrifying reactor. A, nitrogen removal, the solid line denotes NO3 − in the influent, the hollow diamond denotes NO3 − in the effluent, the solid circle denotes NH4 + in the effluent, the solid square denotes Cu2+ in the influent, the hollow star denotes total nitrogen removal efficiency; B, organic carbon removal, the solid line denotes total organic carbon in the influent, the hollow triangle denotes total organic carbon in the effluent, the solid star denotes total organic carbon removal efficiency, and the dashed line denotes hydraulic retention time.

1 mg L−1 . It was unexpected that the reactor maintained favorable performance during days 34 and 44 when the Cu2+ was 15 mg L−1 in the wastewater. The NRE and ORE during this stage were 95.4 ± 1.9 and 85.2 ± 3.3%, respectively. There was no sign of deterioration in the reactor performance. Considering the favorable performance during the previous 44 days and the IC50 of 30.6 ± 1.6 mg L−1 , the strategy of Cu2+ addition was changed. The Cu2+ in the wastewater was directly increased to 30 mg L−1 on day 45. In the following 14 days, the reactor efficiency decreased slightly, but the quality of the effluent remained acceptable, with average NREs and OREs of 89.1 ± 5.0 and 86.8 ± 4.6%, respectively. However, the average effluent NH4 + increased notably from 20.5 mg L−1 in the 8 days before the test to 37.2 mg L−1 in the previous 44 days, then to 71.2 mg L−1 at 30 mgCu2+ L−1 . There were two possible mechanisms for NH4 + accumulation. First, 30 mg L−1 Cu2+ led to microbial toxicity, and the cells started lysing, after which NH4 + was released into the water. Yang et al. [11] confirmed cell lysis in the anammox system by scanning and transmission electron microscopy after the addition of 5 mg L−1 Cu2+ . The second possibility was the occurrence of dissimilatory nitrate reduction to ammonium (DNRA) in the reactor. DNRA is a nitrogen-consuming process that transforms nitrate (NO3 − ) into ammonium (NH4 + ). Many researchers ascertained the existence of DNRA under anaer-

Fig. 6. Evolution of the NH4 + concentrations in the batch assays using denitrifying granules sampled on days 56 and 80.

454

H. Chen et al. / Journal of Hazardous Materials 304 (2016) 448–456

Fig. 7. Evolution of the morphology of the sludge after the addition of Cu2+ . A and B, size and color of the sludge on day −1 ; C, size and color of the sludge on day 56; D and E, size and color of the sludge on day 82.

H. Chen et al. / Journal of Hazardous Materials 304 (2016) 448–456

455

Table 3 Summary of the studies on the influence of Cu2+ on biological wastewater system in recent years. System

Cu2+ (mg L−1 )

Results

References

Phosphorus removal system

≤2.5

Phosphorus removal efficiency was maintained at 95 ± 2.7% Only 79% phosphorus removal efficiency remained COD removal between 60% and 80% and NH4 + -N removal of approximately 20% 50% inhibition of the process The activity of the biomass decreased by 94% The activity of the granules decreased by 50% No notable inhibition The accumulation of both NO2 − and NO3 − was stimulated, and the nitrite accumulation ratio decreased from 63.4% to 39.3%

[45]

3 Aerobic granular system

15

Nitrification process Anammox process Fermentative hydrogen production Sulfate reducing process Shortcut biological nitrogen removal

0.08 5 30 10–30 5

obic conditions [41,42]. In this study, NH4 + in the effluent was 20.3 mg L−1 before the addition of Cu2+ , and the concentration gradually increased with the presence of Cu2+ . Despite the increase of NH4 + in the effluent, the reactor retained acceptable removal efficiency. Hence, Cu2+ was gradually increased from 30 to 35, 50 and 75 mg L−1 in the following period. The NRE and ORE during these days were 91.0 ± 3.4 and 86.9 ± 3.8%, respectively, which were not significantly different from the vales obtained in the former experiments. NH4 + in the effluent was at an average level of 54.0 mg L−1 . Due to the surprising results achieved in the study, the concentration of Cu2+ was not further increased. In this study, the performance of the denitrifying reactor was determined with or without Cu2+ under a constant nitrogen loading rate (NLR) of 7.5 kgN m−3 d−1 . The NLR was relatively low compared with the maximum one obtained by Li et al. [43], which reached 35.1 kgN m−3 d−1 . The tolerance of the denitrifying reactor to Cu2+ stress may be due to the special structure of the granule, which consists of two distinct regions. The outer region is bonded together by readily-extractable EPS and dispersible due to the particles adhering through weak interactions, i.e., EPS ion bridging via multivalent ions and van der Waals forces. However, the inner region, which is composed of relatively compact EPS, is stable, and particles interact strongly through polymer entanglement [44]. Herewith, the multiple-region structure of the granules delayed and hindered the transportation of Cu2+ to the denitrifying granule cores. Selfadaptation of the denitrifying biomass to Cu2+ may be another factor that contributes to the tolerance to Cu2+ stress. The dosage of Cu2+ in the wastewater was progressively increased. Herewith, the denitrifying biomass adapted to the presence of Cu2+ with longterm acclimatization. Table 3 summarizes the response of various wastewater treatment techniques to the long-term influence of Cu2+ in the previous studies. The toxicity ranges of Cu2+ to biological processes depend on the type of the sludge, the operation time and other operation conditions. 3.2.2. SDA of the denitrifying biogranules during the long-term test The SDA of the biomass was monitored at three time points, i.e., at the beginning of the test, on day 56 (30 mg L−1 Cu2+ ) and day 80 (75 mg L−1 Cu2+ ). The corresponding SDAs were 13.7 ± 1.2, 14.7 ± 1.8 and 16.0 ± 2.1 mgN h−1 g− 1 VSS, respectively. Surprisingly, the SDA of the biomass increased instead of decreased with the long-term addition of Cu2+ in the wastewater. NH4 + was monitored during the batch assays (data not shown). NH4 + was under the detection limit in the assays with the biomass cultivated without Cu2+ addition. However, NH4 + continuously increased in the assays on days 56 (30 mg L−1 Cu2+ in the wastewater) and 80 (75 mg L−1 Cu2+ in the wastewater) (Fig. 6).

[46] [47] [11] [19] [48] [49]

The changes in the NH4 + evolution in the batch assays and the gradual increase of the SDA in the experiment indicated that the addition of Cu2+ in the continuous-flow test did not kill or even inhibit the denitrifiers. Instead, the metabolic pathway might be varied due to the presence of Cu2+ . However, the result achieved in the continuous-flow test was inconsistent with previous studies in other systems and, to the best of our knowledge, is the first time that a denitrifying reactor remained stable under such a high level of Cu2+ . Wang et al. [45] found that denitrifying polyphosphateaccumulating organisms (DPAOs) sludge adapted to the long-term addition of 2.5 mg L−1 Cu2+ . However, DPAO sludge activity was noticeably inhibited when the Cu2+ addition was increased to 3 mg L−1 . Moreover, Yang et al. [11] reported that 5 mg L−1 Cu2+ led to a 94% loss of microbial activity during the long-term test, and they suggested that Cu2+ could damage the cell membrane, resulting in the loss of anammox activity. The differences between our results and those of previous studies may due to the sludge type. In this study, although sludge lysis occurred, no inhibition of the denitrifying sludge was observed. Before the measurement of the SDA, the denitrifying sludge was stirred and washed three times. As discussed above, sorption plays an important role in metal uptake [35], Cu2+ may have been removed from the sludge during stirring and washing. Hence, the remaining concentration of Cu2+ significantly decreased to a relatively low level, and eventually, the SDA was not disturbed. The results of the long-term test differed from those of the pre-exposure test, possibly because (1) only 10 mL of sludge was treated during the pre-exposure tests, whereas 1 L of sludge was continuously exposed to Cu2+ in the continuous-flow test, and (2) Cu2+ was sealed in the serum bottles together with the sludge, whereas Cu2+ was flowing in the continuous-flow test; this might determine the distribution depth of Cu2+ in the sludge. 3.2.3. Evolution of the Cu2+ content in the sludge Sludge samples were taken and stored at three time points, i.e., on days -1, 56 and 82, to detect the evolution of the Cu2+ content in the sludge. The result showed an apparent increase in Cu2+ content in the sludge from 0.31 ± 0.01 to 12.38 ± 0.08 and eventually to 16.96 ± 0.07 mg g−1 SS. Because of the accumulation of Cu2+ in the sludge, the surface color of the sludge changed from brown to blackish green and even black, and the sludge broke into small particles (Fig. 7). The addition of Cu2+ would affect the morphology of the sludge and was likely due to the toxicity of Cu2+ , according to Yang et al. [11]. 4. Conclusions Both the short- and long-term effects of Cu2+ on denitrification were investigated in this study. The denitrifying reactor remained stable with Cu2+ concentrations up to 75 mg L−1 , possibly due to long-term acclimation and the special structure of the gran-

456

H. Chen et al. / Journal of Hazardous Materials 304 (2016) 448–456

ule, which could delay and hinder the transportation of Cu2+ into the denitrifying core. Moreover, pre-exposure of the denitrifying biogranules to Cu2+ could lead to an apparent decrease in SDA. These findings can provide utilizable information for denitrification of wastewater containing copper. Acknowledgement The authors thank the National Key Technologies R&D Program of China (No. 2012BAC13B02) for its partial support of this study. References

[23] [24]

[25] [26] [27] [28] [29]

[1] Y.X. Zhao, C.P. Feng, Q.H. Wang, Y.N. Yang, Z.Y. Zhang, N. Sugiura, Nitrate removal from groundwater by cooperating heterotrophic with autotrophic denitrification in a biofilm-electrode reactor, J. Hazard. Mater. 192 (2011) 1033–1039. [2] L. Zhu, J. Jin, H.Z. Lin, K.T. Gao, X.Y. Xu, Succession of microbial community and enhanced mechanism of a ZVI-based anaerobic granular sludge process treating chloronitrobenzenes wastewater, J. Hazard. Mater. 285 (2015) 157–166. [3] E. Sahinkaya, A. Kilic, B. Calimlioglu, Y. Toker, Simultaneous bioreduction of nitrate and chromate using sulfur-based mixotrophic denitrification process, J. Hazard. Mater. 262 (2013) 234–239. [4] Z.Q. Shen, Y.X. Zhou, J. Hu, J.L. Wang, Denitrification performance and microbial diversity in a packed-bed bioreactor using biodegradable polymer as carbon source and biofilm support, J. Hazard. Mater. 250–251 (2013) 431–438. [5] G.X. Wu, X.F. Zhai, C.G. Jiang, Y.T. Guan, Denitrifying kinetics and nitrous oxide emission under different copper concentrations, Water Sci. Technol. 69 (4) (2014) 746–754. [6] H. Chen, L.L. He, A.N. Liu, Q. Guo, Z.Z. Zhang, R.C. Jin, Start-up of granule-based denitrifying reactors with multiple magnesium supplementation strategies, J. Environ. Manage. 155 (2015) 204–211. [7] X. Jin, F. Wang, G. Liu, N. Yan, A key cultivation technology for denitrifying granular sludge, Process Biochem. 47 (2012) 1122–1128. [8] M.K. De Kreuk, M.C.M. van Loosdrecht, Selection of slow growing organisms as means for improving aerobic granular sludge stability, Water Sci. Technol. 49 (11–12) (2004) 9–17. [9] F.G. Fermoso, J. Bartacek, S. Jansen, P.N.L. Lens, Metal supplementation to UASB bioreactors: from cell–metal interactions to full-scale application, Sci. Total Environ. 407 (2009) 3652–3667. [10] D. Karadag, J.A. Puhakka, Enhancement of anaerobic hydrogen production by iron and nickel, Int. J. Hydrogen Energy 35 (2010) 8554–8560. [11] G.F. Yang, W.M. Ni, K. Wu, H. Wang, B.E. Yang, X.Y. Jia, R.C. Jin, The effect of Cu(II) stress on the activity, performance and recovery on the Anaerobic Ammonium-Oxidizing (Anammox) process, Chem. Eng. J. 226 (2013) 39–45. [12] J.B. Glass, V.J. Orphan, Trace metal requirements for microbial enzymes involved in the production and consumption of methane and nitrous oxide, Front. Microbiol. 3 (61) (2012). [13] X. Zhu, Y. Chen, H. Chen, X. Li, Y. Peng, S. Wang, Minimizing nitrous oxide in biological nutrient removal from municipal wastewater by controlling copper ion concentrations, Appl. Microbiol. Biotechnol. 97 (2013) 1325–1334. [14] A. Daverey, Y.C. Chen, Y.C. Liang, J.G. Lin, Short-term effects of monoethanolamine and copper on the activities of Anammox bacteria, Sustain. Environ. Res. 24 (2014) 325–331. [15] K. Gernaey, L. Verschuere, L. Luyten, W. Verstraete, Fast and sensitive acute toxicity detection with an enrichment nitrifying culture, Water Environ. Res. 69 (1997) 1163–1169. [16] M. Gutierrez, J. Etxebarria, L. De las Fuentes, Evaluation of waste water toxicity: comparative study between Microtox® and activated sludge oxygen uptake inhibition, Water Res. 36 (2002) 919–924. [17] G.B. Li, D. Puyol, J.M. Carvajal-Arroyo, R. Sierra-Alvarez, J.A. Field, Inhibition of anaerobic ammonium oxidation by heavy metals, J. Chem. Technol. Biotechnol. 90 (2015) 830–837. [18] Y.P. Ting, H. Imai, S. Kinoshita, Effect of shock-loading of heavy metals on total organic carbon and phosphate removal in an anaerobic-aerobic activated sludge process, World J. Microbiol. Biotechnol. 10 (1994) 308–312. [19] C. Li, H.H.P. Fang, Inhibition of heavy metals on fermentative hydrogen production by granular sludge, Chemosphere 67 (2007) 668–673. [20] A. Franco, E. Roca, J.M. Lema, Granulation in high-load denitrifying upflow sludge bed (USB) pulsed reactors, Water Res. 40 (2006) 871–880. [21] APHA, AWWA, WPCF, Standard Methods for the Examination of Water and Wastewater, 21st ed., American Public Health Association, Washington, DC, USA, 2005. [22] J.Q. Zhang, P. Zheng, M. Zhang, H. Chen, T.T. Chen, Z.F. Xie, J. Cai, G. Abbas, Kinetics of substrate degradation and electricity generation in anodic

[30]

[31] [32] [33]

[34] [35]

[36] [37]

[38] [39]

[40]

[41]

[42]

[43]

[44]

[45]

[46]

[47]

[48]

[49]

denitrification microbial fuel cell (AD-MFC), Bioresour. Technol. 149 (2013) 44–50. C.Y. Lin, Effect of heavy-metals on volatile fatty-acid degradation in anaerobic digestion, Water Res. 26 (1992) 177–183. V.P. Utgikar, B.Y. Chen, N. Chaudhary, H.H. Tabak, J.R. Haines, R. Govind, Acute toxicity of heavy metals to acetate-utilizing mixed cultures of sulfate-reducing bacteria: EC100 and EC50, Environ. Toxicol. Chem. 20 (2001) 2662–2669. C.C. Yeager, Copper and zinc preservatives, in: S.S. Block (Ed.), Disinfection, Sterilization, and Preservation, Lea & Febiger, Philadelphia, 1991, pp. 358–361. C.L. Dupont, G. Grass, C. Rensing, Copper toxicity and the origin of bacterial resistance-new insights and applications, Metallomics 3 (2011) 1109–1118. D.H. Nies, Microbial heavy-metal resistance, Appl. Microbiol. Biotechnol. 51 (1999) 730–750. P. Zheng, X.S. Feng, Biotechnology for Wastes Treatment, China Higher Education Press, 2006, pp. 99–106. J.M. Carvajal-Arroyo, D. Puyol, G.B. Li, A. Lucero-Acuna, R. Sierra-Alvarez, J.A. Field, Pre-exposure to nitrite in the absence of ammonium strongly inhibits anammox, Water Res. 48 (2014) 52–60. Y. Lu, H.E. Allen, Characterization of copper complexation with natural dissolved organic matter (DOM)-link to acidic moieties of DOM and competition by Ca and Mg, Water Res. 36 (2002) 5083–5101. R. Vulkan, U. Mingelgrin, J. Ben-Asher, H. Frenkel, Copper and zinc speciation in the solution of a soil-sludge mixture, J. Environ. Qual. 31 (2002) 193–203. T.D. Hayes, T.L. Theis, The distribution of heavy metals in anaerobic digestion, J. (Water Pollut. Control Fed.) 50 (1978) 61–72. L. Otero-González, J.A. Field, R. Sierra-Alvarez, Inhibition of anaerobic wastewater treatment after long-term exposure to low level of CuO nanoparticles, Water Res. 58 (2014) 160–168. C. White, G.M. Gadd, Copper accumulation by sulfate reducing bacterial biofilms, FEMS Microbiol. Lett. 183 (2000) 313–318. Z. Hu, K. Chandran, D. Grasso, B.F. Smets, Impact of metal sorption and internalization on nitrification inhibition, Environ. Sci. Technol. 37 (2003) 728–734. X. Wu, S.T. Liu, G.L. Dong, X.L. Hou, The starvation tolerance of anammox bacteria culture at 35 ◦ C, J. Biosci. Bioeng. 120 (2015) 450–455. J.M. Carvajal-Arroyo, D. Puyol, G.B. Li, A. Swartwout, R. Sierra-Álvarez, J.A. Field, Starved anammox cells are less resistant to NO2 − inhibition, Water Res. 65 (2014) 170–176. Z.W. Wang, Y. Li, J.Q. Zhou, Y. Liu, The influence of short-term starvation on aerobic granules, Process Biochem. 41 (2006) 2373–2378. J.P. Hernandez, L.E. de-Bashan, Y. Bashan, Starvation enhances phosphorus removal from wastewater by the microalga Chlorella spp. co-immobilized with Azospirillum brasilense, Enzyme Microb. Technol. 38 (2006) 190–198. J.A. Torà, J. Lafuente, J.A. Baeza, J. Carrera, Long-term starvation and subsequent reactivation of a high-rate partial nitrification activated sludge pilot plant, Bioresour. Technol. 102 (2011) 9870–9875. B. Kartal, M.M.M. Kuypers, G. Lavik, J. Schalk, H.J.M. Op den Camp, M.S.M. Jetten, M. Strous, Anammox bacteria disguised as denitrifiers: nitrate reduction to dinitrogen gas via nitrite and ammonium, Environ. Microbiol. 9 (2007) 635–642. F. Sgouridis, C.M. Heppell, G. Wharton, K. Lansdown, M. Trimmer, Denitrification and dissimilatory nitrate reduction to ammonium (DNRA) in a temperate re-connected floodplain, Water Res. 45 (2011) 4909–4922. W. Li, P. Zheng, L. Wang, M. Zhang, H.F. Lu, Y.J. Xing, J.Q. Zhang, R. Wang, J. Song, A. Ghulam, Physical characteristics and formation mechanism of denitrifying granular sludge in high-load reactor, Bioresour. Technol. 142 (2013) 683–687. Z.Z. Zhang, Q.Q. Zhang, Q. Guo, Q.Q. Chen, X.Y. Jiang, R.C. Jin, Anaerobic ammonium-oxidizing bacteria gain antibiotic resistance during long-term acclimatization, Bioresour. Technol. 192 (2015) 756–764. Y.Y. Wang, J. Qin, S. Zhou, X.M. Lin, L. Ye, C.K. Song, Y. Yan, Identification of the function of extracellular polymeric substances (EPS) in denitrifying phosphorus removal sludge in the presence of copper ion, Water Res. 73 (2015) 252–264. X.H. Wang, L.H. Gai, X.F. Sun, H.J. Xie, M.M. Gao, S.G. Wang, Effect of long-term addition of Cu(II) and Ni(II) on the biochemical properties of aerobic granules in sequencing batch reactors, Appl. Microbiol. Biotechnol. 86 (2010) 1967–1975. S.R. Juliastuti, J. Baeyens, C. Creemers, D. Bixio, E. Lodewyckx, The inhibitory effects of heavy metals and organic compounds on the net maximum specific growth rate of the autotrophic biomass in activated sludge, J. Hazard. Mater. 100 (2003) 271–283. H.T.Q. Kieu, E. Müller, H. Horn, Heavy metal removal in anaerobic semi-continuous stirred tank reactors by a consortium of sulfate-reducing bacteria, Water Res. 45 (2011) 3863–3870. J. Yin, H.J. Xu, D.S. Shen, K. Wang, Y. Lin, Effect of Cu(II) shock loads on shortcut biological nitrogen removal in a hybrid biofilm nitrogen removal reactor, Biodegradation 26 (2015) 211–222.