Does zinc in livestock wastewater reduce nitrous oxide (N2O) emissions from mangrove soils?

Does zinc in livestock wastewater reduce nitrous oxide (N2O) emissions from mangrove soils?

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Available online at www.sciencedirect.com

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Does zinc in livestock wastewater reduce nitrous oxide (N2O) emissions from mangrove soils? Guang C. Chen a,b, Nora F.Y. Tam b,c,*, Yong Ye d a

Third Institute of Oceanography, State Oceanic Administration, Xiamen, Fujian, China Department of Biology and Chemistry, City University of Hong Kong, Hong Kong, China c State Key Laboratory in Marine Pollution, City University of Hong Kong, Hong Kong, China d Key Laboratory of the Ministry of Education for Coastal and Wetland Ecosystem, College of the Environment and Ecology, Xiamen University, Xiamen, Fujian, China b

article info

abstract

Article history:

Zinc (Zn) affects nitrogen cycling but the effect of Zn in wastewater on the emission of

Received 17 February 2014

nitrous oxide (N2O) from the soil has not been reported. This study compared N2O emis-

Received in revised form

sions from mangrove soil receiving livestock wastewater containing various Zn2þ con-

15 July 2014

centrations and evaluated how long the effects of Zn would last in these soil-wastewater

Accepted 3 August 2014

microcosms. Significant increases in N2O flux were observed soon after the discharge of

Available online 12 August 2014

wastewater with a low Zn content. On the other hand, the flux was reduced significantly in the wastewater with high Zn levels but such inhibitory effect was not observed after tidal

Keywords:

flushing. Continuous monitoring of the N2O fluxes also confirmed that the inhibitory effect

Zinc

of Zn was confined within a few hours and the fluxes recovered in 6e9 h after the

Nitrous oxide

wastewater was completely drained away. These results indicated that the inhibitory ef-

Mangrove

fect of Zn on N2O fluxes occurred immediately after wastewater discharge and disappeared

Nitrification-denitrification

gradually. In the surface soil, nitrate levels increased with the addition of wastewater but

Wastewater

there was no significant accumulation of NHþ 4  N, irrespective of the Zn content in the wastewater. The study also showed that nitrification potential and immediate N2O emissions were inhibited by high Zn levels in the soil, but the total oxidation of ammonium to nitrate was not affected. © 2014 Elsevier Ltd. All rights reserved.

1.

Introduction

Nitrous oxide (N2O) is one of the key reactive greenhouse gases (GHGs) contributing to global warming, with a direct global warming potential 296 times that of carbon dioxide over

a 100-year time period (IPCC, 2001). In recent years, there has been a growing interest in quantifying the atmospheric N2O emission from coastal wetlands and investigating the regulating factors (Allen et al., 2007; Chmura et al., 2011; Chen et al., 2012), as wetland soil favors nitrificationdenitrification and could be a significant source of N2O.

* Corresponding author. Permanent address: Department of Biology and Chemistry, City University of Hong Kong, Kowloon, Hong Kong, China. Tel.: þ852 34427793; fax: þ852 34420522. E-mail address: [email protected] (N.F.Y. Tam). http://dx.doi.org/10.1016/j.watres.2014.08.003 0043-1354/© 2014 Elsevier Ltd. All rights reserved.

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Mangrove soil has been recognized as a major marine contributor to N2O emissions (Allen et al., 2007; Chen et al., 2012). The emissions from mangrove wetlands are further enhanced by excess nutrient inputs from human activities (Chen et al., 2011). The rapid expansion of coastal populations and consequent rapid development would discharge increasing amounts of waste and wastewater, leading to elevated levels of nutrients and heavy metals in mangrove wetlands (Tam and Wong, 2000; Trott and Alongi, 2000). Mangrove wetlands have recently been identified as an effective means of purifying wastewater, especially for removing nutrients from various types of wastewater rich in carbon and nitrogen, such as shrimp pond effluent, livestock wastewater and municipal sewage (Ye et al., 2001; Gautier et al., 2001; Wu et al., 2008). Direct discharge of wastewater, particularly from livestock or industrial sources, not only adds nutrients but also introduces excess heavy metals into mangrove wetlands. Zinc (Zn) is a common component in livestock wastewater, as animal feed is frequently rich in Zn (Abe et al., 2012). Zn is essential for microorganisms and functions as a co-factor for metallo-proteins and enzymes (Eiland, 1981; Doelman et al., 1994). However, high concentrations of Zn are toxic to organisms through the displacement of essential metals from their native binding sites or through ligand interactions (Bruins et al., 2000). Zn was also toxic to the microorganisms involved in N-cycling and Nmineralization in Zn-spiked or Zn-contaminated soil (Cela  squez-Murrieta et al., 2006; De and Sumner, 2002; Va ~ es et al., 2007; Mertens et al., Brouwere et al., 2007; Magalha 2007a; Ruyters et al., 2010a,b). This toxicity during the microbial process was observed even at Zn levels that were close to background (Wilson, 1977). Nonetheless, studies reporting the effects of Zn contamination on the conversion of inorganic nitrogen to N2O via nitrification and denitrification processes and the atmospheric emission of N2O from marine environment remain limited. Knowledge on how Zn affects N2O emissions from constructed wetlands for wastewater treatment is even scarcer, despite the fact that these emissions could be significantly enhanced by wastewater rich in nitrogen. The present study aims to test whether Zn in wastewater would reduce N2O emission from mangrove soils and how long the Zn-induced reduction on the emission would last, by comparing N2O fluxes from mangrove soil receiving livestock wastewater containing various Zn2þ concentrations in soilwastewater microcosms. It is hypothesized that N2O emission from mangrove soil would be enhanced by the nitrogen in livestock wastewater compared to the control with tap water, but the presence of Zn, especially at high content, in the wastewater would reduce the enhancement effect of nitrogen on N2O emission.

removing the visible stones, benthic fauna and plant tissues, a sample of fresh soil (equivalent to approximately 1.1 kg dry weight) was transferred into a plastic pot with a diameter of 14 cm. A total of 24 pots were prepared, and all pots were submerged into buckets containing artificially prepared seawater at a salinity of 20 parts per thousand for 1 day to simulate a 24-h high tide. The seawater was then drained away by gravity, and the pots were exposed for another day to simulate a 24-h low tide. The artificial seawater was prepared by dissolving commercial sea salt (Instant Ocean, Aquarium Systems, Mentor, Ohio) in deionized water. This artificial seawater had the concentrations of total organic carbon (TOC), ammonium-nitrogen ðNHþ 4  NÞ, nitrogen oxides 3 ðNO x  NÞ and inorganic phosphate ðPO4  PÞ at 5.2, 1.9, <0.05 1 and 0.2 mg g , respectively (Ye et al., 2001). The pots were subjected to this tidal regime for 10 days prior to wastewater treatment. Prior to wastewater addition, four pots were randomly retrieved to assess the N2O fluxes and soil characteristics, and these data were considered as the background values. The remaining 20 pots were divided into five treatment groups, each with four replicates (Table 1). The first group was the control and was treated with tap water (TW). The other four groups, labeled as LS, Z1, Z2 and Z3 treatments, were the synthetic livestock wastewater containing various amounts of zinc chloride. The synthetic wastewater had similar nutrient concentrations as real livestock sewage in Hong Kong (Ye et al., 2001) and was artificially prepared by adding glucose, ammonium chloride, sodium nitrate, urea and potassium dihydrogen phosphate to deionized water (Table 1). LS had the same Zn content as that in TW, while the nominal concentrations of Zn2þ ions in the Z1, Z2 and Z3 wastewaters were 10, 40 and 80 mg L1, respectively, and the respective measured values were 9.8, 34.3 and 65.9 mg L1. The highest Zn level applied in this study, 80 mg L1, was the maximum level recorded in livestock influent in Hong Kong SAR (Chen et al., 2011). During the experiment, all the pots were subjected to the above tidal regime, and wastewater additions were performed on alternating days, during the low tide days (Table 2). In other words, the wastewater or tap water additions were done on odd days (except on 7th, 33rd, 53rd, 73rd, 93rd and 113th days), and 100 mL of each type of wastewaters or tap water was added to the soil surface in each pot half an hour after the start of the low tide period (at approximately 9:00 am). The

Table 1 e Chemical composition of the wastewaters and tap water used in this study. Concentration (mg L1)

2.

Materials and methods

2.1.

Experimental setup

Mangrove soil was collected from Sai Keng (22 250 N, 114 160 E), a typical mangrove swamp in Sai Kung, Hong Kong SAR. The substrate was sandy, consisting of 73.7% sand, 14.8% silt and 11.6% clay and had a pH of 6.39 (Tam and Wong, 1998). After

pH TOC NHþ 4 N NO 3 N Organic N PO3 4 P Zn2þ

Wastewater treatment groups TW

LS

Z1

Z2

Z3

7.72 2.8 0.4 1.25 e 0.3 1.6

6.80 65.9 36.1 2.5 20 55 1.6

6.72 65.9 36.1 2.5 20 55 10

6.38 65.9 36.1 2.5 20 55 40

6.15 65.9 36.1 2.5 20 55 80

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experiment lasted for 115 days, and a total of 5.2 L wastewater or tap water was added to each pot. The total inputs of Zn in Z1, Z2 and Z3 wastewaters were 46.4, 161.8 and 311.5 mg Zn kg1 soil, respectively. During the entire experiment, the wastewater and tap water always leached through the soil within 1.5 h in all treatments, mainly because the soil used in this study was sandy (around 74% are sandy particle) and did not become slurry during the irrigation.

2.2.

Determination of N2O flux

During the dates with wastewater addition, N2O emissions from mangrove soils were measured 4 h after the addition of wastewaters or tap water to ensure that the added waters had soaked into the soil, leached through and completely drained out, that is, approximately 2.5 h after the water had completely leached through the soil. The N2O flux is defined as the “after-discharge flux” to indicate the immediate effect of wastewater addition on gas emission. On the other hand, on the 7th, 33rd, 53rd, 73rd, 93rd and 113th days, the wastewater or tap water additions were not performed and the N2O fluxes were also measured approximately 4 h after tidal flushing. This N2O flux was considered as the “non-discharge flux” which could reflect the effects of ammonia and Zn additions on N2O emission after tidal flooding. This single time measurement was common when determining the emission rates of N2O gas in other constructed wetland treatment systems (Fey et al., 1999). To determine whether the Zn effect was transient or not, N2O fluxes were monitored at 2, 4, 8, 14 and 23 h after the wastewater addition on the 51st, 71st, 91st and 111th days. During each measurement, the pot was placed in an air-tight container with a diameter of 20 cm and an internal volume of 3.5 L (Supplementary Fig. 1). Gas sampling was performed immediately after the container was covered, then every 10 min for a half an hour. During each measurement, a 5 mL gas sample was collected by inserting a 10 mL glass syringe with a hypodermic needle through an air sampling port at the top of container. The concentration of N2O was determined using a Hewlett Packard 6890A gas chromatograph (GC) equipped with a 63Ni electron capture detector (mECD) and an RT-Q Plot column (Restek). Gas analysis and the calculation of the soil-to-air N2O flux were based on the methods described by Chen et al. (2011).

2.3.

NA: not applicable.

NA single-time measurement, 4 h after wastewater addition single-time measurement, 4 h after tidal flushing Continuous measurements, 2, 4, 8, 14 and 23 h after wastewater addition Continuous measurements, 2, 4, 8, 14 and 23 h after tidal flushing NA NA Flushing pots with tidal water Determination of after-discharge N2O flux Determination of non-discharge N2O flux 24-h monitoring after addition of wastewater 24-h monitoring after tidal flushing Wastewater leachate collection Soil sample collection

Odd days, i.e. on alternative days from 1st to 115th days, except on 7th, 33rd, 53rd, 73rd, 93rd and 113th days Even days, i.e. on alternative days from 2nd to 112th days On the 5th, 31st, 51st, 71st, 91st and 111th days 7th, 33rd, 53rd, 73rd, 93rd and 113th days 51st, 71st, 91st and 111th days 53rd, 73rd, 93rd and 113th days 1st, 21st, 41st, 61st, 81st and 101st days 115th day Wastewater or tap water addition

Table 2 e Wastewater addition, tidal flushing and sampling dates in this study.

Experimental action

NA

N2O measurements Dates

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Sampling and analyses of effluent and soil

During the experiment, the effluent that leached out from each pot was collected on the 1st, 21st, 41st, 61st, 81st and 101st days. This water sample was filtered through Whatman No. 42 filter paper, and the concentrations of inorganic nitro2þ  were determined using gen (NHþ 4  N and NO3  N) and Zn Flow Injection Analysis (FIA, Lachat QuikChem 8000, Lachat Instruments, USA) and inductively coupled plasma atomic emission spectrometry (ICP-AES, Optima 2100 DV, Perkin Elmer Inc.), respectively, following standard methods for water and wastewater analyses (Eation et al., 1995). After the last measurement of N2O flux (on the 113th day), an extra addition of wastewater or tap water was applied to each pot (on the 115th day), and a soil core was collected 2 h

w a t e r r e s e a r c h 6 5 ( 2 0 1 4 ) 4 0 2 e4 1 3

after the wastewater drained out using a handheld PVC corer. The soil core was separated into two layers: the surface (0e3 cm) and bottom (3e8 cm) layers. This separation was based on the observation that the surface 3 cm of soil appeared to be more oxic and gray in color, whereas the color of the bottom soil was blackish. The pH and electrical conductivity (Ec) in soil samples were measured as the values in soil water mixture (1:5 weight to deionized water ratio) by a hand-held Conductivity/TDS-pH/mV-Temperature Meter  (WP-81, TPS, Australia). NHþ 4  N and NO3  N in the fresh soil were extracted with 2 M potassium chloride, and their concentrations were determined using FIA. The nitrification potential activity (NPA) of the surface soil and the potential denitrification activity (PDA) of the bottom soil were determined according to the methods described by Chen et al. (2012). NPA was measured by mixing fresh soil sample (equivalent to about 10 g dry weight) with 20 mL lownutrient seawater containing sodium chlorate (200 mM), ammonium chloride (200 mM) and sodium phosphate (15 mM) in a 125 mL flask. The flask was covered with a perforated rubber stopper to allow for oxygen exchange and then incubated in the dark in a mechanical shaker at 25  C for 48 h. After incubation, the soil solution was filtered through a Whatman No. 1 filter paper and the NO 2  N concentration in the filtrate was measured by FIA. For PDA measurement, fresh soil

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sample (equivalent to about 10 g dry weight) was added to a 125 mL air-tight quick fit conical flask and 25 mL solution 1 glucose-C was consisting of 80 mg L1 NO 3  N and 72 mg L added. The air in the head space of the flask was replaced by nitrogen and then 10% of the head space nitrogen was replaced by acetylene. The soil solution was incubated under anaerobic condition at 25  C for 24 h. At the end of incubation, N2O concentration at the head space was sampled by a syringe and analyzed by GC. The PDA rate was calculated based on N2O concentration and gas phase volume. The concentration of total Zn in the air-dried soil sample after sieving through a 0.25 mm mesh sieve was measured using the Microwave/HNO3 digestion technique combined with ICP-AES. The certified Marine Sediment Reference Material (MESS-3, National Research Council Canada) was analyzed for quality assurance, and the recovery was 83.7%. The Zn concentration of the soil was not corrected, as the observed recovery was acceptable.

2.4.

Statistical methods

The differences in N2O flux between the two measurement periods (after-discharge vs. non-discharge) and among five wastewater treatments were tested using a parametric twoway analysis of variance (ANOVA), with sampling days as

Fig. 1 e N2O fluxes in the control and four wastewater treatment groups: (A) after-discharge fluxes on Day 0 (background) and the 5th, 31st, 51st, 71st, 91st and 111th days and (B) non-discharge fluxes on Day 0 and the 7th, 33rd, 53rd, 73rd, 93rd and 113th days. The means and standard deviations of four replicates are shown. Different letters on the same measurement day indicate a significant difference among the five treatments according to one-way ANOVA at p ≤ 0.05.

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3.

Results

indicating that the diurnal fluctuations of N2O flux were treatment-specific, i.e., varied from treatment to treatment. In the Z2 or Z3 treatment groups, the diurnal variations in the N2O flux on the four measurement days (the 51st, 71st, 91st and 111th days) followed similar trends, but the diurnal fluctuation in Z1 treatment showed some variations between the 51st day and the other three sampling days (the 71st, 91st and 111th days). Among the four livestock wastewater treatments, the fluxes in the Z2 and Z3 treatments were significantly lower than those in the LS and Z1 groups during the first two sampling times (2 and 4 h after the addition of wastewater), but the differences became less obvious as the sampling progressed. The non-discharge flux was low in general (less than 10 mmol m2 h1), and the diurnal variations in N2O flux for most treatments and on most sampling days were not significant (Fig. 3).

3.1.

N2O flux from mangrove soil

3.2.

the repeated measures. The same test was used to compare the differences in N2O fluctuation within 24 h after wastewater discharge among the five sampling times and wastewater treatments. The effects of wastewater treatment on the 2þ  of the leachates concentrations of NHþ 4  N, NO3  N and Zn were tested using one-way ANOVA test with sampling days as the repeated measures. The two-way ANOVA test was also employed to examine the differences in the concentrations of inorganic nitrogen and zinc in the soil among the wastewater treatments and between the two soil depths. If the difference was significant at p < 0.05, a post-hoc test was applied to determine where the differences lied. All statistical analyses were performed using SPSS 16.0 for Windows (SPSS Inc., USA).

At the beginning of the experiment, i.e., before wastewater addition, the background values of N2O flux were low, varying from 0.68 to 0.80 mmol m2 h1 (Fig. 1), and were similar among the five treatments (p > 0.05). The soil receiving tap water (TW) exhibited low levels of N2O flux, less than 1 mmol m2 h1 during the entire study, but significantly higher afterdischarge fluxes were detected following the additions of livestock wastewaters containing low Zn levels (10 mg L1) in LS and Z1 treatments. The enhancement effect of livestock wastewater on the N2O flux was reduced by the moderate and high concentrations of Zn, as the fluxes in the Z2 and Z3 treatments were not significantly different from those of the control (TW) on the 5th day. Nevertheless, the Z2 and Z3 treatments showed significantly higher N2O fluxes relative to the control (TW) at each sampling time from the 51st day onwards (Fig. 1), indicating that the effect of Zn disappeared as the experiment proceeded. The non-discharge N2O flux in the control (TW) remained at low levels, similar to the after-discharge emission, but N2O emissions in all livestock wastewater treatments were significantly higher than that in the control, irrespective of the wastewater Zn levels (Fig. 1). Nonetheless, the non-discharge N2O emissions in livestock wastewaters were significantly lower than the after-discharge values (F ¼ 54.2, p < 0.0001). After the addition of wastewater, the 24-h measurements of N2O flux in the control (TW) showed small diurnal fluctuations, and similar patterns were observed on the 51st, 71st, 91st and 111th days (Fig. 2). On the other hand, the N2O fluxes in the livestock wastewater treatment groups increased gradually over the course of 24 h, and the highest flux was observed at the last sampling time, that is, 23 h after the addition of wastewater, with only a few exceptions (Fig. 2). The ANOVA results indicated that N2O flux was significantly affected by time (F ¼ 82.6, p < 0.0001), treatment (F ¼ 154.9, p < 0.0001) and their interaction (F ¼ 11.8, p < 0.0001),

Leachate and soil characteristics

The NHþ 4  N concentrations of the leachates collected from TW were low, but significantly higher values were observed in the four livestock wastewater treatment groups, with mean concentrations ranging from 12.34 to 16.58 mg L1 (Table 3). The highest NHþ 4  N concentration was found in the Z3 treatment group, and the lowest was in the Z1 group (F ¼ 205.5, p < 0.001). Similar to NO 3  N that were below 0.4 mg L1 in all treatments, the concentrations of Zn in all leachates were also low throughout the experiment, even though large amounts of Zn were added in the Z2 (40 mg L1) and Z3 (80 mg L1) treatments through the addition of wastewaters. In average, more than 98% of the Zn was removed from the wastewater and transferred to soil after each wastewater addition during the experiment (Table 4). The percentages of transfer in the Z2 and Z3 treatments (with higher Zn concentrations in the wastewater) were even higher than that in the Z1 treatment with a low Zn concentration. These results indicated that mangrove soil had a strong capacity for retaining wastewater-borne Zn. High Zn levels in the wastewater significantly reduced the removal of NHþ 4 N but did not have any significant effect on NO 3  N removal (Table 4). At the end of the experiment, the concentrations of inor ganic nitrogen (NHþ 4  N and NO3  N) and Zn in the soil varied significantly with soil depth and wastewater treatment (Table 5). Two-way ANOVA revealed that the effects of soil depth, treatment and their interaction on the soil nitrogen and Zn concentrations were all significant at p  0.001, with the exception of the interaction factor for the soil NHþ 4 N   N and NO (F ¼ 0.68, p ¼ 0.61). The concentrations of NHþ 3  4 N in the surface soil were significantly higher than those in the bottom soil. The addition of wastewater caused a significant accumulation of NHþ 4  N in the bottom soil but not in the surface soil. On the contrary, the surface soils that received wastewater exhibited significantly higher levels of NO 3 N

Fig. 2 e Diurnal variations of N2O fluxes over 24 h (2, 4, 8, 14 and 23 h) after the addition of tap water or wastewater on the 51st, 71st, 91st and 111th days. The means and standard deviations of four replicates are shown. Different letters for the same treatment on each sampling day indicate a significant difference among the five sampling times according to one-way ANOVA at p ≤ 0.05. NS: not significant.

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Table 3 e Concentrations of inorganic nitrogen and Zn2þ (mg L¡1) in leachates collected from five treatments during the experiment.

TW LS Z1 Z2 Z3

NO 3 N

NHþ 4 N

Treatment

0.51 (0.05e1.62) a 12.34 (5.25e25.34) 14.47 (6.75e26.83) 15.17 (1.66e31.16) 16.58 (6.60e30.40)

0.11 (0.01e0.43) 0.24 (0.07e0.90) 0.23 (0.01e1.16) 0.32 (0.06e1.66) 0.33 (0.07e1.72)

b c c d

Zn2þ a ab ab b b

0.02 0.02 0.11 0.10 0.08

(0.01e0.03) (0.01e0.04) (0.05e0.28) (0.05e0.45) (0.02e0.24)

a a c c bc

Mean and range values (in brackets) of 24 samples, that is, four replicates at each sampling point with a total of six samplings on the 1st, 21st, 41st, 61st, 81st and 101st days, in each treatment during the experiment are shown. Different letters for each parameter indicate a significant difference among the five treatments according to one-way ANOVA at p  0.05.

Table 4 e Removal percentages of inorganic nitrogen and Zn2þ from livestock wastewater during the experiment.

LS Z1 Z2 Z3

NO 3 N

NHþ 4 N

Treatment 71.88 64.46 60.71 57.49

(42.03e90.73) (40.86e85.49) (31.28e86.09) (25.92e83.82)

c b b a

92.32 (74.46e97.79) 92.59 (81.37e98.99) 89.53 (53.49e97.95) 89.19 (44.24e97.45)

Zn2þ a a a a

e 98.58 (95.56e99.58) a 99.73 (98.98e99.98) b 99.91 (99.74e99.99) b

Means and range values (in brackets) of 24 samples, that is, four replicates at each sampling point with a total of six samplings on the 1st, 21st, 41st, 61st, 81st and 101st days, in each treatment during the experiment are shown. Different letters for each parameter indicate a significant difference among the four treatments according to one-way ANOVA at p  0.05. Removal percentage of each parameter was calculated based on the differences between the concentrations in input wastewater and in leachate.

than the control, but there was no difference among treatments in the bottom soil. The concentrations of total Zn in the surface soil increased with increasing Zn levels in the wastewater, in the order of TW ¼ LS < Z1 < Z2 < Z3, but in the bottom soil, significantly higher Zn was only observed in the Z3 treatment (Table 5). In the Z3 treatment group, the Zn concentration in the surface soil (449.4 mg g1) was significantly higher than that in the bottom soil (45.6 mg g1). These results indicated that Zn inputs from wastewater were mainly retained by the surface soil, with relatively little downward migration. Soil pH was only affected by wastewater treatment (F ¼ 82.61, p ¼ 0.000) but not soil depth. At the end of the experiment, Z1 treatment had the highest soil pH, followed by Z2, and Z3 treatment had the lowest pH that was even lower than the control and LS (Table 5), suggesting that the high Zn level in livestock wastewater increased soil acidity. There were no significant differences in soil electrical conductivity (Ec) among all treatments, which varied from 2.18 to 2.55 mv cm1 in the surface soil and the ranges were slightly lower (1.58e1.85 mv cm1) in the bottom soil (Table 5), indicating the addition of nitrogen and Zn in wastewater would not affect the conductivity values in soil. At the end of the experiment, the nitrification potential (NPA) in the surface soil differed significantly among treatments. The soils receiving livestock wastewaters had significantly higher NPA than the control (TW), and the Z3 treatment had lower activity relative to the other treatments (Fig. 4). A maximum value of NPA was found in the Z1 treatment,

followed by LS and Z2 treatments, suggesting that high Zn levels in wastewater would inhibit NPA activity. On the contrary, there was no significant difference in the soil potential denitrification activity (PDA) in the bottom soil among all treatments (Fig. 4).

4.

Discussion

Both nitrification and denitrification are important drivers for N2O emission from mangrove soil (Allen et al., 2007; Chen et al., 2011). In the present study, the livestock wastewater contained high levels of ammonium but low nitrate, and the final nitrate concentrations in the surface soil receiving livestock wastewater were higher than those of the control (which received tap water), but the ammonium values were all comparable (Table 5). These results indicated that the ammonium in the wastewater was oxidized to nitrate once it entered the soil, and nitrification was an important mechanism driving atmospheric N2O fluxes from mangrove soils in this study. The nitrification potential activity (NPA) in the surface mangrove soil was stimulated by the addition of ammonium-rich livestock wastewater (Fig. 4A), similar to previous studies in intertidal soils that reported changes in composition and activity of soil ammonia-oxidizing commu~es et al., 2005; Lage nity following N enrichment (Magalha et al., 2010). In the present study, the highest NPA was found in the Z1 treatment group (with low Zn content at 10 mg Zn L1), suggesting that nitrification was enhanced by

Fig. 3 e Diurnal variations of N2O fluxes over 24 h (2, 4, 8, 14 and 23 h) after the tidal flushing on the 53rd, 73rd, 93rd and 113th days. The means and standard deviations of four replicates are shown. Different letters for the same treatment on each sampling day indicate a significant difference among the five sampling times according to one-way ANOVA at p ≤ 0.05. NS: not significant.

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¡1  Table 5 e pH, electrical conductivity (Ec1:5, mv cm¡1), concentrations of NHþ ) in the surface 4  N, NO3  N and total Zn (mg g (0e3 cm) and bottom (3e8 cm) soil under various treatments at the end of the experiment.

Soil layer Surface

Bottom

Treatment TW LS Z1 Z2 Z3 TW LS Z1 Z2 Z3

Ec1:5

NHþ 4 N

2.18 (0.57) a 2.25 (0.13) a 2.36 (0.31) a 2.55 (0.25) a 2.18 (90.21) a 1.58 (0.24) a 1.74 (0.12) a 1.85 (0.11) a 1.82 (0.08) a 1.82 (0.17) a

2.27 (0.49) a 3.67 (0.52) a 3.08 (0.97) a 3.14 (0.56) a 2.85 (0.28) a 1.26 (0.36) a 2.97 (0.72) b 2.87 (0.39) b 2.18 (0.22) b 2.24 (0.59) b

pH 6.48 6.59 7.02 6.91 6.30 6.56 6.56 6.99 6.85 6.29

(0.07) (0.07) (0.12) (0.13) (0.12) (0.04) (0.08) (0.08) (0.05) (0.09)

b b c c a b b d c a

NO 3 N 0.32 (0.07) 1.06 (0.54) 1.07 (0.33) 1.47 (0.45) 1.30 (0.34) 0.36 (0.05) 0.50 (0.15) 0.56 (0.10) 0.51 (0.13) 0.40 (0.07)

a b b b b a a a a a

Zn2þ 32.0 28.9 57.7 165.7 449.4 29.4 26.9 28.8 30.6 45.6

(3.5) a (2.0) a (2.6) b (39.3) c (70.6) d (3.4) a (1.0) a (1.2) a (2.5) a (1.9) b

Means and standard deviations (in brackets) of four replicates are shown. Different letters for each parameter at the same soil layer indicate a significant difference among the five treatments according to one-way ANOVA at p  0.05.

the addition of appropriate amounts of Zn due to its function as an essential element for microorganisms. Ruyters et al. (2010a) also found that the NPA in terrestrial soil was enhanced by a small-dose Zn spiking at an approximately total Zn concentration of 100 mg g1. On the other hand, denitrification could also occur in this study as the mangrove soils were subjected to regular tidal flushing with alternating oxic and anoxic conditions. The potential denitrification activity (PDA) in the bottom soil at the

Fig. 4 e Activities of (A) nitrification potential (NPA) of the surface soil and (B) potential denitrification (PDA) of the bottom soil under various treatments. The means and standard deviations of four replicates are shown. Different letters indicate a significant difference among the five treatments according to one-way ANOVA at p ≤ 0.05. NS: not significant.

end of the experiment was comparable among all treatments and control (Fig. 4B), this may be due to the fact that the initial nitrate concentrations in both livestock wastewater and tap water control were relatively low, 2.5 and 1.5 mg N L1, respectively, when compared to ammonium. There were also no significant differences in the nitrate concentrations in the bottom soil among all treatments and control at the end of the experiment (Table 5). Since the PDA was only measured in the bottom soil, but not in the surface soil which was rich in nitrate, the roles of anaerobic denitrification, and even nitrifier denitrification, a pathway of nitrification in which ammonium is oxidized to nitrite, followed by a direct reduction of nitrite to N2O by autotrophic ammonium oxidizing bacteria (Wrage et al., 2001), in N2O emissions should be examined in future studies. The nitrification, and adsorption and hydrolysis of Zn could release protons (Hþ) into the soil solution, leading to acidic pH, while denitrification increase pH (McBride, 1994; Cela and Sumner, 2002; Lin et al., 2012; Ruyters et al., 2013). Therefore, soil pH would be a function of the pH and Zn in wastewater inputs, as well as the balance between nitrification and denitrification processes. In the present study, the highest soil pH was found in Z1 treatment, but this treatment also had the highest NPA in surface soil, suggesting that the denitrification in the surface soil might also be intensive in Z1 treatment. Wong (2004) also reported that ammonia addition could increase pH and the abundances of both nitrifiers and denitrifiers in surface mangrove soil. The highest level of Zn and the lowest water pH in Z3 treatment might contribute more to the lowest soil pH than the acidity produced from nitrification. Soil pH, on the other hand, was found to affect the N2O production in both nitrification and denitrification in soils (Granli and Bockman, 1994), and the proportion of the different nitrogen gases (N2O, NO and N2) released during denitrification as well (Wijler and Delwiche, 1954). However, the soil pH in the present study was close to neutral values (6.29e7.02), therefore, pH was unlikely to have any significant effect on N2O emission. The mangrove soil had a strong capacity for retaining wastewater-borne Zn which was mainly adsorbed onto the surface soil in this study, as heavy metals in wastewater could rapidly precipitate as low-solubility salts of sulfides, carbonates and oxyhydroxides in wetland soil (Peltier et al., 2003).

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Tam and Wong (1996) also found that most of heavy metals in wastewater accumulated in the surface layer after they were filtered through the mangrove soil. Zinc could reduce the N2O emission from soils due to its toxicity to nitrifier (De Brouwere et al., 2007; Ruyters et al., 2010a). In this study, lower afterdischarge N2O fluxes were found in the Z2 and Z3 treatment groups than the Z1 and LS groups (Figs. 1, 2 and 5) , as high Zn levels in the former two treatments significantly reduced the NPA in the surface mangrove soil (Fig. 4A). These results indicated that Zn in wastewater inhibited the oxidation of ammonium to nitrate, thereby reducing N2O emissions from mangrove soil. Cela and Sumner (2002) also reported that soil nitrification was partially inhibited when the soil waterextractable Zn2þ fraction ranged from 0.125 to 0.5 mg L1, and the process was completely inhibited at Zn2þ > 0.5 mg L1. Although soil nitrification was reduced by the high level Zn, the presence of other ions, such as proton, calcium (Ca2þ) and magnesium (Mg2þ), could protect microorganisms against Zn stress (Mertens et al., 2007a). The addition of ammonium to the soil also accelerated the adaptation of ammonia-oxidizing microorganisms to Zn toxicity (Ruyters et al., 2010a). Soil microorganisms could adapt to metal stresses by developing a variety of resistance mechanisms (Bruins et al., 2000). Mertens et al. (2007b) reported that the soil nitrifying bacterial community exposed to a long-term contamination of Zn was able to resist Zn toxicity, and the NPA in soil containing 260 or 960 mg Zn g1 was similar to that observed in

Fig. 5 e Relationships between cumulative NHþ 4  N inputs from livestock wastewater and (A) after-discharge N2O fluxes from mangrove soil and (B) non-discharge N2O fluxes from mangrove soil to show the effect of ammonium and Zn in wastewater on N2O emission.

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uncontaminated soil. In the present study, the inhibition of N2O emission by moderate (40 mg Zn L1) to high (80 mg Zn L1) levels of Zn took place immediately after the addition of wastewater but the Zn effect disappeared gradually over time, as reflected by the emission from the Z2 and Z3 treatment groups became comparable to that in the LS and Z1 treatment groups 6e12 h after the wastewater addition. These results suggested that the inhibited N2O emission from mangrove soil could be recovered quickly, probably in less than 12 h. The comparable non-discharge N2O fluxes among the four treatments (Figs. 1, 3 and 5) further confirmed that the effect of Zn was transient. Previous studies also reported similar transient Zn inhibitory effects on the nitrification rate in soil and found that the recovery was due to microbial adaptation associated with changes in the nitrifying microbial community, which could occur from less than one year to two years after Zn contamination (Rusk et al., 2004; Mertens et al., 2006; Ruyters et al., 2010a, 2013). It is possible that the nitrifying microorganisms, particularly ammonia oxidation bacteria (AOB), gradually adapted to Zn contamination, leading to a recovery of the N2O emissions. Not only nitrifying microorganisms, the denitrification process such as N2O reduction in soil, the last step of denitrification, was also inhibited by Zn toxicity, leading to more N2O emissions from the soil, but the process could be recovered with microbial adaptation and the development of Zn tolerant denitrifying community (De Brouwere et al., 2007; Ruyters et al., 2010b). The N2O production via nitrifier denitrification pathway was positively correlated to the total  squezconcentration of Zn in metal-contaminated soil (Va Murrieta et al., 2006). In the present study, since denitrification was important in N2O emission from mangrove soils, the inhibition of Zn on denitrification and its recovery could also affect the N2O emission patterns in this study. However, the Zn effect on denitrification and its recovery were not clear, as PDA was measured in the bottom soil where only Z3 treatment had more Zn than the other treatments. The responses of soil anaerobic heterotrophic denitrification and nitrifier denitrification to initial nitrogen concentrations in wastewater, with and without the presence of Zn, remained to be explored. Metal toxicity to nitrifying community to some extend depends on the soil cation exchange capacity (CEC), and decreases with increasing CEC due to decreasing metal bioavailability (Smolders et al., 2009). Although CEC was not determined in the present study, the measured soil conductivity (Ec) values, closely related with soil CEC, were not significantly different among all treatments (Table 5); therefore the effect of Ec on Zn availability and toxicity would not explain the adaption of microorganisms to Zn stress in this study. It is possible that the soil Ec was more affected by tidal seawater than the wastewater treatments as all microcosms were flooded with seawater in alternate days. Previous studies have demonstrated that nutrient inputs significantly increased the atmospheric greenhouse gas fluxes from mangrove soils (Allen et al., 2007; Chen et al., 2010, 2011). The present study further proved that the addition of wastewater rich in inorganic nitrogen significantly enhanced the N2O emission from mangrove soil, and the gas emission from mangrove soil receiving nitrogen-rich wastewater was still

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significant even under high Zn stresses. It is therefore essential to reduce discharges from various anthropogenic activities, particularly nutrient-rich wastewater, and local authorities must set up more stringent discharge standards. Regular monitoring of greenhouse gas emissions from mangrove wetlands, especially those in close proximity to humans, is recommended. Moreover, this study suggested that single time measurement might be inadequate in representing the response of soil N2O emission to the effect of nitrogen addition and Zn stress, and continuous monitoring is indispensable. Daelman et al. (2013) also suggested a higherfrequency sampling to monitor the short term variability of N2O emission, as well as a long-term sampling scheme covering more factors, such as temperature, for better estimation of N2O emission from wastewater treatment plants.

5.

Conclusions

The present study demonstrated that the discharge of ammonium and appropriate Zn levels in livestock wastewater enhanced N2O emission and stimulated the nitrification potential activity in mangrove soil, suggesting that nitrification was an important mechanism involved in this gas emission. This enhancement was reduced by the high Zn levels in wastewater immediately after the addition of wastewater, but the effect was transient that disappeared gradually over time with a relative fast recovery in N2O emission. However, the mechanisms behind the interactions of ammonium and Zn on N2O emission were not explored, and the residual effects of Zn accumulated in the soil were also not examined. More indepth researches on the relative importance of the different pathways involved in N2O emission, including nitrification, anaerobic heterotrophic denitrification and nitrifier denitrification, as well as the short- and long-term effect of Zn and nitrogen inputs at different levels in wastewater, on these pathways are needed.

Acknowledgments The work described in this paper was supported by the State Key Laboratory in Marine Pollution, City University of Hong Kong (Project number: 9360136) and the project on the degradation of persistent organic pollutants in Shenzhen Bay supported by the Science, Technology and Innovation Committee of the Shenzhen Municipality (Project Number: 9680078). The National Natural Science Foundation of China (41206108) also contributed. The authors are grateful to Mr. Benz Chan and Ms. Amy Chong for their technical support during the GC and FIA analyses, as well as Y. C. Kwan and Y. Y. Sha for their assistance with field and lab sampling.

Appendix A. Supplementary data Supplementary data related to this article can be found at http://dx.doi.org/10.1016/j.watres.2014.08.003.

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