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Environmental Pollution 255 (2019) 113219

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Environmental Pollution

journal homepage: www.elsevier .com/locate/envpol

Effects of shrimp-aquaculture reclamation on sediment nitratedissimilatory reduction processes in a coastal wetland of southeasternChina*

Dengzhou Gao a, b, Min Liu a, Lijun Hou b, *, Y.F. Lai Derrick c, Weiqi Wang d, Xiaofei Li d,Aying Zeng d, Yanling Zheng a, b, Ping Han a, b, Yi Yang a, b, Guoyu Yin a

a College of Geographical Sciences, East China Normal University, 500 Dongchuan Road, Shanghai, 200241, Chinab State Key Laboratory of Estuarine and Coastal Research, East China Normal University, 500 Dongchuan Road, Shanghai, 200241, Chinac Department of Geography and Resource Management, Institute of Environment, Energy and Sustainability, The Chinese University of Hong Kong, Shatin,New Territories, Hong Kong, Chinad Key Laboratory for Humid Subtropical Eco-geographical Processes of the Ministry of Education, Fujian Normal University, 8 Shangsan Road, Fuzhou,350007, China

a r t i c l e i n f o

Article history:Received 27 June 2019Received in revised form28 August 2019Accepted 7 September 2019Available online 11 September 2019

Keywords:DenitrificationAnaerobic ammonium oxidationDNRAEnvironmental implicationsShrimp aquacultureCoastal wetland

* This paper has been recommended for acceptanc* Corresponding author.

E-mail address: ljhou@sklec.ecnu.edu.cn (L. Hou).

https://doi.org/10.1016/j.envpol.2019.1132190269-7491/© 2019 Elsevier Ltd. All rights reserved.

a b s t r a c t

The conversion of natural saltmarshes to shrimp aquaculture ponds can potentially influence thebiogeochemical cycling of nutrients in coastal wetlands, but its impact on the dynamics of sedimentdissimilatory nitrate (NO3

�) reduction remains poorly understood. In this study, three sediment NO3�

reduction processes including denitrification (DNF), anaerobic ammonium oxidation (ANAMMOX), anddissimilatory NO3

� reduction to ammonium (DNRA) were examined simultaneously in a natural salt-marsh and two shrimp culture ponds (5- and 18-year-old) in July and November, using nitrogen (N)isotope-tracing experiments. Our results showed that sediment potential DNF, ANAMMOX and DNRArates were generally higher in the shrimp culture ponds than the natural saltmarsh in the two seasons.The rates of all three processes generally increased with the age of shrimp ponds, with the magnitude ofincrease being less pronounced for DNF and ANAMMOX than DNRA. The contribution of DNRA to totalNO3

� reduction increased significantly following saltmarsh conversion to shrimp ponds, suggesting thatDNRA became an increasingly important biogeochemical process under shrimp culture. DNRA competedwith DNF and limited reactive N loss to some extent after natural saltmarshes converted to shrimpculture ponds. The results of redundancy analysis revealed that the availability of substrates and sulfidesin sediments, rather than the bacteria gene abundance, were the most important factor influencing theNO3

� reduction processes. Overall, our findings highlighted that shrimp-aquaculture reclamation mayaggravate nitrogen loading in coastal wetlands by promoting the production of bioavailable ammonium.

© 2019 Elsevier Ltd. All rights reserved.

1. Introduction

Over the past several decades, global reactive nitrogen (N)production has increased substantially due to industrial and agri-cultural activities (Canfield et al., 2010). Approximately 20e30% ofthe reactive N enters eventually into estuarine and coastal eco-systems through rivers, groundwater runoff and atmosphericdeposition (Hardison et al., 2015; Chen et al., 2016a, b). Estuarine

e by Sarah Harmon.

and coastal saltmarshes, located at the interface between terrestrialand oceanic zones, provide many important ecosystem services,including the removal of excess reactive N (Hou et al., 2013;Wankelet al., 2017). The dominance of specific N-conversion pathwayshelps determine whether an ecosystem is a net sink of excessreactive N (Plummer et al., 2015; Zheng et al., 2017). Thus, microbialN transformation processes in coastal wetlands and their envi-ronmental implications have attracted much attention recently.

Among various N transformation processes, dissimilatory ni-trate (NO3

�) reduction is considered to be highly important inregulating reactive N in the coastal ecosystems (Herbert, 1999;Damashek and Francis, 2018). The reduction of NO3

� mainly

D. Gao et al. / Environmental Pollution 255 (2019) 1132192

involves three anaerobic metabolisms, namely denitrification(DNF), anaerobic ammonium oxidation (ANAMMOX), and dissimi-latory NO3

� reduction to ammonium (DNRA) (Thamdrup andDalsgaard, 2002; Roland et al., 2018). Both DNF and ANAMMOXwould lead to gaseous N production and a net loss of ecosystem N,while DNRAwould convert NO3

� into more bioavailable ammonium(NH4

þ) and cause a net retention of reactive N within the systems(Silver et al., 2005; Wankel et al., 2017). DNF has been found to be adominant microbial pathway of NO3

� reduction in many estuarineand coastal wetlands (Hou et al., 2015; Plummer et al., 2015), whilesome researchers have reported that DNRA is the main mechanismregulating reactive N in the coastal intertidal zones (Dong et al.,2011; Giblin et al., 2013; Cao et al., 2016). Numerous studies haveexamined the spatio-temporal variability of dissimilatory NO3

�

reduction processes across the estuarine and coastal zones, andfound that oxygen level, temperature, salinity, sulfide, and theavailability of carbon (C) and N substrates play crucial roles incontrolling the process rates (Dong et al., 2011; Deng et al., 2015;Plummer et al., 2015; Smith et al., 2015; Gao et al., 2017). Forinstance, although both DNF and ANAMMOX are well adapted tosuboxic/anaerobic environments, one process could be morefavorable than the other as a result of differences in substrateavailability (Dalsgaard et al., 2005; Babbin et al., 2014). While bothDNRA and DNF are competing for NO3

� and nitrite (NO2�), the former

and the latter processes are expected to dominate in environmentswith high concentrations of TOC and NO3

�, respectively (Dong et al.,2011). Sediment microbial communities could also mediate thebiogeochemical cycling of N, with the abundance of nirS, anammoxbacterial 16S rRNA (ANAMMOX 16S), and nrfA genes being indica-tive of DNF, ANAMMOX, and DNRA activities, respectively, incoastal wetlands (Zheng et al., 2016).

The reclamation of coastal saltmarshes for agriculture is acommon practice around the world, and has already caused aremarkable deterioration in the functions of wetland ecosystems(Barbier et al., 2011; Murray et al., 2019). The construction of land-based aquaculture ponds, being one of the most widespread typesof agricultural reclamation, can exert a profound and compleximpact on the nutrient cycles of wetlands owing to the significantchanges in hydrology and management modes (Wu et al., 2014;

Fig. 1. Location of the study

Murphy et al., 2016). Intensive shrimp aquaculture has becomeincreasingly popular globally in order to meet the rising demandfor seafood products (Yang et al., 2017a). Our previous work hasshown that shrimp aquaculture could significantly increase C and Nsubstrates due to the continuous supply of feeds (Gao et al., 2018).Also, shrimp ponds are found to dramatically alter the CH4 and N2Odynamics of coastal wetlands (Yang et al., 2017a). In general,intensive shrimp aquaculture is maintained through the dailyaddition of feeds. Only a small proportion (4.0e27.4%) of feeds isactually utilized by fishes and shrimps (Chen et al., 2016a, b), whilethe majority of residual feeds are accumulated in the sediments. Ahigh N loading in aquaculture ponds is detrimental to the culturedanimals and overall ecological health (Wu et al., 2014). Whilesediment dissimilatory NO3

� reduction can play a critical role inmediating N cycling, the effects of saltmarsh conversion to shrimpponds as well as other environmental factors on the relativeimportance of DNF, ANAMMOX and DNRA processes are still poorlyunderstood. Given that China has the world's largest aquacultureindustry with a total area and production of 2.6� 106 ha and2.3� 109 kg, respectively, in 2015 (Chen et al., 2016a, b), we con-ducted a research in the Min River estuary of southeastern China toinvestigate the effects of shrimp aquaculture on sediment DNF,ANAMMOX and DNRA processes. The specific objectives of thisstudy were (1) to investigate the changes in sediment NO3

� reduc-tion processes and the associated bacterial gene abundances alonga chronosequence of shrimp aquaculture ponds, (2) to elucidate themain environmental factors governing the sediment DNF, ANAM-MOX and DNRA processes, and (3) to assess the relative contribu-tions of three specific processes to total NO3

� reduction and theirenvironmental implications following the conversion of coastalsaltmarshes to shrimp ponds.

2. Materials and methods

2.1. Study area and sample collection

The studywas carried out in the Shanyutanwetland ofMin Riverestuary, southeast China (26�0003600e26�0304200N,119�3401200e119�4104000E, Fig. 1). This area has a southern

area and sampling sites.

D. Gao et al. / Environmental Pollution 255 (2019) 113219 3

subtropical monsoonal climate, withmean annual temperature andprecipitation of 19.6 �C and 1350mm, respectively (Zhang et al.,2015). This wetland has a high diversity of plant species, withCyperus malaccensis, Phragmites australis and exotic Spartina alter-niflora being the main vegetation species (Gao et al., 2017). Sedi-ments are mainly characterized as silt loam, with the volumecontents of 62e78% for silt, 19.05e37.39% for clay and 0.28e4.77%for sand (Zhang et al., 2015). Over the past decades, large-scaleconversion of coastal saltmarshes (mainly dominated byC. malaccensis) to shrimp culture ponds has occurred due to therising human population and demand for seafood (Yang et al.,2017a). Based on field investigation and satellite image analysis,the shrimp ponds in the study area were mainly reclaimed in twodifferent years in 1998 and 2011, respectively, which formed ashrimp culture chronosequence (Fig. 1). A common approach forthe reclamation of natural saltmarsh in the study area is to builddike through digging sediments at the edge of ponds, so the initialsediment physico-chemical characteristics in shrimp ponds weresimilar to natural saltmarshes. Therein, natural saltmarshes areinundated twice daily (a typical semidiurnal tide) with a tide rangeof approximately 2.5e6m (Zhang et al., 2015). The shrimp rearingperiod of these ponds spans from June to November, and the meanwater depth is maintained at about 1.4m. During the cultureperiod, the daily feeding rate is kept at approximately30e55 kg ha�1. Pond water is discharged into the adjacent sea aftershrimp harvest (Yang et al., 2017b).

To assess the influence of coastal saltmarsh conversion toshrimp culture ponds on sediment dissimilatory NO3

� reductionprocesses, sediment core samples (0e5 cm) were collected twice inJuly and November 2016 from six replicate locations in each of thethree types of stands, including the natural C. malaccensis salt-marsh, and the 18- and 5-year-old shrimp ponds that werereclaimed in 1998 and 2011, respectively (Fig. 1). Here, sedimentcores were collected with Plexiglas tubes (20 cm long and 7 cm indiameter) that were installed into a surface-operated coring deviceequipped with a core cylinder and a one-way check valve to pre-serve the integrity of sediment (Yang et al., 2017b). After collection,a cutting ring was pressed into the sediment for the determinationof bulk density and water content (Zhang et al., 2015). Then, sedi-ment cores were placed in sterile bags, stored in a temperature-controlled box at 4 �C, and transported back to the laboratorywithin 4 h. In the laboratory, each core sediment was immediatelyhomogenized and divided into three parts: the first part was frozenat �80 �C for DNA extraction, the second part was stored at 4 �C formeasurement of dissimilatory NO3

� reduction rates via slurry ex-periments, and the third part was freeze-dried for analysis ofphysico-chemical properties.

2.2. Analysis of sediment properties

Sediment water content and bulk density were determined byoven drying method and a cutting-ring method, respectively(Zhang et al., 2015). The pH and electrical conductivity (EC) weremeasured by a pHmeter (IQ150, IQ Scientific Instruments, USA) andan EC meter (2265FS, Spectrum Technologies Inc., USA), respec-tively. Sediment total organic carbon (TOC) was analyzed by anelemental analyzer (Vario EL, Elementar, Germany) after acidifyingthe samples with excess 1M HCl (Zhang et al., 2015). Sedimentinorganic N (NH4

þ, NO3� and NO2

�) were extracted by 2 M KCl, andtheir concentrations were determined by flow injection analysis(Skalar Analytical SANþþ, Netherlands) (Yin et al., 2017). Theconcentrations of total extractable Fe, ferrous iron (Fe(II)), andferric iron (Fe(III)) in sediments were determined by the ferrozine

method (Lovley and Phillips, 1987). Sulfide concentrations in thesediment were measured by the spectrophotometric method usingmethylene blue (Cline, 1969).

2.3. Determination of sediment potential DNF, ANAMMOX andDNRA rates

Sediment slurry experiments were conducted to determine thepotential rates of DNF and ANAMMOX based on the nitrogenisotope-tracing technique (Thamdrup and Dalsgaard, 2002; Denget al., 2015). In brief, sediment slurries were prepared by mixingfresh sediments and in situwater with a ratio of 1:7 (w:v), and thenpurged with helium for 30min and transferred into helium-flushed12mL vials (Exetainer, Labco). Subsequently, these vials were pre-incubated for 36 h to eliminate background residual NO3

�, NO2�

and oxygen at near field temperature (32 �C for July and 20 �C forNovember). After pre-incubation, these vials were divided intothree groups, which were spiked with 0.1mL sterile anoxic solu-tions of 15N (i: 15NH4

þ (15N at 99.6%), ii: 15NH4þ þ 14NO3

� (15N at99.6%), iii: 15NO3

� (15N at 99.6%)) through the septum, with the finalconcentration of 15N in each vial being approximately 100 mM (Houet al., 2013). Then, 0.2mL of ZnCl2 solution (50%) was added to halfof the vials to stop the reactions and serve as initial samples. Theremaining half of the vials were further incubated for 8 h, and theninjected with 0.2mL of 50% ZnCl2 solution at the end of incubationto terminate the reaction. Potential DNF were determined by 15Ntracer techniques based on the assumption of N2 as the only endproduct because the ratio of N2O to N2 from DNF in aquatic eco-systems is very small (Dong et al., 2002). Both 29N2 and 30N2 con-centrations in the incubation vials were determined by membraneinlet mass spectrometry (MIMS), and the potential rates of DNF andANAMMOX were calculated by the difference in produced 29N2 and30N2 between the final and initial samples as described by Hou et al.(2012) and Deng et al. (2015). The contributions of DNF andANAMMOX to total 29N2 production were quantified as follows:

P29 ¼D29 þ A29 (1)

where P29 (nmol N g�1 h�1) represents the total production rates of29N2; D29 and A29 (nmol N g�1 h�1) are 29 production rates fromDNF and ANAMMOX, respectively. Here, the ratio of 14N to 15N in N2produced during the incubation follows random isotope pairing(Risgaard-Petersen et al., 2003), and D29 was expressed as follows:

D29 ¼ P30 � 2� ð1� FnÞ � F�1n (2)

where P30 (nmol N g�1 h�1) represents total 30N2 production rates;Fn (%) is the fraction of 15N in NO3

�, which was estimated by theadded 15NO3

� and the residual NO3� concentrations in incubation

slurries. The potential DNF and ANA rates were calculated asfollows:

Dt ¼D29 þ 2� D30 (3)

A29 ¼ P29 � D29 (4)

where Dt and A29 (nmol N g�1 h�1) are the potential DNF and ANArates, respectively.

Potential DNRA rates were measured using the ammoniumoxidation membrane inlet mass spectrometry (OX/MIMS) method(Yin et al., 2014). Sediment slurries were prepared in the same wayas the DNF and ANAMMOX experiments. After pre-incubation, allthe vials were spiked with 0.1mL of 15NO3

� (15N at 99.6%, finalconcentration of approximately 100 mM 15N in each vial). Similarly,

D. Gao et al. / Environmental Pollution 255 (2019) 1132194

half of the vials were supplied with 0.2mL of ZnCl2 solution (50%)to serve as initial samples. The remaining half of the vials werefurther incubated for 8 h before adding 0.2mL of ZnCl2 solution(50%) to serve as final samples. Concentrations of 15NH4

þ producedduring the incubation were determined with OX/MIMS to estimatethe potential DNRA rates (Yin et al., 2014). Potential DNRA rateswere quantified as follows:

RDNRA ¼h�

15NHþ4

�Final

��15NHþ

4

�Initial

i�V �W�1 � T�1 (5)

where RDNRA (nmol N g�1 h�1) denotes total potential rates of

DNRA; ð15NHþ4 ÞFinal and ð15NHþ

4 ÞInitial (nmol N L�1) are the con-centrations of 15NH4

þ in the final and initial sediment slurry,respectively; V (L), W (g) and T (h) represent the incubation vialvolume, sediment dry weight and incubation time, respectively.

2.4. Sediment DNA extraction and quantitative PCR

Sediment total DNA was extracted using a Powersoil™ DNAIsolation Kit (MOBIO, USA) based on the manufacturer's protocol.The q-PCR analysis of extracted DNA was performed to determinethe nirS, ANAMMOX 16S, and nrfA gene abundance with an ABI7500 Sequence Detection System (Applied Biosystems, Canada)according to the SYBR green qPCRmethod. The primers used for thenirS, ANAMMOX 16S, and nrfA genes were cd3aF/R3cd, Amx-808-F/Amx-1040-R, and nrfA-2F/nrfA-2R, respectively. The detailed in-formation of the primers and q-PCR conditions for these genes isgiven in Table S1. The standard curves for the nirS, ANAMMOX 16Sand nrfA genes were constructed using a 10-fold dilution series ofthe standard plasmid DNA (Zheng et al., 2016). The amplificationefficiencies of nirS, ANAMMOX 16S and nrfA genes were 96.8%,97.0% and 93.3%, respectively. The gene abundances were calcu-lated using constructed standard curves and expressed in copiesper gram of dry sediment.

2.5. Statistical analysis

In this study, all data sets were checked for normality and log-transformed if necessary prior to statistical analysis. One-wayANOVA followed by Tukey LSD's test was performed to test forsignificant differences in sediment properties, NO3

� reduction pro-cesses and associated gene abundance among different types ofstands. Redundancy analysis (RDA) and Pearson's correlationanalysis were performed to explore the relationships betweenenvironmental factors, gene abundances, and NO3

� reduction rates.All the statistical analyses were performed using the SPSS 19.0

Table 1Sediment physicochemical properties (means± standard deviation; n¼ 6) in natural salt

July

Natural saltmarsh 5-year-old shrimpponds

18-year-old sh

Water content (%) 48.2± 2.7 b 49.9± 2.0 ab 53.2± 3.3 a

Bulk density (g cm�3) 1.5± 0.08 a 1.5± 0.06 a 1.3± 0.09 b

pH 6.7± 0.2 a 6.6± 0.1 a 6.6± 0.3 a

EC (mS cm�1) 2.0± 0.4 a 1.9± 0.2 a 1.1± 0.07 b

TOC (g kg�1) 16.9± 0.9 c 22.4± 2.0 b 28.6± 3.4 a

NH4þ (mg kg�1) 16.8± 0.8 b 59.5± 7.1 a 68.5± 8.4 a

NO3� (mg kg�1) 1.5± 0.1 c 2.1± 0.2 b 2.5± 0.2 a

NO2� (mg kg�1) 29± 4 c 44± 5 b 54± 4 a

Fe(II) (g kg�1) 8.3± 3.9 b 12.2± 1.9 ab 14.5± 1.4 a

Fe(III) (g kg�1) 12.0± 4.3 a 8.9± 1.6 ab 7.5± 1.8 b

Sulfide (mg kg�1) 19.6± 4.4 b 67.1± 24.2 a 76.2± 13.5 a

Different letters indicate significant differences (p< 0.05) at different stands in the same

software package and Canoco for Windows 4.5 software.

3. Results

3.1. Sediment properties

Shrimp aquaculture had significant effects on most of themeasured sediment properties in both July and November (Table 1).The mean sediment water content in the 18-year-old shrimp pondswas generally higher than that in the 5-year-old shrimp ponds andnatural saltmarsh (Table 1). Sediment bulk density and EC weresignificantly lower in the 18-year-old shrimp ponds (p< 0.05)(Table 1), but were similar between the 5-year-old shrimp pondsand natural saltmarsh. Sediment pH ranged from 6.4 to 6.7, with nosignificant difference among the shrimp ponds and natural salt-marsh (Table 1). Sediment TOC, NH4

þ, NO3� and NO2

� in the naturalsaltmarsh varied between 16.9 and 18.2 g C kg�1, 16.8e18.8mgNkg�1, 1.5e1.6mgN kg�1, and 20e29 mg N kg�1, respectively, whilethose in the shrimp ponds were significantly higher, especially inthe 18-year-old ones (p< 0.05) (Table 1). Sediment Fe(II) and Fe(III)were significantly higher and lower, respectively, in the shrimpculture ponds as compared to the natural saltmarsh (p< 0.05)(Table 1). The concentrations of sulfide in the shrimp ponds rangedbetween 67.1 and 109.2mg S kg�1, which were significantly higherthan those in the natural saltmarsh (19.6e22.1mg S kg�1) (p< 0.05)(Table 1).

3.2. Sediment nirS, ANAMMOX 16S, and nrfA gene abundance

The abundance of nirS gene at our study sites ranged from1.45� 106 to 1.57� 106 copies g�1 in July and from 1.38� 106 to1.47� 106 copies g�1 in November. Significantly higher nirS geneabundance was found in the 18-year-old shrimp ponds (p< 0.05),while no significant difference was observed between the naturalsaltmarsh and 5-year-old shrimp ponds (Fig. 2). In addition, shrimpaquaculture had no significant effects on the ANAMMOX 16S geneabundance (5.31� 105 to 6.73� 105 copies g�1 in July, 4.56� 105 to5.11� 105 copies g�1 in November) (Fig. 2). Sediment nrfA geneabundance in the natural saltmarsh varied between 6.74� 105 and8.95� 105 copies g�1, which was generally lower than that in the 5-year-old shrimp ponds (1.11� 106 to 1.18� 106 copies g�1) and 18-year-old shrimp ponds (1.21� 106 to 1.24� 106 copies g�1 copiesg�1) (p< 0.05) (Fig. 2). In general, no significant seasonal variationsof nirS, ANAMMOX 16S and nrfA gene abundances were observed inboth the natural saltmarsh and the shrimp ponds (except for nrfAgene in the natural saltmarsh) (Fig. 2).

marsh and shrimp culture ponds.

November

rimp ponds Natural saltmarsh 5-year-old shrimpponds

18-year-old shrimpponds

48.9± 5.5 b 52.8± 6.9 b 60.6± 1.2 a

1.4± 0.1 a 1.4± 0.2 a 1.2± 0.03 b

6.4± 0.2 a 6.5± 0.3 a 6.6± 0.3 a

3.0± 0.5 a 2.2± 0.3 b 1.8± 0.4 b

18.2± 1.3 b 28.5± 2.9 a 30.4± 2.9 a

18.8± 2.4 b 76.2± 7.8 a 88.9± 7.7 a

1.6± 0.1 c 2.2± 0.2 b 2.6± 0.5 a

20± 3 c 54± 5 b 88± 11 a

6.9± 2.1 b 18.7± 4.8 a 15.1± 2.9 a

11.9± 3.1 a 7.9± 1.7 ab 7.8± 3.2 b

22.1± 3.2 b 94.1± 17.9 a 109.2± 17.8 a

season.

Fig. 2. Sediment nirS, ANAMMOX 16S, and nrfA gene abundance in the natural saltmarsh and shrimp culture ponds. Different letters indicate significant differences (p< 0.05)among different stands in the same season, and the asterisks denote significant differences (p < 0.05) in different seasons at the same stand. Error bars represent the standarddeviation (n¼ 6).

D. Gao et al. / Environmental Pollution 255 (2019) 113219 5

3.3. Sediment dissimilatory NO3� reduction processes

Shrimp aquaculture dramatically altered sediment NO3� reduc-

tion processes in both July and November (Fig. 3). Sediment po-tential DNF rates ranged from 8.9 to 14.3 nmol N g�1 h�1 in July andfrom 2.5 to 5.1 nmol N g�1 h�1 in November, while the potentialANAMMOX rates varied from 1.1 to 2.1 nmol N g�1 h�1 in July andfrom 0.6 to 1.4 nmol N g�1 h�1 in November. The process rates ofDNF and ANAMMOX in the shrimp culture ponds were generallyhigher than those in the natural saltmarsh (p< 0.05), but no sig-nificant difference was observed between the 5- and 18-year-oldshrimp culture ponds (Fig. 3). In addition, sediment potential DNRArates were significantly higher in the shrimp culture ponds(2.9± 0.9 nmol N g�1 h�1 in July and 2.0± 0.6 nmol N g�1 h�1 inNovember) than in the natural saltmarsh (1.2± 0.3 nmol N g�1 h�1

in July and 0.7± 0.4 nmol N g�1 h�1 in November) (p< 0.05), andshowed an increasing trend with the age of ponds (Fig. 3).Seasonally, the mean rates of DNF, ANAMMOX and DNRA weregenerally higher in July than in November (p< 0.05) (Fig. 3). Theprocesses of DNF, ANAMMOX and DNRA contributed to 57.2e82.2%,

Fig. 3. Sediment potential DNF, ANAMMOX and DNRA rates in the natural saltmarsh and sdifferent stands in the same season, and the asterisks denote significant differences (p < 0.0(n¼ 6).

10.3e17.3% and 7.5e27.4%, respectively, of the total NO3� reduction

(Fig. 4). The relative proportions of DNF and ANAMMOX to totalNO3

� reduction were generally lower in the shrimp culture pondsthan in the natural saltmarsh. In contrast, the contribution of DNRAto total NO3

� reduction was significantly higher in the shrimp cul-ture ponds than in the natural saltmarsh (p< 0.05) (Fig. 4).

3.4. Influences of sediment properties on NO3� reduction rates

RDA results showed that the sum of RDA1 and RDA2 accountedfor ca. 89.1% and 92.4% of the variations in NO3

� reduction processesin July and November, respectively. Sediment samples in the nat-ural saltmarsh were clearly distinct from those in the shrimp cul-ture ponds along RDA1 and RDA2 (Fig. 5). Sediment potential DNFrate was significantly associated with bulk density, TOC, NH4

þ, NO3�,

NO2� and sulfide in July (p< 0.05), and significantly related to EC,

TOC, NH4þ, NO3

�, NO2�, Fe(II), Fe(III), sulfide and nrfA gene in

November (p< 0.05) (Fig. 5; Table S2). The ANAMMOX rate wassignificantly correlated with EC, TOC, NH4

þ, NO3�, NO2

� and sulfide inboth July and November (p< 0.05) (Fig. 5; Table S2). Sediment

hrimp culture ponds. Different letters indicate significant differences (p < 0.05) among5) in different seasons at the same stand. Error bars represent the standard deviation

Fig. 4. Relative contributions of DNF, ANAMMOX and DNRA to total NO3� reduction in the natural saltmarsh and shrimp culture ponds. Different lowercase letters indicate sig-

nificant differences (p< 0.05) among different stands in the same season. Error bars represent the standard deviation (n¼ 6).

Fig. 5. Ordination diagram showing the results of RDA of NO3� reduction processes, associated gene abundances, and soil physicochemical characteristics. The hollow circles

represent individual sediment samples from the three stands in the Min River Estuary (1e6: natural saltmarsh; 7e12: 5-year-old shrimp ponds; 13e18: 18-year-old shrimp ponds).

D. Gao et al. / Environmental Pollution 255 (2019) 1132196

potential DNRA rate was positively correlated with water content,TOC, NH4

þ, NO3�, NO2

�, Fe(II), sulfide and nrfA gene, but negativelycorrelated with bulk density, EC, and Fe(III) (p< 0.05) (Fig. 5;Table S2).

4. Discussion

4.1. Effects of shrimp aquaculture on NO3� reduction processes

The conversion of coastal saltmarshes to aquaculture ponds canexert significant impacts on sediment physico-chemical properties,in particular the availability of C and N substrates, resulting in aseries of changes in the biogeochemical processes of nutrients(Murphy et al., 2016; Yang et al., 2017a). Our results showed that

sediment potential DNF, ANAMMOX and DNRA rates were gener-ally higher in the shrimp culture ponds than in the natural salt-marsh (Fig. 4), which could be attributed to the changes insediment properties. DNF and DNRA processes in the aquatic eco-systems are generally governed by the availability of substrates andelectron donors, with the rates of both processes being positivelycorrelated with sediment NO3

�, NO2� and TOC (Plummer et al., 2015;

Shan et al., 2016). Although ANAMMOX activity does not require adirect energy source, it is found to be highly related to the amountof C and N substrates in sediments (Trimmer et al., 2003; Hou et al.,2015). This relationship might be because organic matter miner-alization can provide NH4

þ substrate for the ANAMMOX process(Damashek and Francis, 2018). In the present study, shrimp aqua-culture significantly increased sediment TOC, NH4

þ, NO3� and NO2

�

D. Gao et al. / Environmental Pollution 255 (2019) 113219 7

concentrations (Table 1), as a result of the decomposition of a largenumber of residual feeds and excrements (Wu et al., 2014).Increasing sediment C and N substrates in the shrimp culture pondscould increase the rates of dissimilatory NO3

� reduction, as sup-ported by the positive correlations observed between the rates ofpotential DNF, ANAMMOX and DNRA, and the concentrations ofsediment TOC, NH4

þ, NO3� and NO2

� (Fig. 5; Table S2).Previous studies have shown that NO3

� reduction processes aremediated essentially by microbial activities, which imply that theabundance of nirS, ANAMMOX 16S, and nrfA genes can be used as aproxy in characterizing the rates of DNF, ANAMMOX and DNRA,respectively (Hou et al., 2015; Smith et al., 2015; Zheng et al., 2016).However, in this study, the abundance of nirS and ANAMMOX 16Sgenes did not exhibit a clear increasing trend from the naturalsaltmarsh to the shrimp culture ponds (Fig. 2). A positive rela-tionship was only observed between nrfA gene abundance and thepotential DNRA rate (Fig. 5; Table S2). Our results of correlationanalysis showed that the concentrations of various substrates (TOC,NH4

þ, NO3� and NO2

�) in sediments were more significantly corre-lated with DNRA than nrfA gene abundance (Fig. 5; Table S2),suggesting that bacteria gene abundance might be not the mostdominant factor in controlling the process rates following theconversion of natural saltmarshes to shrimp culture ponds. Itshould be noted that nrfA gene was not found in all bacteria per-forming DNRA (Stremi�nska et al., 2012). Additionally, gene abun-dance at the DNA level cannot fully reflect microbial activities (Shanet al., 2016), so further experiments should be conducted todetermine the role of microbes in sediment NO3

� reduction in futurestudies.

As for the shrimp culture chronosequence, we found that theconcentrations of sediment C and N substrates were considerablyhigher in the 18-year-old shrimp ponds than in the 5-year-old ones(Table 1), which were expected to cause a higher rate of sedimentdissimilatory NO3

� reduction in the former. However, the rates ofDNF and ANAMMOX processes did not differ between the older andyounger ponds (Fig. 3). Themore pronounced increase in the rate ofDNRA than that of DNF and ANAMMOX in the older shrimp cultureponds could be related to the differences in sediment properties.Firstly, a higher TOC/NO3

� ratio can promote DNRA but suppressDNF while both processes compete for electron acceptors (i.e., NO3

�

and NO2�), because DNRA has a higher efficiency in utilizing elec-

tron acceptors in an environment with high C and low NO3� con-

centrations (Kraft et al., 2014; Shan et al., 2016). In our study, theTOC/NO3

� ratios were generally higher in the shrimp culture ponds(1.2� 104 in 5-year shrimp ponds and 1.3� 104 in 18-year shrimpponds), particularly in the older ones, than those in the naturalsaltmarsh (9.9� 103). Also, a high sulfide concentration is known tostimulate DNRA and inhibit the final two steps of the DNF process(i.e. NO to N2O, and N2O to N2) (Brunet and Garcia-Gil, 1996; Sengaet al., 2006). In the shrimp culture ponds, the decomposition ofresidual feeds can consume a large amount of oxygen and form ahighly reducing environment in the sediment, which is beneficial tothe production of sulfide (Yang and Silver, 2016). Moreover, Fe(II)can usually be used and oxidized by some chemolithoautotrophicmicrobes to carry out DNRA through the reduction of NO3

� or NO2�

(Giblin et al., 2013; Robertson et al., 2016). In the present study, wefound a positive correlation between Fe(II) concentrations andpotential DNRA rates, supporting the important role of iron oxidesin this N transformation process (Fig. 5; Table S2). Previous re-searchers have also found that the ANAMMOX activity can beenhanced under optimal concentrations of TOC, NH4

þ, NO2�, NO3

�

and sulfide, but suppressed when these substrates are in excess(Jensen et al., 2008; Bettazzi et al., 2010; Fern�andez et al., 2012). Thecombination of high concentrations of sulfide, Fe(II), C and N

substrates, as well as a high ratio of TOC/NO3� in the ponds with a

longer history of shrimp aquaculture played a crucial role in pro-moting the DNRA process. Plummer et al. (2015) have similarlyreported maximum DNF and ANAMMOX activities at intermediateconcentrations of TOC and sulfide, although their explanationsvaried for the optima observed in different environments (Eyre andFerguson, 2009). Given the above results, a mechanistic under-standing of the impact of a wide range of sediment C, N and sulfideconcentrations on NO3

� reduction processes in the shrimp cultureponds deserves further investigation.

Seasonally, NO3� reduction rates in the shrimp culture ponds

were generally higher in July than in November (Fig. 3), in spite ofthe lower TOC, NH4

þ, NO3� and NO2

� concentrations (Table 1). As theshrimp rearing period in our study area occurred between June andNovember, sediment nutrient contents were significantly higher inNovember than in July owing to the accumulation of residual feeds(Yang et al., 2017b). However, the seasonal variations of NO3

�

reduction rates did not follow the temporal pattern of substratesupply (Fig. 3), but were more strongly related to the changes intemperature (Smyth et al., 2013; Canion et al., 2014). In the presentstudy, the measured in situ temperatures in July and Novemberwere 31.2e33.9 �C and 19.6e21.1 �C, respectively. A higher tem-perature in July would likely promote organic matter decomposi-tion and oxygen consumption, leading to a more favorable anoxiccondition for dissimilatory NO3

� reduction processes (Hou et al.,2015). On the other hand, temperature may regulate NO3

� reduc-tion rates through its effect on the metabolism of related micro-organisms (Yin et al., 2017). Yet, we found no significant seasonalvariations of nirS, ANAMMOX 16S, and nrfA gene abundances in thisstudy (Fig. 2), further implying that gene abundance was not a keyfactor controlling NO3

� reduction processes. Overall, temperatureand the concentrations of sulfide, Fe(II), and other substrates insediments were the dominant factors regulating the DNF, ANAM-MOX and DNRA processes in response to shrimp aquaculture.

4.2. Environmental implications of NO3� reduction in shrimp ponds

Increasing levels of reactive N can lead to various environmentalproblems (Galloway et al., 2008). Since nearly 20e30% of the totalanthropogenic reactive N, mainly in the form of NO3

�, is eventuallytransported to the estuarine and coastal zones (Seitzinger et al.,2010; Chen et al., 2016a, b), the transformation of NO3

� is particu-larly important for the N removal and retention in these ecosystems(Dong et al., 2011). The reclamation of natural saltmarsh for aqua-culture activities is one of the main anthropogenic disturbancesthat could threaten the intrinsic N balance of coastal wetlands(Murphy et al., 2016; Yang et al., 2017a). In the present study, DNFwas themost dominant NO3

� transformation pathway (57.2e82.2%),while the contribution of ANAMMOX (10.3e17.3%) and DNRA(7.5e27.4%) to NO3

� reduction was similar to the range of valuesreported for other estuarine ecosystems (Giblin et al., 2013; Donget al., 2009; Plummer et al., 2015; Zheng et al., 2016). Amongthese three metabolisms, DNF and ANAMMOX play an importantfunction in ameliorating reactive N loads and subsequently eutro-phication (Rose et al., 2014), while DNRA plays a different role byreducing NO3

� into more bioavailable NH4þ, which is subsequently

retained in the ecosystems (Wankel et al., 2017). Although DNRA isalso a potential mechanism of nitrous oxide production, thedetermination of this process in natural ecosystems is very limited(Murray et al., 2018). Therefore, DNRA process is not as an expla-nation for N loss in our study. We found that shrimp aquaculture,especially after a long period of operation, significantly increasedthe contribution of DNRA to total NO3

� reduction, while at the sametime reduced the relative contributions of DNF and ANAMMOX(Fig. 4). Our findings showed that shrimp aquaculture favored the

D. Gao et al. / Environmental Pollution 255 (2019) 1132198

dominance of DNRA in the NO3� reduction process, in consistence

with the results of Murphy et al. (2016) that DNRAwas significantlyenhanced after clam cultivation.

Previous studies have used the N Retention Index (NIRI, calcu-lated as DNRA/(DNFþANAMMOX)) to assess the relative amount ofreactive N being retained in versus removed from the ecosystem(Algar and Vallino, 2014; Plummer et al., 2015). In this study, theNIRI in the shrimp culture ponds was generally higher than that inthe natural marshland (Fig. 6), which suggested that shrimpaquaculture played a significant role in N retention. Plummer et al.(2015) found that increasing sulfide and organic carbon availabilitycould enhance NIRI, which was also supported by the positivecorrelations observed between TOC and sulfide concentrations andNIRI in our study (Fig. S1). Based on the average bulk density ofsurface sediments in the natural saltmarsh and shrimp cultureponds (Table 1), we estimated that the amounts of NH4

þ convertedfrom NO3

� in the natural marshland, 5-year-old and 18-year-oldshrimp ponds were approximately 7.2, 16.7, and 28.7 g N m�2 yr.�1,respectively. In addition, N loss in shrimp culture ponds (116.6 g Nm�2 yr.�1 in 5-year shrimp ponds and 130.6 g N m�2 yr.�1 18-yearshrimp ponds) was also higher than that of natural saltmarsh(88.9 g N m�2 yr.�1). However, the reactive N retention in theshrimp culture ponds increased significantly by 235% as comparedto the natural saltmarsh, but N loss increased only by 39%. Afternatural saltmarshes convert to shrimp culture ponds, DNRA com-petes with DNF and may limit the loss of reactive N (Murphy et al.,2016). These results implied that if natural saltmarshes are beingincreasingly converted to shrimp culture ponds, there might be agradual increase in the retention of reactive N (Murphy et al., 2016).This would result in greater sediment N loads and fluxes, andsubsequently a deterioration of water quality and the outbreak ofshrimp diseases (Wu et al., 2014; Castillo-Soriano et al., 2017). Yanget al. (2017b) observed that the sediment NH4

þ release rates in theshrimp culture ponds of Min River Estuary were generally high ascompared to the fluxes of other nutrients (i.e. NO3

�, NO2� and PO4

3�),which might be in part associated with the greater DNRA activityand higher NIRI. However, some of other microbial N trans-formation, e.g., N mineralization, NH4

þ immobilization and

Fig. 6. Sediment Nitrogen Retention Index (NIRI) in natural saltmarsh and shrimpculture ponds. Different letters indicate significant differences (p< 0.05) amongdifferent stands in the same season, and the asterisks denote significant differences(p< 0.05) in different seasons at the same stand. Error bars represent the standarddeviation (n¼ 6).

nitrification, also mediate the NH4þ availability, which should be

studied further (Lin et al., 2017). At the end of each culture cycle,large quantities of wastewater with high NH4

þ levels are dischargedfrom the aquaculture ponds into the adjacent coastal waters, whichcould directly increase the risk of eutrophication (Wu et al., 2014).On the other hand, increasing reclamation along the coastline hascaused serious degradation of natural saltmarshes, indirectlyweakening the function of reactive N removal as well as the pro-vision of ecosystem services by coastal wetlands (Sun et al., 2015).In this study, we have examined the impact of shrimp aquacultureon sediment dissimilatory NO3

� reduction processes only, while werecognize that there are many other biogeochemical processes thatcould be involved in N transformation and transportation in theecosystems. Our findings contribute to a better mechanistic un-derstanding of reactive N conversion in the shrimp culture ponds,and highlight the importance of taking into account the DNRAprocess in ecological impact assessments of wetland reclamationprojects.

5. Conclusion

Our study provides an early insight on the changes in dissimi-latory NO3

� reduction processes caused by the conversion of naturalsaltmarshes to shrimp culture ponds. Shrimp culture generallyincreased the potential DNF, ANAMMOX and DNRA rates ascompared to the natural saltmarsh. With the pond age, the po-tential NO3

� reduction rates generally increased, with the magni-tude of increase being greater for DNRA than DNF and ANAMMOX.Moreover, shrimp culture significantly increased the contributionof DNRA to total NO3

� reduction, while simultaneously reducing theimportance of DNF and ANAMMOX. Sediment temperature and theavailability of substrates were found to be the key factors governingthe variations of NO3

� reduction rates. Overall, our results showedthat DNRA was an important NO3

� reduction process retainingreactive N in the coastal shrimp ponds, whichmight limit the loss ofreactive N and further aggravate various eco-environmental prob-lems associated with N enrichment.

Conflicts of interest

The authors declare that they have no conflict of interest.

Acknowledgments

This work was supported by the Natural Science Foundation ofChina (grant numbers: 41725002, 41671463, 41761144062,41601530, and 41730646). It was also funded by Chinese NationalKey Programs for Fundamental Research and Development (No.2016YFA0600904 and 2016YFE0133700) and FundamentalResearch Funds for the Central Universities. Thanks are given toeditor and anonymous reviewers for constructive comments onthis manuscript.

Appendix A. Supplementary data

Supplementary data to this article can be found online athttps://doi.org/10.1016/j.envpol.2019.113219.

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