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Nitrogen removal in an ecological ditch receiving agricultural drainage in subtropical central China
Ecological Engineering 82 (2015) 487–492
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Ecological Engineering journal homepage: www.elsevier.com/locate/ecoleng
Nitrogen removal in an ecological ditch receiving agricultural drainage in subtropical central China Liang Chena,b , Feng Liua,b,* , Yi Wanga,b , Xi Lia,b , Shunan Zhanga,b,c , Yong Lia,b , Jinshui Wua,b,* a Changsha Research Station for Agricultural & Environmental Monitoring, Institute of Subtropical Agriculture, Chinese Academy of Science, Hunan 410125, PR China b Key Laboratory of Agro-ecological Process in Subtropical Region, Institute of Subtropical Agriculture, Chinese Academy of Science, Hunan 410125, PR China c Graduate University of the Chinese Academy of Sciences, Beijing 100039, PR China
A R T I C L E I N F O
A B S T R A C T
Article history: Received 22 November 2014 Received in revised form 11 April 2015 Accepted 23 May 2015 Available online xxx
Nitrogen (N) loss from agricultural field can cause eutrophication in downstream freshwater systems. Ecological ditches (eco-ditches) which are engineered to mitigate N loss from agricultural runoff. This study presents an analysis of water quality data for a 200 m long eco-ditch treating agricultural drainage in subtropical central China. Inflow concentrations of total N (TN) ranged from 2.3 to 3.1 mg l1 and typically contained 1.1–1.7 mg l1 nitrate (NO3-N) and <0.4 mg l1 ammonium (NH4+-N). Mean concentration removal efficiencies (h) in the eco-ditch for TN, NO3-N, and NH4+-N were 75.8, 63.7, and 77.9%, respectively. Mean area-based first-order removal rate constant (J) and removal rate constant (k) of TN were 942.1 mg-N m2 d1 and 1.11 m d1, respectively. Outflow from the eco-ditch typically contained <0.85 mg l1 TN and <0.70 mg l1 NO3-N. Rate of N loss in the eco-ditch was highly seasonal, generally peaking in the summer months (May–August). Results show that the eco-ditch can be effective at reducing transport of non-point source of N in-situ, particularly in warm environments. Its use should be expanded to similar areas, although managing the eco-ditch to maximize N removal in the long-term will require dynamic management. ã2015 Published by Elsevier B.V.
1. Introduction Agricultural intensification is recognized as a major source of increased nitrogen (N) in aquatic systems (McIsaac and Libra, 2003). Chemical N fertilizers are widely used for pursuing high crop yield. However, fertilizer additions most often exceed the N requirements of crops, thus resulting in a higher amount of ammonium (NH4+) and nitrate (NO3) leached from soils during storm events and flushed away in agricultural runoff (Tilman, 1999). Runoff from agricultural lands can potentially carry high nutrient loads to receiving waters and contribute to eutrophication, hypoxia, and ecological damage (Howarth et al., 2011). It has been reported that total N (TN) levels greater than 0.5 mg l1 can result in large masses of nuisance algae unless other factors limit algae growth (Biggs, 2000), and nitrate concentrations above 2.0 mg l1 can cause toxicity in a variety of
* Corresponding authors at: Changsha Research Station for Agricultural & Environmental Monitoring, Institute of Subtropical Agriculture, Chinese Academy of Science, Hunan 410125, PR China. Tel.: +86 73184615224; fax: +86 73184619736. E-mail addresses: [email protected] (F. Liu), [email protected] (J. Wu). http://dx.doi.org/10.1016/j.ecoleng.2015.05.012 0925-8574/ ã 2015 Published by Elsevier B.V.
freshwater organisms (Camargo et al., 2005). Therefore, developing best management practices (BMPs) that minimize nutrient runoff from agricultural fields is needed. Vegetated ecological ditches (eco-ditches) are a BMP being closely examined for nutrient mitigation (Kröger et al., 2007; Moore et al., 2010; Liu et al., 2013). Ditches are often designed and constructed as the main component of an agricultural irrigated system, which aims to irrigate or drain excess water from the fields. They also act as major pathways of surface and subsurface flow N from agricultural lands to receiving waters (Kröger et al., 2007; Herzon and Helenius, 2008). Eco-ditches are transformed from conventional agricultural drainage ditches, which consist of drainage channels, substrate, vegetation, and flow control facilities (Wu et al., 2013). They are well placed to mitigate contaminants in a manner similar to a surface-flow constructed wetland, filtering pollutants from runoff before exiting into downstream receiving systems. Studies have demonstrated vegetated eco-ditches’ capacity to mitigate pesticides (Bennett et al., 2005; Cooper et al., 2004; Moore et al., 2011). More recently, nutrient mitigation potential of eco-ditches has been investigated. Kröger et al. (2007) determined the inorganic N reduction capacity of agricultural
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ditches under natural, variable rainfall conditions in northern Mississippi, and results indicated that drainage ditches reduced the initial 2.5% inorganic N load to less than 1.5%, which is a 57% reduction over 2-years study period. Wu et al. (2011) reported a mean reduction of 52%, 53% and 58% for TN, NO3-N, and NH4+-N, respectively, for vegetated eco-ditches where the bottom part was packed with adsorbent. The main mechanisms of eco-ditches for N-removal are sediment retention, plant uptake, and microbial metabolic activities, and vegetation inside ditches are important to effectively slow down the water velocity, increase agrochemical retention, and subsequently provide better conditions for pollution dissipation in ditches (Hu et al., 2010; Kröger et al., 2007). Studies have demonstrated the potential for eco-ditches to decrease N loads exiting agricultural systems, but the use of degradation models to describe N mitigation performance is limited. Area-based first-order kinetic models are usually applied to describe of N removal in surface-flow wetland systems (Beutel et al., 2009; Karpuzcu and Stringfellow, 2012). Incorporation of monitoring data into established models for eco-ditch systems is needed to support dynamic management decisions related to changing performance of ecologically engineered ditches. Consequently, the aim of this study was to document the N mitigation performance of an eco-ditch based on a detailed evaluation of water quality for a year. The vegetated eco-ditch, located in the Tuojia catchment in central China, is mainly receives drainage water from paddy fields. Seasonal N concentration in conjunction with temporal hydrological variability provided the potential to generate N loads flowing within the eco-ditch. Efficiency and kinetics of N (TN, NH4+-N and NO3-N) removal were determined and analyzed.
2. Materials and methods 2.1. Study site The Tuojia catchment is an upstream subcatchment of the Jinjing River (28 300 –28 390 N, 113180 –113 260 E, elevation of 46– 452 m), located in Changsha County, Hunan Province, China. The site has a subtropical climate, with an annual mean air temperature of 17.5 C, a mean annual rainfall of 1330 mm, and an annual potential evaporation of 1300 mm. An agricultural drainage ditch (approximately 850 m in length) was the main channel of water from paddy fields to the Tuojia catchment at the study site. The ditch receives a contributing area of 28.6 ha, with forest and rice agriculture as the main land-use types in the catchments accounting for approximately 52.8% and 40.9% of the total land area, respectively (Fig. 1). Generally, no fertilizer is applied in the forest. Rice (Oryza sativa) is transplanted twice a year (mid-April and mid-July) and harvested at the end of June and in the middle of October. During rice planting, there are typically two fertilizer applications in each rice season, and paddy fields receive approximately 2146 kg N ha1 yr1 and 507 kg P ha1 yr1 annually. Intensive fertilizer applications usually result in high nutrient concentrations in the surface water of paddy fields. The eco-ditch was designed and constructed in the downstream section of the original drainage ditch in 2009 to reduce nutrient losses from paddy fields (Liu et al., 2013). The eco-ditch was 200 m in length, with a mean top width of 4.0 m, mean bottom width of 2.8 and an average depth of 1.0 m. The eco-ditch includes a series of weirs used as water control structures and various hydrophytes. Dominant vegetation at the eco-ditch includes Canna indica,
Fig. 1. Study site location in subtropical central China. Map of the eco-ditch (highlighted in frame), which is located to the west of the Tuojia River.
L. Chen et al. / Ecological Engineering 82 (2015) 487–492
Hydrocotyle vulgaris, Sparganium stoloniferum, Myriophyllum sp., and Juncus sp. Water discharge was monitored at the outlet of ecoditch using the observed flow velocity and the flow cross-section area method which described by our previous study (Wang et al., 2014). Site maintenance was undertaken when the eco-ditch began running.
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of N is not known, and it is reasonable to assume that P = N = 4, which is a mean N-value proposed for free-water surface wetlands (Kadlec and Wallace, 2008; Kadlec and Wallace, 2008). 3. Results and discussion 3.1. Nitrogen removal efficiency
2.2. Field sampling and lab analysis Water quality characterization of the eco-ditch is presented in Fig. 2. As can be seen from this figure, inflow concentration of TN, 0.4
(a)
0.3 -1
CNH +-N (mg L ) 4
During the 2013 study period (January–December), weekly water samples were collected at the inlet and outlet of the ecoditch. Water samples were collected in glass 500 ml bottles to guarantee the volume requirements in laboratory analysis. All bottles were rinsed with sample water prior to collection of a depth-integrated sample. Samples were immediately stored at 4 C after sampling and transported to the laboratory on the day of sampling. Samples were received by the laboratory, logged in and inspected for damage, and stored at 4 C until filtering and analysis. All filtration and preservation of samples were completed within 24 h. TN, NO3-N and NH4+-N concentrations were measured using an automatic flow injection analyzer (Fia-star 5000, Foss Tecator, Sweden) according to the standard methods for Water and Wastewater Monitoring and Analysis (SEPA, 2002). Briefly, water samples were filtered through a 0.45-mm membrane and the filtrates were used to directly determine the NO3-N and NH4+-N concentrations using the fully automated injection system. TN concentrations were determined using the fully automated injection system after digestion with a K2S2O8-NaOH solution.
0.2
Inlet Outlet
0.1
0.0 Jan Feb M ar Apr M ay Jun Jul Aug Sep Oct Nov Dec Jan 2.0
(b)
2.3. Kinetic and removal calculations
h¼
C in C out 100 C in
1.5 -1
CNO --N (mg L ) 3
Rates of removal for TN, NO3-N and NH4+-N were quantified using three common approaches for wetland systems (Kadlec and Knight, 1996). First, concentration removal efficiency (h, %) was calculated according to Eq. (1), where Cin and Cout are the inlet and outlet concentrations of TN, NO3-N or NH4+-N (mgl1) respectively.
Q ðC in C out Þ A
Inlet Outlet
0.5
(1)
Second, the areal removal rate (J, mg-N/m2/d) was calculated according to Eq. (2), where Q is the flow rate (m3 d1), A is the surface area of the eco-ditch (m2). J¼
1.0
0.0 Jan Feb M ar Apr M ay Jun Jul Aug Sep Oct Nov Dec Jan 4.0
(c)
(2)
C out C 1 ¼ p C in C ð1 þ k=PqÞ
-1
Finally, the modified first-order areal removal rate constants, k (m d1) were determined using the P–k–C* model (Eq. (3)), which is a relaxed tanks-in-series flow (TIS) model (Kadlec and Wallace, 2008).
CTN (mg L )
3.0
2.0
Inlet Outlet
(3)
where C* is the background concentration (mg l1), P is the apparent number of tanks in series, and q is the hydraulic loading (m d1). C* accounts for substances generated in the wetland by biological activity, sediment release, etc., as well as the nondegradable fraction of the contaminant. From observing our N data, it seemed that a C* concentration of 0, 0 and 0.2 mg l1 for NO3-N, NH4+-N and TN, respectively, was appropriate, as outflow concentrations of N tend toward a low concentration. P is a free-fitting parameter with the condition P < N, where N is the tracer TIS number. In the absence of a tracer test, the actual number
1.0
0.0 Jan Feb M ar Apr M ay Jun Jul Aug Sep Oct Nov Dec Jan
Month Fig. 2. Concentrations of (a) ammonium (NH4+-N), (b) nitrate (NO3-N) and (c) total nitrogen (TN) in the inlet and outlet of the eco-ditch. Monitoring period was from January 2013 to December 2013. Solid circles represent monthly inlet concentrations and open circles represent outlet concentrations.
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NO3-N and NH4+-N were typically greater than outflow concentration. Total N levels entering the eco-ditch over the study period (January–December, 2013) varied from 2.32 mg l1 to 3.05 mg l1, and were dominated by nitrate (42.6–60.3%). Inlet concentration of NO3-N and NH4+-N varied from 1.12 mg l1 to 1.68 mg l1 and 0.24 mg l1 to 0.33 mg l1, respectively (Fig. 2). Outflow concentrations were always lower for TN, NO3-N and NH4+-N. TN levels in the eco-ditch outflow ranged from 0.41 mg l1 to 0.83 mg l1, and NO3-N and NH4+-N levels in eco-ditch outflow were consistently <0.7 mg l1 and <0.1 mg l1, respectively. TN, NO3N and NH4+-N in eco-ditch inflow increased sharply as the transplanting season (April–May and July–August) progressed (Fig. 2). This may due to fertilizer application were applied during this period. The eco-ditch investigated in this study was observed high N removal efficiency. Mean TN, NO3-N and NH4+-N removal efficiencies were 75.8, 63.7 and 77.9%, respectively (Table S1). Moreover, changes of the N removal efficiencies are shown in Fig. 3a. Removal efficiencies (h) of TN, NO3-N tended to increase during summer periods, but this trend is not fully suited for NH4+N (Fig. 3). This seasonal pattern of h suggests the biological N transformation rate increased during warm summer periods. This was to be expected, since biological process is strongly affected by temperature. The primary mechanism for the loss of NO3-N in wetlands is denitrification, the microbial reduction NO3-N of to N2, which typically accounts for 60–95% of the NO3-N removal in wetlands (Spieles and Mitsch, 1999). Denitrification rates increase dramatically with temperature, within a lower and upper bounds of around 5 C and 70 C, respectively (Vymazal, 2007). Sirivedhin and Gray (2006) documented a two-order of magnitude increase in denitrification rates in wetland sediments when temperature was increased from 4 C to 25 C. Furthermore, it is the growing season for the wetland plants at the warm periods, so plant uptake would be another important way of N-removal. Inflow and outflow NH4+N concentrations were typically low (<0.3 mg l1). Because of the positive charge on the ammonium ion, NH4+-N is easily absorbed by sediments or detritus and can be released easily when water
90
3.2. Areal N removal rate Historical observations at a range of wetlands show that levels of a number of pollutants, including NO3-N, generally decrease with distance thorough wetlands, suggesting that pollutant removal in wetlands is first order in nature (Kadlec and Knight, 1996). In this study, areal-based first-order removal rates for TN, NO3-N and NH4+-N were calculated from flow and water quality measurements (Eq. (2)). Averaged removal rates (J) were 942.1, 407.8, and 96.1 mg-N m2 d1 for TN, NO3-N, and NH4+-N, respectively (Table S1). The N removal efficiencies of 60–80% and areal removal rates of 400 mg-N m2 d1 for NO3-N, which estimated for the eco-ditch, were comparable to those reported for other surface-flow wetlands treating nitrate-dominated wastewaters. Karpuzcu et al. (2012) conducted a study on wetlands located in the San Joaquin River Valley (California, USA) receiving agricultural drainage from an irrigated field and reported areal NO3-N removal rates ranged from 142 to 350 mg-N m2 d1. Pilot-scale testing in the Prado Wetlands in Southern California measured around 80% NO3-N removal efficiencies during summer operations; average areal removal rates for NO3-N were around 500 mg-N m2 d1 (Horne, 2001). In the San Joaquin Marsh, a pound/marsh treatment wetland in Southern California, removal efficiencies averaged 80% for NO3-N and 60% for TN, and areal removal rates were around 300 mg-N m2 d1 (Fleming-Singer and Horne, 2006). Areal N removal rates also showed a strong seasonal trend (Fig. 3b). Removal rates were variable across months, and the trend of TN, NO3-N and NH4+-N removal rates were similar. The areal removal rates for TN, NO3-N and NH4+-N increased during the warm months and decreased during the cold months. Generally, areal removal rates were higher started in early April, and then peaked, finally began to reduce in September.
90
90
(b)
(a)
(%)
chemistry conditions change. Therefore, NH4+-N levels in the inflow, and its removal are more likely influenced by fieldresidence time.
(c)
85
80
85
80
70
80
75
60
75
70
50
70
65
40
65
30 60 Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec Jan Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec Jan Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec Jan 3000 250 1600 (f) (d) (e ) 1400 2500 200 1200 2000 1000 150
J (m g m -2 d-1 )
60
100
800
1500
600
1000
400
50
500 200
0
0
0 Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec Jan
Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec Jan
Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec Jan
Mon th
Mon th
Mon th
Fig. 3. Seasonal pattern of (a) nitrogen removal efficiency (h) and (b) areal removal rate (J) of the eco-ditch during the monitoring period (January to December in 2013).
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3.3. First-order areal N removal rate constant
491
2.5
(a) *
kNH +-N =3.6846 -10.14 56CNH +-N 4 4 2 R =0.2296 p<0.0001
1.5
-1
k (m d )
2.0
1.0
0.5
0.0 0.24
0.26
0.28
0.30
0.32
0.34
CNH +-N (mg L-1 ) 4
2.5
(b) k NO
-N
=-4.2821+3.7917CNO
3
-N
R =0.6442 p<0.0001
1.5
-1
k (m d )
3
2
2.0
1.0
0.5
0.0 1.2
1.4
1.6
1.8
CNO3--N (mg L-1 )
4.0
(c) k T N=-7.6184+3.3359CT N R2 =0.6410 p<0.0001
3.0
-1
k (m d )
The P–k–C model (Eq. (3)) was used to calculate k, the modified first-order areal removal rate constant (Table S2). Average values for kTN, kNO3-N and kNH4+-N were approximately 1.11, 0.62 and 0.79 m d1, respectively (Table S1). Lots of work has been developed concerning the treatment of municipal wastewater effluents which contain high biochemical oxygen demand (BOD) in treatment wetlands (Kadlec and Knight, 1996). In general, k values observed in this eco-ditch were greater than those observed in wetlands systems receiving municipal wastewater. Our results were consistent with other studies, in that wetlands treating low BOD, high NO3-N drainage water tend to have higher observed removal rates that wetlands treating high BOD low NO3-N municipal wastewater (Beutel et al., 2009; Fleming-Singer and Horne, 2006). Agricultural runoff dominated by runoff from cropland, which typically has low concentrations of organic carbon, low concentrations of NH4+-N, and higher relative concentrations of NO3-N (Beutel et al., 2009). Higher apparent removal kinetics in systems treating agricultural runoff with low BOD may be due to ammonification and nitrification altering net NO3-N changes across wetlands (Karpuzcu et al., 2012). Andersen et al. (2010) deemed that the removal rate constants were influenced by a number of factors such as depth, hydraulic loading rate and inflow concentration. Numerous studies have shown that nitrate removal in wetlands is typically modeled as a first-order process relative to inflowing nitrate concentration (Kadlec and Knight, 1996). In this study, values of k were found to correlate significantly with inflow concentrations (Fig. 4). kTN and kNO3-N correlated positively with CTN and CNO3-N, while kNH4+-N negatively correlated with CNH4+-N. However, our results were not consistent with Beutel et al. (2009) which reported that k in wetlands showed no correlation with inflow NO3-N concentration. In that study, they found a significantly correlation between k values and inflow and outflow temperature. This was to be expected, since denitrification is highly sensitive to temperature. Our study also found that k correlated positively with water temperature, the kN values were showed higher in warm month (April–August) than cold month (December–February) (Fig. S1). Moreover, k 3.4. Implications for management Based on the results of this study, potential benefits of the integration of eco-ditches into agricultural watersheds to reduce N concentrations are apparent. The eco-ditch in the current study is managed similarly to a free water surface wetland, to maximize N removal rates. There are currently no regulatory limits for N in agricultural drainage, and the eco-ditch is not managed to achieve a specific outflow concentration criterion. Previous studies suggested 0.5 mg l1 can be used as the critical concentration above which nuisance algae problem can occur in downstream ecosystems (Biggs, 2000). Based on this criterion, it should optimize the eco-ditch to achieve the target TN concentration of 0.5 mg l1. Kadlec and Wallace (2008) considered that empirical relationships between removal rates and loading may not be useful for wetland design, but we do find them important for operation and management. For instance, at a given loading rate, how much removal might one expect? Reedy et al. (1999) considered that greater loading rates provide greater opportunities for increased nutrient removal. Our study demonstrated that as the areal Nloading (mg-N m2 d1) increases areal removal rates tend to increase (Fig. S3). However, some experimental wetlands documented that percent removal of N concentration tends to decrease with areal removal rates increase (Mitsch and Gosselink, 2000).
2.0
1.0
0.0 2.2
2.4
2.6
2.8
3.0
3.2
-1
CTN (mg L ) Fig. 4. Relationship between (a) NH4+-N, (b) NO3-N and (c) TN areal rate constants (k) and influent concentrations. The linear regression equations, goodness of fit (R2) and their significance (p < 0.05) were indicated.
This relationship did not entirely found in the investigated ecoditch. Besides NH4+-N, the removal efficiency of NO3-N and TN tends to increase as the influent loading rate increase (Fig. S4). Looking at eco-ditch performance and characteristics during the long-term by using multiple lines of evidence helps provide context and improve understanding of N removal performance and system characteristics at the operational wetland-scale (Dunne et al., 2013). For example, we generally found higher N-removal
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during warmer summer/fall periods relative to cooler winter/ spring periods. Using seasonal data in a dynamic operational context, we can increase hydraulic loading during warmer periods when the system removes highest rates of N. 4. Conclusions In this study, we characterized N-removal in an eco-ditch receiving drainage from paddy fields. Results indicated that the eco-ditch removed significant amounts of N, mostly NO3-N, from agricultural runoff. Averaged removal efficiency (h), removal rate (J) and rate constants (k) of TN were 75.8%, 942.1 mg-N m2 d1 and 1.11 m d1, respectively. Moreover, N-removal rates were highly sensitive to temperature and exhibited seasonal trends. Removal rates in warm moths (April–August) were 2–4 times higher than in cooler months. The eco-ditch application for the treatment of drainage with high nutrients from paddy fields achieved good results, and will therefore be employed locally and in similar areas. However, to sustain high N-removal performance for the future, dynamic management of this system is required. Acknowledgments This study was financially supported by the Key CAS Programs (KZZD-EW-11, KZZD-EW-10-5, 100-Talents), the National Science and Technology Supporting Project (2014BAD14B01, 2012BAD14B17), and the China Postdoctoral Science Foundation Funded Project (2014M560648). We gratefully acknowledge the anonymous reviewers for their constructive comments and suggestions. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j. ecoleng.2015.05.012. References Andersen, J.H., Murray, C., Kaartokallio, H., 2010. A simple method for confidence rating of eutrophication status classifications. Mar. Pollut. Bull. 60, 919–924. Bennett, E.R., Moore, M.T., Cooper, C.M., Smith Jr., S., Shields Jr., F.D., Drouillard, K.G., Schulz, R., 2005. Vegetated agricultural drainage ditches for the mitigation of pyrethroid associated runoff. Environ. Toxicol. Chem. 24, 2121–2127. Beutel, M.W., Newton, C.D., Brouillard, E.S., Watts, R.J., 2009. Nitrate removal in surface-flow constructed wetlands treating dilute agricultural runoff in the lower Yakima Basin, Washington. Ecol. Eng. 35, 1538–1546. Biggs, B.J.F., 2000. Eutrophication of streams and rivers: dissolved nutrientchlorophyll relationships for benthic algae. J. N. Am. Benthol. Soc. 19, 17–31. Camargo, J.A., Alonso, A., Salamanca, A., 2005. Nitrate toxicity to aquatic animals: a review with new data for freshwater invertebrates. Chemosphere 58, 1255–1267. Cooper, C.M., Moore, M.T., Bennett, E.R., Smith Jr., S., Farris, J.L., Milam, C.D., 2004. Innovative uses of vegetated drainage ditches for reducing agricultural runoff. Water Sci. Technol. 49, 117–123.
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Nitrogen removal in an ecological ditch receiving agricultural drainage in subtropical central China Liang Chena,b , Feng Liua,b,* , Yi Wanga,b , Xi Lia,b , Shunan Zhanga,b,c , Yong Lia,b , Jinshui Wua,b,* a Changsha Research Station for Agricultural & Environmental Monitoring, Institute of Subtropical Agriculture, Chinese Academy of Science, Hunan 410125, PR China b Key Laboratory of Agro-ecological Process in Subtropical Region, Institute of Subtropical Agriculture, Chinese Academy of Science, Hunan 410125, PR China c Graduate University of the Chinese Academy of Sciences, Beijing 100039, PR China
A R T I C L E I N F O
A B S T R A C T
Article history: Received 22 November 2014 Received in revised form 11 April 2015 Accepted 23 May 2015 Available online xxx
Nitrogen (N) loss from agricultural field can cause eutrophication in downstream freshwater systems. Ecological ditches (eco-ditches) which are engineered to mitigate N loss from agricultural runoff. This study presents an analysis of water quality data for a 200 m long eco-ditch treating agricultural drainage in subtropical central China. Inflow concentrations of total N (TN) ranged from 2.3 to 3.1 mg l1 and typically contained 1.1–1.7 mg l1 nitrate (NO3-N) and <0.4 mg l1 ammonium (NH4+-N). Mean concentration removal efficiencies (h) in the eco-ditch for TN, NO3-N, and NH4+-N were 75.8, 63.7, and 77.9%, respectively. Mean area-based first-order removal rate constant (J) and removal rate constant (k) of TN were 942.1 mg-N m2 d1 and 1.11 m d1, respectively. Outflow from the eco-ditch typically contained <0.85 mg l1 TN and <0.70 mg l1 NO3-N. Rate of N loss in the eco-ditch was highly seasonal, generally peaking in the summer months (May–August). Results show that the eco-ditch can be effective at reducing transport of non-point source of N in-situ, particularly in warm environments. Its use should be expanded to similar areas, although managing the eco-ditch to maximize N removal in the long-term will require dynamic management. ã2015 Published by Elsevier B.V.
1. Introduction Agricultural intensification is recognized as a major source of increased nitrogen (N) in aquatic systems (McIsaac and Libra, 2003). Chemical N fertilizers are widely used for pursuing high crop yield. However, fertilizer additions most often exceed the N requirements of crops, thus resulting in a higher amount of ammonium (NH4+) and nitrate (NO3) leached from soils during storm events and flushed away in agricultural runoff (Tilman, 1999). Runoff from agricultural lands can potentially carry high nutrient loads to receiving waters and contribute to eutrophication, hypoxia, and ecological damage (Howarth et al., 2011). It has been reported that total N (TN) levels greater than 0.5 mg l1 can result in large masses of nuisance algae unless other factors limit algae growth (Biggs, 2000), and nitrate concentrations above 2.0 mg l1 can cause toxicity in a variety of
* Corresponding authors at: Changsha Research Station for Agricultural & Environmental Monitoring, Institute of Subtropical Agriculture, Chinese Academy of Science, Hunan 410125, PR China. Tel.: +86 73184615224; fax: +86 73184619736. E-mail addresses: [email protected] (F. Liu), [email protected] (J. Wu). http://dx.doi.org/10.1016/j.ecoleng.2015.05.012 0925-8574/ ã 2015 Published by Elsevier B.V.
freshwater organisms (Camargo et al., 2005). Therefore, developing best management practices (BMPs) that minimize nutrient runoff from agricultural fields is needed. Vegetated ecological ditches (eco-ditches) are a BMP being closely examined for nutrient mitigation (Kröger et al., 2007; Moore et al., 2010; Liu et al., 2013). Ditches are often designed and constructed as the main component of an agricultural irrigated system, which aims to irrigate or drain excess water from the fields. They also act as major pathways of surface and subsurface flow N from agricultural lands to receiving waters (Kröger et al., 2007; Herzon and Helenius, 2008). Eco-ditches are transformed from conventional agricultural drainage ditches, which consist of drainage channels, substrate, vegetation, and flow control facilities (Wu et al., 2013). They are well placed to mitigate contaminants in a manner similar to a surface-flow constructed wetland, filtering pollutants from runoff before exiting into downstream receiving systems. Studies have demonstrated vegetated eco-ditches’ capacity to mitigate pesticides (Bennett et al., 2005; Cooper et al., 2004; Moore et al., 2011). More recently, nutrient mitigation potential of eco-ditches has been investigated. Kröger et al. (2007) determined the inorganic N reduction capacity of agricultural
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ditches under natural, variable rainfall conditions in northern Mississippi, and results indicated that drainage ditches reduced the initial 2.5% inorganic N load to less than 1.5%, which is a 57% reduction over 2-years study period. Wu et al. (2011) reported a mean reduction of 52%, 53% and 58% for TN, NO3-N, and NH4+-N, respectively, for vegetated eco-ditches where the bottom part was packed with adsorbent. The main mechanisms of eco-ditches for N-removal are sediment retention, plant uptake, and microbial metabolic activities, and vegetation inside ditches are important to effectively slow down the water velocity, increase agrochemical retention, and subsequently provide better conditions for pollution dissipation in ditches (Hu et al., 2010; Kröger et al., 2007). Studies have demonstrated the potential for eco-ditches to decrease N loads exiting agricultural systems, but the use of degradation models to describe N mitigation performance is limited. Area-based first-order kinetic models are usually applied to describe of N removal in surface-flow wetland systems (Beutel et al., 2009; Karpuzcu and Stringfellow, 2012). Incorporation of monitoring data into established models for eco-ditch systems is needed to support dynamic management decisions related to changing performance of ecologically engineered ditches. Consequently, the aim of this study was to document the N mitigation performance of an eco-ditch based on a detailed evaluation of water quality for a year. The vegetated eco-ditch, located in the Tuojia catchment in central China, is mainly receives drainage water from paddy fields. Seasonal N concentration in conjunction with temporal hydrological variability provided the potential to generate N loads flowing within the eco-ditch. Efficiency and kinetics of N (TN, NH4+-N and NO3-N) removal were determined and analyzed.
2. Materials and methods 2.1. Study site The Tuojia catchment is an upstream subcatchment of the Jinjing River (28 300 –28 390 N, 113180 –113 260 E, elevation of 46– 452 m), located in Changsha County, Hunan Province, China. The site has a subtropical climate, with an annual mean air temperature of 17.5 C, a mean annual rainfall of 1330 mm, and an annual potential evaporation of 1300 mm. An agricultural drainage ditch (approximately 850 m in length) was the main channel of water from paddy fields to the Tuojia catchment at the study site. The ditch receives a contributing area of 28.6 ha, with forest and rice agriculture as the main land-use types in the catchments accounting for approximately 52.8% and 40.9% of the total land area, respectively (Fig. 1). Generally, no fertilizer is applied in the forest. Rice (Oryza sativa) is transplanted twice a year (mid-April and mid-July) and harvested at the end of June and in the middle of October. During rice planting, there are typically two fertilizer applications in each rice season, and paddy fields receive approximately 2146 kg N ha1 yr1 and 507 kg P ha1 yr1 annually. Intensive fertilizer applications usually result in high nutrient concentrations in the surface water of paddy fields. The eco-ditch was designed and constructed in the downstream section of the original drainage ditch in 2009 to reduce nutrient losses from paddy fields (Liu et al., 2013). The eco-ditch was 200 m in length, with a mean top width of 4.0 m, mean bottom width of 2.8 and an average depth of 1.0 m. The eco-ditch includes a series of weirs used as water control structures and various hydrophytes. Dominant vegetation at the eco-ditch includes Canna indica,
Fig. 1. Study site location in subtropical central China. Map of the eco-ditch (highlighted in frame), which is located to the west of the Tuojia River.
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Hydrocotyle vulgaris, Sparganium stoloniferum, Myriophyllum sp., and Juncus sp. Water discharge was monitored at the outlet of ecoditch using the observed flow velocity and the flow cross-section area method which described by our previous study (Wang et al., 2014). Site maintenance was undertaken when the eco-ditch began running.
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of N is not known, and it is reasonable to assume that P = N = 4, which is a mean N-value proposed for free-water surface wetlands (Kadlec and Wallace, 2008; Kadlec and Wallace, 2008). 3. Results and discussion 3.1. Nitrogen removal efficiency
2.2. Field sampling and lab analysis Water quality characterization of the eco-ditch is presented in Fig. 2. As can be seen from this figure, inflow concentration of TN, 0.4
(a)
0.3 -1
CNH +-N (mg L ) 4
During the 2013 study period (January–December), weekly water samples were collected at the inlet and outlet of the ecoditch. Water samples were collected in glass 500 ml bottles to guarantee the volume requirements in laboratory analysis. All bottles were rinsed with sample water prior to collection of a depth-integrated sample. Samples were immediately stored at 4 C after sampling and transported to the laboratory on the day of sampling. Samples were received by the laboratory, logged in and inspected for damage, and stored at 4 C until filtering and analysis. All filtration and preservation of samples were completed within 24 h. TN, NO3-N and NH4+-N concentrations were measured using an automatic flow injection analyzer (Fia-star 5000, Foss Tecator, Sweden) according to the standard methods for Water and Wastewater Monitoring and Analysis (SEPA, 2002). Briefly, water samples were filtered through a 0.45-mm membrane and the filtrates were used to directly determine the NO3-N and NH4+-N concentrations using the fully automated injection system. TN concentrations were determined using the fully automated injection system after digestion with a K2S2O8-NaOH solution.
0.2
Inlet Outlet
0.1
0.0 Jan Feb M ar Apr M ay Jun Jul Aug Sep Oct Nov Dec Jan 2.0
(b)
2.3. Kinetic and removal calculations
h¼
C in C out 100 C in
1.5 -1
CNO --N (mg L ) 3
Rates of removal for TN, NO3-N and NH4+-N were quantified using three common approaches for wetland systems (Kadlec and Knight, 1996). First, concentration removal efficiency (h, %) was calculated according to Eq. (1), where Cin and Cout are the inlet and outlet concentrations of TN, NO3-N or NH4+-N (mgl1) respectively.
Q ðC in C out Þ A
Inlet Outlet
0.5
(1)
Second, the areal removal rate (J, mg-N/m2/d) was calculated according to Eq. (2), where Q is the flow rate (m3 d1), A is the surface area of the eco-ditch (m2). J¼
1.0
0.0 Jan Feb M ar Apr M ay Jun Jul Aug Sep Oct Nov Dec Jan 4.0
(c)
(2)
C out C 1 ¼ p C in C ð1 þ k=PqÞ
-1
Finally, the modified first-order areal removal rate constants, k (m d1) were determined using the P–k–C* model (Eq. (3)), which is a relaxed tanks-in-series flow (TIS) model (Kadlec and Wallace, 2008).
CTN (mg L )
3.0
2.0
Inlet Outlet
(3)
where C* is the background concentration (mg l1), P is the apparent number of tanks in series, and q is the hydraulic loading (m d1). C* accounts for substances generated in the wetland by biological activity, sediment release, etc., as well as the nondegradable fraction of the contaminant. From observing our N data, it seemed that a C* concentration of 0, 0 and 0.2 mg l1 for NO3-N, NH4+-N and TN, respectively, was appropriate, as outflow concentrations of N tend toward a low concentration. P is a free-fitting parameter with the condition P < N, where N is the tracer TIS number. In the absence of a tracer test, the actual number
1.0
0.0 Jan Feb M ar Apr M ay Jun Jul Aug Sep Oct Nov Dec Jan
Month Fig. 2. Concentrations of (a) ammonium (NH4+-N), (b) nitrate (NO3-N) and (c) total nitrogen (TN) in the inlet and outlet of the eco-ditch. Monitoring period was from January 2013 to December 2013. Solid circles represent monthly inlet concentrations and open circles represent outlet concentrations.
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NO3-N and NH4+-N were typically greater than outflow concentration. Total N levels entering the eco-ditch over the study period (January–December, 2013) varied from 2.32 mg l1 to 3.05 mg l1, and were dominated by nitrate (42.6–60.3%). Inlet concentration of NO3-N and NH4+-N varied from 1.12 mg l1 to 1.68 mg l1 and 0.24 mg l1 to 0.33 mg l1, respectively (Fig. 2). Outflow concentrations were always lower for TN, NO3-N and NH4+-N. TN levels in the eco-ditch outflow ranged from 0.41 mg l1 to 0.83 mg l1, and NO3-N and NH4+-N levels in eco-ditch outflow were consistently <0.7 mg l1 and <0.1 mg l1, respectively. TN, NO3N and NH4+-N in eco-ditch inflow increased sharply as the transplanting season (April–May and July–August) progressed (Fig. 2). This may due to fertilizer application were applied during this period. The eco-ditch investigated in this study was observed high N removal efficiency. Mean TN, NO3-N and NH4+-N removal efficiencies were 75.8, 63.7 and 77.9%, respectively (Table S1). Moreover, changes of the N removal efficiencies are shown in Fig. 3a. Removal efficiencies (h) of TN, NO3-N tended to increase during summer periods, but this trend is not fully suited for NH4+N (Fig. 3). This seasonal pattern of h suggests the biological N transformation rate increased during warm summer periods. This was to be expected, since biological process is strongly affected by temperature. The primary mechanism for the loss of NO3-N in wetlands is denitrification, the microbial reduction NO3-N of to N2, which typically accounts for 60–95% of the NO3-N removal in wetlands (Spieles and Mitsch, 1999). Denitrification rates increase dramatically with temperature, within a lower and upper bounds of around 5 C and 70 C, respectively (Vymazal, 2007). Sirivedhin and Gray (2006) documented a two-order of magnitude increase in denitrification rates in wetland sediments when temperature was increased from 4 C to 25 C. Furthermore, it is the growing season for the wetland plants at the warm periods, so plant uptake would be another important way of N-removal. Inflow and outflow NH4+N concentrations were typically low (<0.3 mg l1). Because of the positive charge on the ammonium ion, NH4+-N is easily absorbed by sediments or detritus and can be released easily when water
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3.2. Areal N removal rate Historical observations at a range of wetlands show that levels of a number of pollutants, including NO3-N, generally decrease with distance thorough wetlands, suggesting that pollutant removal in wetlands is first order in nature (Kadlec and Knight, 1996). In this study, areal-based first-order removal rates for TN, NO3-N and NH4+-N were calculated from flow and water quality measurements (Eq. (2)). Averaged removal rates (J) were 942.1, 407.8, and 96.1 mg-N m2 d1 for TN, NO3-N, and NH4+-N, respectively (Table S1). The N removal efficiencies of 60–80% and areal removal rates of 400 mg-N m2 d1 for NO3-N, which estimated for the eco-ditch, were comparable to those reported for other surface-flow wetlands treating nitrate-dominated wastewaters. Karpuzcu et al. (2012) conducted a study on wetlands located in the San Joaquin River Valley (California, USA) receiving agricultural drainage from an irrigated field and reported areal NO3-N removal rates ranged from 142 to 350 mg-N m2 d1. Pilot-scale testing in the Prado Wetlands in Southern California measured around 80% NO3-N removal efficiencies during summer operations; average areal removal rates for NO3-N were around 500 mg-N m2 d1 (Horne, 2001). In the San Joaquin Marsh, a pound/marsh treatment wetland in Southern California, removal efficiencies averaged 80% for NO3-N and 60% for TN, and areal removal rates were around 300 mg-N m2 d1 (Fleming-Singer and Horne, 2006). Areal N removal rates also showed a strong seasonal trend (Fig. 3b). Removal rates were variable across months, and the trend of TN, NO3-N and NH4+-N removal rates were similar. The areal removal rates for TN, NO3-N and NH4+-N increased during the warm months and decreased during the cold months. Generally, areal removal rates were higher started in early April, and then peaked, finally began to reduce in September.
90
90
(b)
(a)
(%)
chemistry conditions change. Therefore, NH4+-N levels in the inflow, and its removal are more likely influenced by fieldresidence time.
(c)
85
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30 60 Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec Jan Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec Jan Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec Jan 3000 250 1600 (f) (d) (e ) 1400 2500 200 1200 2000 1000 150
J (m g m -2 d-1 )
60
100
800
1500
600
1000
400
50
500 200
0
0
0 Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec Jan
Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec Jan
Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec Jan
Mon th
Mon th
Mon th
Fig. 3. Seasonal pattern of (a) nitrogen removal efficiency (h) and (b) areal removal rate (J) of the eco-ditch during the monitoring period (January to December in 2013).
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3.3. First-order areal N removal rate constant
491
2.5
(a) *
kNH +-N =3.6846 -10.14 56CNH +-N 4 4 2 R =0.2296 p<0.0001
1.5
-1
k (m d )
2.0
1.0
0.5
0.0 0.24
0.26
0.28
0.30
0.32
0.34
CNH +-N (mg L-1 ) 4
2.5
(b) k NO
-N
=-4.2821+3.7917CNO
3
-N
R =0.6442 p<0.0001
1.5
-1
k (m d )
3
2
2.0
1.0
0.5
0.0 1.2
1.4
1.6
1.8
CNO3--N (mg L-1 )
4.0
(c) k T N=-7.6184+3.3359CT N R2 =0.6410 p<0.0001
3.0
-1
k (m d )
The P–k–C model (Eq. (3)) was used to calculate k, the modified first-order areal removal rate constant (Table S2). Average values for kTN, kNO3-N and kNH4+-N were approximately 1.11, 0.62 and 0.79 m d1, respectively (Table S1). Lots of work has been developed concerning the treatment of municipal wastewater effluents which contain high biochemical oxygen demand (BOD) in treatment wetlands (Kadlec and Knight, 1996). In general, k values observed in this eco-ditch were greater than those observed in wetlands systems receiving municipal wastewater. Our results were consistent with other studies, in that wetlands treating low BOD, high NO3-N drainage water tend to have higher observed removal rates that wetlands treating high BOD low NO3-N municipal wastewater (Beutel et al., 2009; Fleming-Singer and Horne, 2006). Agricultural runoff dominated by runoff from cropland, which typically has low concentrations of organic carbon, low concentrations of NH4+-N, and higher relative concentrations of NO3-N (Beutel et al., 2009). Higher apparent removal kinetics in systems treating agricultural runoff with low BOD may be due to ammonification and nitrification altering net NO3-N changes across wetlands (Karpuzcu et al., 2012). Andersen et al. (2010) deemed that the removal rate constants were influenced by a number of factors such as depth, hydraulic loading rate and inflow concentration. Numerous studies have shown that nitrate removal in wetlands is typically modeled as a first-order process relative to inflowing nitrate concentration (Kadlec and Knight, 1996). In this study, values of k were found to correlate significantly with inflow concentrations (Fig. 4). kTN and kNO3-N correlated positively with CTN and CNO3-N, while kNH4+-N negatively correlated with CNH4+-N. However, our results were not consistent with Beutel et al. (2009) which reported that k in wetlands showed no correlation with inflow NO3-N concentration. In that study, they found a significantly correlation between k values and inflow and outflow temperature. This was to be expected, since denitrification is highly sensitive to temperature. Our study also found that k correlated positively with water temperature, the kN values were showed higher in warm month (April–August) than cold month (December–February) (Fig. S1). Moreover, k 3.4. Implications for management Based on the results of this study, potential benefits of the integration of eco-ditches into agricultural watersheds to reduce N concentrations are apparent. The eco-ditch in the current study is managed similarly to a free water surface wetland, to maximize N removal rates. There are currently no regulatory limits for N in agricultural drainage, and the eco-ditch is not managed to achieve a specific outflow concentration criterion. Previous studies suggested 0.5 mg l1 can be used as the critical concentration above which nuisance algae problem can occur in downstream ecosystems (Biggs, 2000). Based on this criterion, it should optimize the eco-ditch to achieve the target TN concentration of 0.5 mg l1. Kadlec and Wallace (2008) considered that empirical relationships between removal rates and loading may not be useful for wetland design, but we do find them important for operation and management. For instance, at a given loading rate, how much removal might one expect? Reedy et al. (1999) considered that greater loading rates provide greater opportunities for increased nutrient removal. Our study demonstrated that as the areal Nloading (mg-N m2 d1) increases areal removal rates tend to increase (Fig. S3). However, some experimental wetlands documented that percent removal of N concentration tends to decrease with areal removal rates increase (Mitsch and Gosselink, 2000).
2.0
1.0
0.0 2.2
2.4
2.6
2.8
3.0
3.2
-1
CTN (mg L ) Fig. 4. Relationship between (a) NH4+-N, (b) NO3-N and (c) TN areal rate constants (k) and influent concentrations. The linear regression equations, goodness of fit (R2) and their significance (p < 0.05) were indicated.
This relationship did not entirely found in the investigated ecoditch. Besides NH4+-N, the removal efficiency of NO3-N and TN tends to increase as the influent loading rate increase (Fig. S4). Looking at eco-ditch performance and characteristics during the long-term by using multiple lines of evidence helps provide context and improve understanding of N removal performance and system characteristics at the operational wetland-scale (Dunne et al., 2013). For example, we generally found higher N-removal
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during warmer summer/fall periods relative to cooler winter/ spring periods. Using seasonal data in a dynamic operational context, we can increase hydraulic loading during warmer periods when the system removes highest rates of N. 4. Conclusions In this study, we characterized N-removal in an eco-ditch receiving drainage from paddy fields. Results indicated that the eco-ditch removed significant amounts of N, mostly NO3-N, from agricultural runoff. Averaged removal efficiency (h), removal rate (J) and rate constants (k) of TN were 75.8%, 942.1 mg-N m2 d1 and 1.11 m d1, respectively. Moreover, N-removal rates were highly sensitive to temperature and exhibited seasonal trends. Removal rates in warm moths (April–August) were 2–4 times higher than in cooler months. The eco-ditch application for the treatment of drainage with high nutrients from paddy fields achieved good results, and will therefore be employed locally and in similar areas. However, to sustain high N-removal performance for the future, dynamic management of this system is required. Acknowledgments This study was financially supported by the Key CAS Programs (KZZD-EW-11, KZZD-EW-10-5, 100-Talents), the National Science and Technology Supporting Project (2014BAD14B01, 2012BAD14B17), and the China Postdoctoral Science Foundation Funded Project (2014M560648). We gratefully acknowledge the anonymous reviewers for their constructive comments and suggestions. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j. ecoleng.2015.05.012. References Andersen, J.H., Murray, C., Kaartokallio, H., 2010. A simple method for confidence rating of eutrophication status classifications. Mar. Pollut. Bull. 60, 919–924. Bennett, E.R., Moore, M.T., Cooper, C.M., Smith Jr., S., Shields Jr., F.D., Drouillard, K.G., Schulz, R., 2005. Vegetated agricultural drainage ditches for the mitigation of pyrethroid associated runoff. Environ. Toxicol. Chem. 24, 2121–2127. Beutel, M.W., Newton, C.D., Brouillard, E.S., Watts, R.J., 2009. Nitrate removal in surface-flow constructed wetlands treating dilute agricultural runoff in the lower Yakima Basin, Washington. Ecol. Eng. 35, 1538–1546. Biggs, B.J.F., 2000. Eutrophication of streams and rivers: dissolved nutrientchlorophyll relationships for benthic algae. J. N. Am. Benthol. Soc. 19, 17–31. Camargo, J.A., Alonso, A., Salamanca, A., 2005. Nitrate toxicity to aquatic animals: a review with new data for freshwater invertebrates. Chemosphere 58, 1255–1267. Cooper, C.M., Moore, M.T., Bennett, E.R., Smith Jr., S., Farris, J.L., Milam, C.D., 2004. Innovative uses of vegetated drainage ditches for reducing agricultural runoff. Water Sci. Technol. 49, 117–123.
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