Skip to main content
NCBI home page
ACS AuthorChoice logoLink to ACS AuthorChoice
. 2020 Jul 26;54(15):9584–9592. doi: 10.1021/acs.est.0c00869

Environmental Tradeoffs between Nutrient Recycling and Greenhouse Gases Emissions in an Integrated Aquaculture–Agriculture System

Thomas Groenveld †, Naftali Lazarovitch †, Yair Y Kohn ‡, Ilya Gelfand †,*
PMCID: PMC7460073  PMID: 32790417

Abstract

graphic file with name es0c00869_0006.jpg

The unlimited nitrogen (N) availability that has characterized crop production in the last few decades is accompanied by environmental burdens, including the greenhouse gas (GHG) emissions associated with fertilizer production, post-application nitrate (NO3–) pollution of water bodies, and emissions of reactive gaseous N forms into the atmosphere. Here, we quantified the environmental tradeoffs of replacing mineral N fertilizer with NO3– and ammonium (NH4+) originating from effluent water of aquaculture in a cucumber (Cucumis sativus) cultivation system. While the yield, nitrogen use efficiency (NUE), and NO3– leaching were similar between the cucumbers fertilized and irrigated (fertigated) by aquaculture effluent water containing 100 mg of NO3–-N L–1 (AN), by aquaculture effluent water supplemented with NH4+ (AN+), or by tap water with NO3– and NH4+ added (FN+), there were significant differences in the nitrous oxide (N2O) emissions between the systems. The N2O emissions peaked after each irrigation event followed by an exponential decline. The cumulative N2O emissions were between 60 and 600 g N2O-N ha–1, smaller than predicted based on a fertilizer application rate of 600 kg N ha–1 and were in the order AN+ ≫ FN+ > AN.

Introduction

With the growing global population, worldwide food demand is increasing along with concerns about the environmental impact of food production.1 A nutritionally balanced diet cannot depend only on cereals and meat consumption but must include vegetables,2,3 resulting in an increased demand for vegetable production. In 2018, ∼9.4 × 108 Mg of fresh vegetables were produced globally on an area of ∼6.4 × 107 ha, and this production is predicted to expand further.4 One promising method that can increase the sustainability of vegetable production is to integrate it with intensive fish production.5,6 Integration of aquaculture and agriculture allows the recycling of nutrients excreted by fish due to nonefficient nutrient use (i.e., nitrogen (N) and phosphorus (P)7), which otherwise require removal at a centralized wastewater treatment plant or are discharged into the environment, leading to contamination of surface and ground waters.8,9 An additional benefit is that the use of effluent water for vegetable fertigation (fertilization through irrigation) allows the recycling of otherwise wasted water for secondary crop production, increasing the water use efficiency (WUE) of the integrated system. This is especially appealing in semi-arid countries with limited water availability.

Continuous depletion of the wild fish supply due to overfishing has led to an increase in aquaculture, including the use of recirculating aquaculture systems (RASs) around the world,10 potentially ensuring an uninterrupted supply of effluent water for vegetable fertigation. RASs have the potential to reach a wide range of nitrate (NO3–) concentrations suitable for fertigation,11 and the modularity of such systems makes it possible to install them in series with fertigated crop production systems.5 While water reuse benefits of RASs are well described,12 fertigation with N originating from fish production is not well understood as the addition of large quantities of readily available N and carbon (C) sources13 may enhance denitrification and nitrous oxide (N2O) emissions from the integrated system,14 contributing to their negative environmental impact. Furthermore, the reuse of effluent N from RASs allows for a reduction in dependence on mineral fertilizers and the carbon dioxide (CO2) emissions associated with their production.15 The N cycle in agriculture is inherently leaky with ∼50–70% of applied N lost to the environment.15 Two major forms of N pollution from horticulture are N2O16 and NO3–.17 The NO3– leaches out of the root zone and contaminates groundwater and water bodies,8 while N2O is the third major greenhouse gas (GHG) and the agent of ozone (O3) destruction in the stratosphere.18 Agriculture is responsible for the majority of new reactive N in the environment,19,20 and mineral fertilizers are the biggest input of new N into the agricultural system.21

In general, microbial nitrification and denitrification are responsible for most N2O production in agroecosystems,22 including aquaculture.23 However, a range of other microbial metabolic pathways that produce or consume N2O complicates the determination of the specific sources of N2O emissions measured in the field and their controlling factors.24 Due to the similarity of the required environmental conditions, some pathways can only be determined with a degree of certainty through isotopic studies.25,26 Nitrous oxide emission due to heterotrophic nitrification and the aerobic denitrification associated with it are examples of pathways that are often grouped together with others (heterotrophic denitrification and autotrophic nitrification) and are frequently neglected from overviews of N2O production pathways,24,27−29 though included in others.30−33 Finally, traces of anammox bacteria have been found in aquaculture systems,34 but a properly designed biofilter is required to use this pathway for efficient N removal.35

Regardless of the specific N2O production pathway, the emission factors (EFs) or the fertilizer N-induced N2O emissions can be estimated for aquaculture, as has been done for agriculture.36 The EF estimation for aquaculture, however, is not done often, and knowledge about GHG emissions from the aquaculture and integrated systems is lacking. Recently, Hu et al.23 estimated that ∼6% of all anthropogenic N2O emissions come from aquaculture using an EF of 1.8% taken from a study of activated sludge wastewater treatment plants.37 The United Nations Environment Programme (UNEP) report38 citing the same publication used an EF of 1%. Such an inconsistency in estimations of EF for aquaculture is due to data scarcity. A handful of EF estimations for aquaculture have been published, generally reporting low values between 0.5 and 0.8%.39−41 Because the optimal conditions for fish and plant cultivation are different (pH,42 nutrient concentrations,43 and temperature11) and recycling drainage water of plants back to the fish (i.e., aquaponic system) requires a much more expensive infrastructure,6 the majority of future integrated aquaculture–agriculture systems (IAAS) are expected to be constructed with unidirectional water and nutrient flow.5,11 Furthermore, such a system allows for the integration of the sludge treatment (feces and food waste removed from the RAS) into the IAAS, which will further increase the NUE of the integrated system. It has been suggested that, after undergoing anaerobic digestion, this sludge is a valuable potential source of NH4+ for plant cultivation.11,44

We are not aware of any published research assessing direct N2O emissions from the plant production module of an IAAS where the plants were grown in soil. Thus, the overall objective of this study was to understand the N balance and cycle within the vegetable production component of an IAAS with an emphasis on sources of gaseous N2O emissions. To this end, an IAAS11 was modified to inject the supernatant from anaerobically digested sludge with a high NH4+ concentration into irrigation water to change the NH4+:NO3– ratio to that found in the mineral fertilizer used as the control treatment. The effect of the fertigation by the aquaculture effluent with and without NH4+ addition, on yield, NUE, NO3– leaching, and N2O emissions from a cucumber cultivation was assessed in comparison to cultivation using a mineral fertilizer with the same nutrient composition. The specific aim was to quantify the N2O emissions and to compare the overall environmental impact of the fertigation in the integrated and conventional systems with an emphasis on N pollution pathways.

Materials and Methods

Recirculating Aquaculture System

The experimental setup was based on the one used by Groenveld et al.11 In brief, cucumbers (Cucumis sativus) were grown in 18 lysimeters filled with a local sandy soil and were fertigated from three different sources: aquaculture wastewater with NO3– (AN), aquaculture wastewater with NO3– and NH4+ added from the anaerobic digester (AN+), and fresh water with fertilizer (MOR+, ICL, Israel) containing NO3– and NH4+ (FN+).

The two RASs used to produce AN and AN+ irrigation water each consisted of a polypropylene fish tank (800 L), a clarifier (300 L), and a moving bed bioreactor (MBBR; 400 L) with a flow rate of 0.5 m3 h–1. Barramundi fish (Lates calcarifer) were grown in the RASs for 6 months prior to the start of the experiment at which time they were weighed. Each system was provided with 500 g of food daily (45% protein, Raanan Fish Feed, Israel) for the duration of the experiment, which was 2% of the initial fish biomass and was the sole source of N into the system. Sludge was removed by flushing the clarifier every 2 h (resulting in about 30 L of sludge removed per kg feed) into an anaerobic sludge collection tank (500 L). The organic matter sunk to the bottom of the tank where it was anaerobically digested, and the reduced N remained in the supernatant as NH4+.

Nitrification of the NH4+ excreted by the fish took place in the MBBR on plastic biobeads (Aridal Ltd., Israel) where dissolved oxygen (DO) was maintained at about 90% saturation (∼7.5 mg L–1). A heating coil in the reactor kept water in the fish tank between 28 and 30 °C. The MBBRs of the two RASs were connected with a pump that exchanged water between them so that their NO3–-N concentration would be the same at around 100 mg L–1. Once a day, 150 L of water was transferred from the MBBR through a 0.13 mm screen filter (with continual backwash to the biofilter) to an intermediary irrigation tank and was replaced with fresh water. The AN irrigation water was left as is, and the supernatant from the sludge tank, which had been diluted to a NH4+-N concentration of 100 mg L–1, was added to AN+ irrigation water to bring the NH4+: NO3– ratio to about 1:10, similar to that of the synthetic fertilizer. The irrigation tanks were emptied before refilling each day.

Liquid NPK (4:2.5:6) fertilizer with microelements (MOR+ (NH4+:NO3– 1:10), ICL, Israel) was used to prepare fertilized water for the FN+ treatment. This solution was remade twice a week to eliminate NH4+ oxidation within the irrigation tank. The prepared solutions were continually mixed and cooled to 24 °C but not aerated. Potassium sulfate (K2SO4, Solucros, Belgium), potassium phosphate (MKP, Haifa Chemicals, Israel), and a cocktail of micronutrients (Super Koratin, ICL, Israel) were added to all the irrigation treatments to keep their levels similar (Table S2).

The pH in the RASs was kept at 6.5 by dosing potassium hydroxide (KOH), while tap water had a pH of about 8.2 and was brought down to 6.5 for the FN+ treatment by dosing hydrochloric acid (HCl); both processes were done by means of automated dispensers (Profilux, GHL, Germany).

Irrigation water was sampled every 2 days to measure the electrical conductivity (EC) (DDS 120 W, Bante Instruments, China), NO3–-N (Reflectoquant nitrate test 116971, Merck KGaA, Germany), NH4+-N (ammonium test 11117, Merck KGaA, Germany), NO2–-N (nitrite test 14658, Merck KGaA, Germany), and pH (HI 9126, Hanna Instruments, RI USA). The DO levels were measured in the irrigation water tank itself (Handy Polaris, OxyGuard, Denmark). Drainage water of each lysimeter was sampled every 2 days and tested for NO3–-N, NH4+-N, pH, and EC by the same methods listed for irrigation water. Total organic C (TOC) was measured in irrigation water every 5 days and three times in drainage water at periodic intervals (multi N/C 2100S; Analytik Jena, Germany). N as a dissolved organic material (DOM-N) was assumed to be about 6% of N in the aquaculture effluent as this was measured in earlier experiments with the same feeding and water exchange rates.11 The supernatant from the sludge digestion tanks was tested twice a week for NH4+ (Reflectoquant ammonium test 116977, Merck KGaA, Germany). Samples from all of the irrigation water treatments were tested twice for other nutrients essential for plant growth (K, Ca, Mg, Na, Mn, Fe, Zn, Cu, and B) by means of atomic absorption (AA240FF, Varian, USA); P and K by means of spectrophotometry, (Lambda 25, Perkin Elmer, USA); and Cl by means of titration, (665 Dosimat, Metrohm, Switzerland)) to confirm that these were being added in the right proportion (Table S2).

Plant Cultivation

Cucumbers (Sanyal, Soli, Israel) were grown in round 50 L lysimeters of 40 cm in diameter fertigated from each of the three treatments in six replicates (total of 18 lysimeters). The growth medium was sandy soil (3% 1–2 mm, 29% 0.5–1 mm, 52% 0.25–0.5 mm, 15% 0.125–0.25 mm, and 2% <0.125 mm) packed at an initial bulk density of 1.54 kg L–1. Below the sand, a layer of highly permeable geotextile (polyester fibers, Perlon, Germany) separated the sand from a 5 cm layer of plastic biobeads (Aridal Ltd., Israel) to ensure that the bottom boundary condition was equal for the whole container. The 35 cm sand above this layer ensured that the upper root zone was sufficiently aerated. The soil–air oxygen concentration was measured continuously (KE-50, Figaro, Japan) at a 5 cm depth next to the base of the plant. From the water collection layer, the drainage flowed into a separate container from which it was pumped out, weighed, and sampled to perform water and N balances. The daily transpiration was calculated for each lysimeter by subtracting the amount of drainage from the amount of irrigation water.

One cucumber seedling was transplanted into each lysimeter on 5/5/19; from here on, dates are referred to as days after transplant (DAT). The lysimeters were irrigated with water from their respective treatments 3 weeks prior to the start of the experiment. After the transplantation, each replicate was irrigated five times per day with 1 L of water till the measured transpiration reached 2 L per day per lysimeter after which each lysimeter was irrigated at 2.5 times the measured transpiration, rounded up to 1 L increments.11 The soil surface was covered with white geotextile (polyester fibers, Perlon, Germany) to reduce evaporation while allowing free gas movement. One runner of each plant was trained up a single string, and all side branches were removed by cutting after the first node of each side branch. Cucumbers were harvested daily from 21 DAT till the end of the experiment at 51 DAT. From 21 DAT, one replicate of each treatment was destructively sampled weekly to determine the dry biomass and N content of the plant; therefore, the number of replicates decreased over time from six at the beginning of the experiment to one for the final measurements. Leaves, shoots, and fruits were dried separately and were tested for the C and N content every week by combustion (OEA-CHNS Flash, 2000, Thermo Fisher Scientific, MA, USA) after grinding to pass through a 200 μm screen. Lysimeters of replicates that were destructively sampled were repacked with fresh soil (to an initial bulk density of 1.54 kg L–1), returned to their former location in the greenhouse, and continued to be fertigated according to their respective treatment.

N2O Emissions

Measurements

A slightly modified static chamber method was used for estimating N2O emissions after Gelfand et al.45 The 10 cm tall permanent collar was placed adjacent to the plant and irrigation water emitter, inserted ∼2–3 cm into the soil, and left in place for the duration of the experiment in every lysimeter. During the measurements, the collar was covered with a lid, which had inlet and outlet sampling ports and a pressure stabilization coil made out of a ∼1 mm diameter tube.46 The total static chamber volume during sampling was ∼3 L. The chamber was connected to an ICOS N2O/CH4 analyzer (M1-919; ABB-LGR, Cary, NC, USA), and air from the chamber was continuously circulated through the analyzer, which measured the N2O concentration inside the chamber and recorded the average every 5 s. The N2O fluxes were calculated as the linear change of the gas concentration in the chamber headspace over an incubation period of ∼1 min, corrected for ambient air temperature. Measurements of N2O started from DAT 7 after plant acclimation and were carried out at least once a week, usually for two consecutive days, so that each lysimeter was sampled multiple times throughout the day. Due to the high irrigation frequency, emission measurements of each individual lysimeter were performed at different time intervals from the fertigation, which is reflected in high flux variability.

Dissolved N2O

Nitrous oxide dissolved in irrigation water was measured by placing a 250 mL water sample into a glass container of a 1 L volume with two pipes inserted into the sealing lid, bringing the headspace air to and from the N2O analyzer. The pipe with the returning air extended to below the surface of the sample so that it bubbled up through it. After measuring the temperature of water, the container was closed, the sample was shaken profusely for 1 min, and then remained still till the N2O reading stabilized, which was considered the point at which the air space concentration was at equilibrium with that of water. The difference between the atmospheric and equilibrium concentration and the water temperature was used to calculate the amount of N2O dissolved in the 250 mL sample, according to Weiss and Price.47

Degassing Estimation

To test whether the irrigation water degassing was responsible for part of the observed N2O emissions, the emission rates over time after fertigating with water of different dissolved N2O concentrations were compared. Different dissolved N2O concentrations were prepared from the same AN+ treatment irrigation water tank by heating water to 90 °C, transferring it between two containers to ensure a large gas exchange area, and finally cooling it to ∼30 °C. The NO3–, NH4+, and nitrite (NO2–) concentrations of heated and untreated water remained the same, but the heated sample had no dissolved N2O. After measuring the dissolved N2O concentration in the two water samples, 1 L of each was used to irrigate a lysimeter without plants, which had been fertigated with water from the AN+ treatment for more than a month. The N2O emissions were measured repeatedly (>five times) in the 1.5 h period following the irrigation, and the entire process was repeated three times.

Carbon Addition Experiments

To test whether the N2O production was limited by available C in the FN+ treatment, dextrose (C6H12O6) was added to water from this treatment at 2.5 g L–1. Water was incubated at room temperature overnight, and concentrations of dissolved inorganic N (NO3–, NH4+, NO2–, and N2O) were measured before and after the incubation and were identical. Two lysimeters of the FN+ treatment without plants that had been fertigated with water from the FN+ treatment for more than a month were irrigated with 1 L of water with added C each. One similar control lysimeter was irrigated with FN+ water without the added C. The N2O emissions were measured repeatedly following the irrigation; after 7 and 18 h, this cycle was repeated in the same lysimeters.

N2O Emissions Model

As the biggest factor influencing the N2O emissions was the time after irrigation (Figure S1), an exponential model was fit to all the emission data measured during the experiment for each treatment

graphic file with name es0c00869_m001.jpg 1

where E is the rate of N2O emission (μg N m–2 min–1), t is time after the irrigation event (min), E0 is the rate of N2O emission at t = 0 (μg m–2 min–1), and k is the rate of decline (min–1). The E0 and k parameters were determined for each set of measurement data by minimizing the sum of square errors of modeled to measured data by means of the generalized reduced gradient (GRG) nonlinear solving method in the solver function of MS Excel (Microsoft Inc.) without any additional constraints. This model of N2O emission with time after an irrigation event was used to calculate the cumulative N2O emissions for each treatment by integrating the emission rates between irrigation events. The use of linear interpolation between the measurements, as is commonly done to calculate cumulative flux for field measurements,48 was not possible due to pulse behavior of the post-irrigation flux (Figure S1). The methodology for the estimation of the pathways involved in the N2O production is reported in Table S1.

Carbon Footprint, Nitrogen Use Efficiency, and Yield-Scaled N2O Emission Calculations

For estimation of the C footprint, we calculated the global warming impact (GWI), including interpolated N2O emissions and the C cost of the N fertilizer (after Gelfand and Robertson49) using a 100 year time horizon for all GWI calculations.50 We used a factor of 298 g of a carbon dioxide equivalent (CO2e) per g N2O for N2O GWI estimation and 8 g of CO2e per g N of ammonium nitrate fertilizer production.51 The GWI was normalized to Mg of fresh yield to facilitate comparison with other studies.52

As other works have been done concerning the GWI of RASs44,53,54 and given that the focus of the current study was the environmental impact of integrating such systems with horticulture, only the GWI of the vegetable production module was calculated, assuming that water would be discharged into the sewage system if not used for vegetable production. The GWI was calculated as the sum of the GWIs of N2O emissions and fertilizer production. Finally, the NUE was calculated as the N retained in the harvested plant biomass over the N applied (g N in biomass g–1 N applied);52 the fact that the N in aquaculture was already used once for fish production (within the fish feed) was not taken into consideration in this calculation.

Data Analysis

The irrigation treatment data was tested for normality with the Kolmogorov–Smirnov test, and for homogeneity of variance with Levene’s test, a one-way analysis of variance (ANOVA) with a post hoc Tukey test was used to determine significantly different interactions between tested groups when the ANOVA p value was below 0.05. The data was analyzed using on-line resources.55,56

Results and Discussion

Irrigation water’s NO3– and NH4+ concentrations were steady over time, and their average values are shown in Table 1. The NH4+ concentration was slightly lower for AN+ and higher for FN+ than the target, but the total amounts of N fertigated were not significantly different. The cumulative transpiration (Figure 1A) showed a linear increase over time, indicating continual healthy plant growth. Transpiration reached 2 L per plant at 20 DAT; the resulting irrigation over transpiration ratio (I/T) average for all treatments after that time is shown on the same graph, and it remained close to the target I/T of 2.5.

Table 1. Average Concentrations of Different N Forms, Total Organic Carbon (TOC), and Dissolved Oxygen (DO) that Were Measured in Irrigation Water (the Standard Deviation in Parentheses, n Varies between 12 and 50)a.

  mg L–1
system/parameter AN AN+ FN+
TOC 40.0 (5.0) a 39.0 (5.0) a 9.0 (3.0) b
DO 8.4 (1.3) a 6.7 (0.6) b 7.8 (0.5) a
NO3– 100.0 (17.0) a 99.0 (10.0) a 96.0 (11.0) a
NH4+ 1.2 (0.5) a 8.3 (2.1) b 13.9 (4.3) c
NO2– 0.6 (0.3) a 3.4 (1.2) b 3.6 (4.6) b
N2O 0.0018 (0.0028) a 0.0082 (0.0069) b 0.0003 (0.0000) a
total N 102 a 111 a 114 a
Open in a new tab
a

Lowercase letters identify significant differences between the treatments for the individual parameter.

Figure 1.

Figure 1

Open in a new tab

(A) Average cumulative transpiration and the average irrigation over transpiration rate (I/T; L L–1) and (B) cucumber yield (fresh weight) and N use efficiency (NUE; g N applied g–1 N in plant) for all treatments.

The average cumulative cucumber yields per treatment (Figure 1B) were similar for all treatments. The NUE (Figure 1B) was calculated for one lysimeter per treatment per week on the day that it was destructively sampled and was ∼40%. The slight increase of the NUE over time is due to the fact that the N was supplied in relation to the transpiration level, while an increasing proportion of the N was incorporated into the fruit, which transpires less per unit of N taken up than the vegetative parts of the plant.57 This caused the N application/N uptake ratio to decrease over time despite the steady I/T regime.

On average, we accounted for 87% of applied N in measured nongaseous N forms (Figure 2). As N2O emissions were calculated to not exceed 0.1% (see below), the 13% of N missing from the balance can be assumed to be N2 and NO but was unquantified in our study. The allocation of the N fertigated to different plant parts and to drainage water was calculated by linear interpolation of the treatment-averaged dry matter N concentration. After 21 DAT, the N allocation to the fruits increased rapidly in relation to the leaves and stems. The difference between the fertigated N and what was accounted for in drainage water and plant uptake was assumed to be the amount of total gaseous N lost and was similar between treatments. A growth curve of the different parts of the cucumber plants per treatment is presented in Figure S2.

Figure 2.

Figure 2

Open in a new tab

Average allocation of the total applied N in fertigation (dashed line) for each of the treatments. NH4+ and N-DOM levels in the drainage are too small to be seen.

The ratio of drainage NO3– to that taken up by the plant was significantly higher in the first 3 weeks than during the rest of the short growing season; at DAT 20, already 45% of the total NO3– drainage had leached out, while only 20% of the N uptake by the plant had taken place. Such excessive fertigation to ensure proper seedling establishment is a common practice in horticulture (personal communication with local farmers), and addressing this issue could potentially prevent a significant part of NO3– leaching. A technical solution could be to increase the irrigation frequency while decreasing the irrigation amount. Some potential problems with this solution include the magnification of any lack of homogeneity in the flow rate between drippers, and in desert climates, the temperature of water remaining in the dripline between irrigation events can get very hot, exposing the sensitive plant root zone to frequent pulses of scalding water.

On average, 38% of the N applied was leached as NO3–-N, or 1 g kg–1 fresh weight of cucumber harvested, as opposed to 19% by Groenveld et al.11 in which the same variety of cucumbers was grown on a shallow bed of perlite at the same I/T but at a 3 times higher irrigation frequency. Grewal et al.58 reported that 59% of the applied N was removed with the drainage in a hydroponic cucumber production system that had a lower normalized yield; this, however, may include gaseous N emissions. Yao et al.59 did not measure the NO3– leached, but a mass balance estimate would put the NO3– leaching at around 10 g of NO3–-N kg–1 cucumber.

The average of N2O emission measurements was higher for AN+ than for the other treatments, but the variability was very large (Figure 3B). Measurements of the N2O emissions from a single lysimeter over time after an irrigation event show that there was a strong decline in emissions over time after the fertigation event (Figure 3A). The exponential model (eq 1) fit the data well, as can be seen from the R2 values reported in Table 2, and was subsequently used to describe the change in N2O emissions with time after irrigation per treatment for all data collected (parameters reported in Table 2). The fit to all the data was not as good as it was for the single event (Table 2 and Figure 3B) due to the overall variability of gaseous N emissions in time and space, although no trend was observed in the change in N2O emissions throughout the growing season. Nonetheless, as it takes into consideration the observed reduction in emissions with time after the fertigation event, the exponential model was a more reliable tool to use in approximating the cumulative N2O emissions from all studied systems than either average values or single values representing the time at which they were taken. Measurements of N2O emissions throughout the growing season can be seen in Figure S1.

Figure 3.

Figure 3

Open in a new tab

(A) N2O emissions, measured over time after an irrigation event, of one lysimeter of each treatment; fit with the exponential model whose parameters and R2 values are listed in Table 2. (B) All of the N2O oxide measurements over time (note that the Y-axis scale changes) with the exponential model fit to each treatment whose parameters and R2 values are listed in Table 2. The axes of panel (B) are limited to 10 μg N m–2 min–1 and 100 min to make it more readable.

Table 2. Number of Emission Measurements of the Total Dataset (n), the Parameters of the Exponential Model’s (eq 1) Fit to the Whole Season Measurements, and the R2 Values for this Model’s Fit to the Emissions Measured after a Single Irrigation Event and the Whole Season Data.

 
 exponential model
 R2
treatment n E0 k single event all data
AN 104 1.77 0.10 0.70 0.48
AN+ 134 7.30 0.04 0.91 0.31
FN+ 103 0.38 0.01 0.98 0.12
Open in a new tab
a

The parameters for the best fit were determined by means of the Solver.

The amount of N2O dissolved in the AN and AN+ water was higher than in the FN+ treatment, which was in equilibrium with the atmospheric concentration (Table 1). Part of the large initial post-irrigation peak in N2O emissions was hypothesized to be due to irrigation water degassing. This was confirmed by measuring the emissions following the irrigation with water of different dissolved N2O concentrations. Figure 4A shows that the N2O emissions, following irrigation with degassed water, were lower and remained stable over time as opposed to water with higher dissolved N2O concentrations with the E0 and k parameters of eq 1 increasing and decreasing, respectively, with increasing irrigation water N2O concentration. The concentration of the degassed water (∼0.2 μg N2O-N L–1) was lower than what was reported to be at equilibrium with the atmospheric concentration (Table 1) because of the warmer water temperature (∼30 °C).

Figure 4.

Figure 4

Open in a new tab

Experiments on lysimeters without plants to test hypotheses concerning N2O emission pathways. (A) N2O emissions over time after irrigation with water of different dissolved N2O concentrations. The exponential model (eq 1) was fit to each dataset with the R2 value of the modeled fit in parentheses after the dissolved N2O (μg L−1) concentration reported in the legend. (B) N2O emissions following fertigation with synthetic fertilizer with and without added carbon (irrigation events marked with red arrows). The emissions over time are shown in dots with the standard deviation reported on the +C treatment (n = 2), and the cumulative emissions are shown with lines whose colors match those of the dots reported in the legend.

The only difference between the AN+ and FN+ treatments was the absence of C in FN+ water, and it was hypothesized that N2O production was C-limited in the FN+ treatment. This was tested by comparing the N2O emissions of lysimeters fertigated with water from FN+ with and without C (Figure 4B). The N2O emissions did not increase after the first irrigation event, but following the second and third events, there was a clear increase similar to Liang et al.60 who demonstrated that C availability can alter the N2O flux in laboratory incubations. Dependence of the N2O flux on the C availability in a mostly aerobic environment suggests that heterotrophic N2O production pathways other than denitrification occur in this system, a point that is discussed in-depth in the Supporting Information.

The cumulative N2O emissions are shown in Figure 5, and it is possible that we slightly underestimated the cumulative N2O emissions due to the model’s (eq 1) failure to represent the sharp reduction of N2O emissions in the first 5 min after irrigation (e.g., AN+; Figure 3A). The increase in rate of N2O emission after DAT 25 is due to the increased amount of irrigation (i.e., irrigation events) on the basis of which the emissions were calculated.

Figure 5.

Figure 5

Open in a new tab

Cumulative N2O emissions calculated by means of the exponential model (eq 1) with the parameters listed in Table 2.

The N2O emissions in this study were lower than previously reported.61,62 The calculated results indicate that the highest sum of emissions was from the AN+ treatment at 62 mg N m–2, while emissions from AN and FN+ were similar (Figure 5). This amounts to 0.4, 4.5, and 0.6 g of N2O-N emission per Mg of cucumbers produced for treatments AN, AN+, and FN+, respectively. The EFs were also very low, only 0.1% for the system with the highest emissions (AN+; Figure 5), compared to an average of 1% assumed for fertilized agriculture.20 However, Hashida et al.63 observed lower N2O emissions of 9 mg m–2 in a tomato cultivation in fresh soil during the first season, which increased to 31 and 290 mg m–2 in the subsequent two half seasons (all data normalized to match the 50 day season length of the experiment described in this paper). They attributed this to increased N substrate availability over time, although an alternative explanation could be the development of microbial films in the soil. Daum and Schenk16 reported an EF of about 0.9% in a cucumber cultivation on rock wool in Holland with an overall 25% lower yield compared to our results. Yao et al.59 reported even higher N2O emissions of 174 g Mg–1 for cucumbers grown in China with similar N application rates. However, their N application type and timing were very different from those reported here. They applied fertilizer in the form of urea or manure in a few large pulses, which potentially reduced the NUE and is reflected in 7 times lower yields, as compared to our study.

Fertigation with NO3– from the RAS effluent (AN), as opposed to a synthetic N source (FN+), prevents 37 kg of CO2e emissions per Mg of fresh cucumbers produced. The reduction in the case of the AN+ was slightly lower at 32 kg per Mg of fresh cucumbers produced due to the increased N2O emissions resulting from the sludge incorporation. As there was no apparent added benefit to the higher NH4+:NO3– ratio in terms of yield, an alternative way to use the sludge for fertigation would be to treat it in a second nitrifying biofilter.64 Such systems have been shown to be capable of converting high NH4+ concentrations to NO3–65 while potentially reducing total N2O emissions.66 Our estimations of the GWI reduction do not include the CO2e emissions avoided by using the RAS effluent, which would otherwise require treatment by a wastewater treatment plant. Such a treatment is estimated to emit between 230 and 830 g of the CO2e m–3 effluent.67 In the IAAS system, 5865 m3 ha–1 was used for irrigation, and inclusion of the CO2e savings associated with the treatment of this water would increase climate change mitigation of the integrated systems by 30–100% (10–35 kg of CO2e emissions per Mg of fresh cucumbers produced). Moreover, fish production and consumption accounts for 3.8% of the annual per capita impact for aquatic eutrophication and is a larger environmental load than the aquaculture contribution to the climate change.68 The reuse of effluent water for horticulture thus will reduce not only the GWI of the integrated system but also its eutrophication potential.

Acknowledgments

We thank Matthias Overduin, Angelika Yurina, and Mais Dzhafarov for help with the field work at the Arava R&D station where the experiments were performed. Financial support for this work was provided by the Israel Ministry of Agriculture and Rural Development, Israel (grant number 3000013059), the Keren Kayemeth Le Israel-Jewish National Fund, Israel, the Goldinger Trust, USA, the Daniel Koshland Fund for Interdisciplinary Research, USA, and the JCA Foundation.

Supporting Information Available

The Supporting Information is available free of charge at https://pubs.acs.org/doi/10.1021/acs.est.0c00869.

  • Supplementary methods, results and discussion all relate to the contribution of different pathways of N2O production to the total N2O emission from the studied systems (Figures S1–S3 and Table S1), and concentrations of nutrients other than nitrogen in water (Table S2) (PDF)

Author Contributions

T.G., I.G., and N.L. designed the experiments, T.G. conducted the experiments and analyzed the data, and T.G. and I.G. wrote the paper with contributions from all co-authors.

The authors declare no competing financial interest.

Supplementary Material

es0c00869_si_001.pdf (303.9KB, pdf)

References

  1. Tilman D.; Balzer C.; Hill J.; Befort B. L. Global Food Demand and the Sustainable Intensification of Agriculture. Proc. Natl. Acad. Sci. U. S. A. 2011, 108, 20260–20264. 10.1073/pnas.1116437108. [DOI] [PMC free article] [PubMed] [Google Scholar]
  2. Eshel G.; Shepon A.; Noor E.; Milo R. Environmentally Optimal, Nutritionally Aware Beef Replacement Plant-Based Diets. Environ. Sci. Technol. 2016, 50, 8164–8168. 10.1021/acs.est.6b01006. [DOI] [PubMed] [Google Scholar]
  3. Garnett E. E.; Balmford A.; Sandbrook C.; Pilling M. A.; Marteau T. M. Impact of Increasing Vegetarian Availability on Meal Selection and Sales in Cafeterias. Proc. Natl. Acad. Sci. 2019, 116, 20923. 10.1073/pnas.1907207116. [DOI] [PMC free article] [PubMed] [Google Scholar]
  4. FAO FAOSTAT; http://www.fao.org/faostat/en/#data/QC (accessed Feb 10, 2020).
  5. Suhl J.; Dannehl D.; Kloas W.; Baganz D.; Jobs S.; Scheibe G.; Schmidt U. Advanced Aquaponics: Evaluation of Intensive Tomato Production in Aquaponics vs. Conventional Hydroponics. Agric. Water Manag. 2016, 178, 335–344. 10.1016/j.agwat.2016.10.013. [DOI] [Google Scholar]
  6. Love D. C.; Fry J. P.; Li X.; Hill E. S.; Genello L.; Semmens K.; Thompson R. E. Commercial Aquaponics Production and Profitability: Findings from an International Survey. Aquaculture 2015, 435, 67–74. 10.1016/j.aquaculture.2014.09.023. [DOI] [Google Scholar]
  7. Timmons M. B.; Ebeling J. M.. Recirculating Aquaculture; Ithaca Publishing Company: 2013. [Google Scholar]
  8. Erisman J. W.; Bleeker A.; Galloway J.; Sutton M. S. Reduced Nitrogen in Ecology and the Environment. Environ. Pollut. 2007, 150, 140–149. 10.1016/j.envpol.2007.06.033. [DOI] [PubMed] [Google Scholar]
  9. Dodds W. K.; Bouska W. W.; Eitzmann J. L.; Pilger T. J.; Pitts K. L.; Riley A. J.; Schloesser J. T.; Thornbrugh D. J. Eutrophication of U.S. Freshwaters: Analysis of Potential Economic Damages. Environ. Sci. Technol. 2009, 43, 12. 10.1021/es801217q. [DOI] [PubMed] [Google Scholar]
  10. FAO The State of World Fisheries and Aquaculture 2016; Rome: 2016. [Google Scholar]
  11. Groenveld T.; Kohn Y. Y.; Gross A.; Lazarovitch N. Optimization of Nitrogen Use Efficiency by Means of Fertigation Management in an Integrated Aquaculture-Agriculture System. J. Cleaner Prod. 2019, 212, 401–408. 10.1016/j.jclepro.2018.12.031. [DOI] [Google Scholar]
  12. van Rijn J. The Potential for Integrated Biological Treatment Systems in Recirculating Fish Culture—A Review. Aquaculture 1996, 139, 181–201. 10.1016/0044-8486(95)01151-X. [DOI] [Google Scholar]
  13. van Rijn J. Waste Treatment in Recirculating Aquaculture Systems. Aquac. Eng. 2013, 53, 49–56. 10.1016/j.aquaeng.2012.11.010. [DOI] [Google Scholar]
  14. Wongkiew S.; Hu Z.; Chandran K.; Lee J. W.; Khanal S. K. Nitrogen Transformations in Aquaponic Systems: A Review. Aquac. Eng. 2017, 76, 9–19. 10.1016/j.aquaeng.2017.01.004. [DOI] [Google Scholar]
  15. Smil V.Enriching the Earth; The MIT Press: Cambridge, Massachusetts, 2001. [Google Scholar]
  16. Daum D.; Schenk M. K. Gaseous Nitrogen Losses from a Soilless Culture System in the Greenhouse. Plant Soil 1996, 183, 69–78. 10.1007/BF02185566. [DOI] [Google Scholar]
  17. Cameron K. C.; Di H. J.; Moir J. L. Nitrogen Losses from the Soil/Plant System: A Review. Ann. Appl. Biol. 2013, 162, 145–173. 10.1111/aab.12014. [DOI] [Google Scholar]
  18. Crutzen P. J. The Role of NO and NO2 in the Chemistry of the Troposhere and Stratosphere. Annu. Rev. Earth Planet. Sci. 1979, 7, 443–472. 10.1146/annurev.ea.07.050179.002303. [DOI] [Google Scholar]
  19. Tian H.; Yang J.; Xu R.; Lu C.; Canadell J. G.; Davidson E. A.; Jackson R. B.; Arneth A.; Chang J.; Ciais P.; Gerber S.; Ito A.; Joos F.; Lienert S.; Messina P.; Olin S.; Pan S.; Peng C.; Saikawa E.; Thompson R. L.; Vuichard N.; Winiwarter W.; Zaehle S.; Zhang B. Global Soil Nitrous Oxide Emissions since the Preindustrial Era Estimated by an Ensemble of Terrestrial Biosphere Models: Magnitude, Attribution, and Uncertainty. Glob. Chang. Biol. 2019, 25, 640–659. 10.1111/gcb.14514. [DOI] [PubMed] [Google Scholar]
  20. Reay D. S.; Davidson E. A.; Smith K. A.; Smith P.; Melillo J. M.; Dentener F.; Crutzen P. J. Global Agriculture and Nitrous Oxide Emissions. Nat. Clim. Chang. 2012, 2, 410–416. 10.1038/nclimate1458. [DOI] [Google Scholar]
  21. Robertson G. P.; Vitousek P. M. Nitrogen in Agriculture: Balancing the Cost of an Essential Resource. Annu. Rev. Environ. Resour. 2009, 34, 97–125. 10.1146/annurev.environ.032108.105046. [DOI] [Google Scholar]
  22. Robertson G. P.; Groffman P. M.. Nitrogen Transformations. In Soil Microbiology, Ecology and Biochemistry; Paul E. A., Ed.; Academic Press: Burlington, Massachusetts, USA, 2015; pp 421–446, 10.1016/b978-0-12-415955-6.00014-1. [DOI] [Google Scholar]
  23. Hu Z.; Lee J. W.; Chandran K.; Kim S.; Khanal S. K. Nitrous Oxide (N2O) Emission from Aquaculture : A Review. Environ. Sci. Technol. 2012, 46, 6470–6480. 10.1021/es300110x. [DOI] [PubMed] [Google Scholar]
  24. Butterbach-Bahl K.; Baggs E. M.; Dannenmann M.; Kiese R.; Zechmeister-Boltenstern S. Nitrous Oxide Emissions from Soils: How Well Do We Understand the Processes and Their Controls?. Philos. Trans. R. Soc. Lond. B. Biol. Sci. 2013, 368, 20130122. 10.1098/rstb.2013.0122. [DOI] [PMC free article] [PubMed] [Google Scholar]
  25. Müller C.; Laughlin R. J.; Spott O.; Rütting T. Quantification of N2O Emission Pathways via a 15N Tracing Model. Soil Biol. Biochem. 2014, 72, 44–54. 10.1016/j.soilbio.2014.01.013. [DOI] [Google Scholar]
  26. Kool D. M.; Dolfing J.; Wrage N.; Van Groenigen J. W. Nitrifier Denitrification as a Distinct and Significant Source of Nitrous Oxide from Soil. Soil Biol. Biochem. 2011, 43, 174–178. 10.1016/j.soilbio.2010.09.030. [DOI] [Google Scholar]
  27. Law Y.; Ye L.; Pan Y.; Yuan Z. Nitrous Oxide Emissions from Wastewater Treatment Processes. Philos. Trans. R. Soc. Lond. B. Biol. Sci. 2012, 367, 1265–1277. 10.1098/rstb.2011.0317. [DOI] [PMC free article] [PubMed] [Google Scholar]
  28. Baggs E. M. Soil Microbial Sources of Nitrous Oxide: Recent Advances in Knowledge, Emerging Challenges and Future Direction. Curr. Opin. Environ. Sustain. 2011, 3, 321–327. 10.1016/j.cosust.2011.08.011. [DOI] [Google Scholar]
  29. Prosser J. I.; Hink L.; Gubry-Rangin C.; Nicol G. W. Nitrous Oxide Production by Ammonia Oxidisers: Physiological Diversity, Niche Differentiation and Potential Mitigation Strategies. Glob. Chang. Biol 2020, 26, 103. 10.1111/gcb.14877. [DOI] [PubMed] [Google Scholar]
  30. Kampschreur M. J.; Temmink H.; Kleerebezem R.; Jetten M. S. M.; van Loosdrecht M. C. M. Nitrous Oxide Emission during Wastewater Treatment. Water Res. 2009, 43, 4093–4103. 10.1016/j.watres.2009.03.001. [DOI] [PubMed] [Google Scholar]
  31. Guo J.; Peng Y.; Wang S.; Ma B.; Ge S.; Wang Z.; Huang H.; Zhang J.; Zhang L. Pathways and Organisms Involved in Ammonia Oxidation and Nitrous Oxide Emission. Crit. Rev. Environ. Sci. Technol. 2013, 43, 2213–2296. 10.1080/10643389.2012.672072. [DOI] [Google Scholar]
  32. Ji B.; Yang K.; Zhu L.; Jiang Y.; Wang H.; Zhou J.; Zhang H. Aerobic Denitrification: A Review of Important Advances of the Last 30 Years. Biotechnol. Bioprocess Eng. 2015, 20, 643–651. 10.1007/s12257-015-0009-0. [DOI] [Google Scholar]
  33. Müller C.; Clough T. J. Advances in Understanding Nitrogen Flows and Transformations: Gaps and Research Pathways. J. Agric. Sci. 2014, 152, 34–44. 10.1017/S0021859613000610. [DOI] [Google Scholar]
  34. Van Kessel M. A. H. J.; Harhangi H. R.; Flik G.; Jetten M. S. M.; Klaren P. H. M.; Op Den Camp H. J. M. Anammox Bacteria in Different Compartments of Recirculating Aquaculture Systems. Biochem. Soc. Trans. 2011, 39, 1817–1821. 10.1042/BST20110743. [DOI] [PubMed] [Google Scholar]
  35. Van Duc L.; Song B.; Ito H.; Hama T.; Otani M.; Kawagoshi Y. High Growth Potential and Nitrogen Removal Performance of Marine Anammox Bacteria in Shrimp-Aquaculture Sediment. Chemosphere 2018, 196, 69–77. 10.1016/j.chemosphere.2017.12.159. [DOI] [PubMed] [Google Scholar]
  36. Shcherbak I.; Millar N.; Robertson G. P. Global Metaanalysis of the Nonlinear Response of Soil Nitrous Oxide (N2O) Emissions to Fertilizer Nitrogen. Proc. Natl. Acad. Sci. 2014, 111, 9199–9204. 10.1073/pnas.1322434111. [DOI] [PMC free article] [PubMed] [Google Scholar]
  37. Ahn J. H.; Kim S.; Park H.; Rahm B.; Pagilla K.; Chandran K. N2O Emissions from Activated Sludge Processes, 2008-2009: Results of a National Monitoring Survey in the United States. Environ. Sci. Technol. 2010, 44, 4505–4511. 10.1021/es903845y. [DOI] [PubMed] [Google Scholar]
  38. UNEP Drawing Down N2O To Protect Climate and the Ozone Layer; United Nations Environment Programme: Nairobi, Kenya, 2013, 978–92–807-3358-7. [Google Scholar]
  39. Liu S.; Hu Z.; Wu S.; Li S.; Li Z.; Zou J. Methane and Nitrous Oxide Emissions Reduced Following Conversion of Rice Paddies to Inland Crab-Fish Aquaculture in Southeast China. Environ. Sci. Technol. 2016, 50, 633–642. 10.1021/acs.est.5b04343. [DOI] [PubMed] [Google Scholar]
  40. Paudel S. R.; Choi O.; Khanal S. K.; Chandran K.; Kim S.; Lee J. W. Effects of Temperature on Nitrous Oxide (N2O) Emission from Intensive Aquaculture System. Sci. Total Environ. 2015, 518-519, 16–23. 10.1016/j.scitotenv.2015.02.076. [DOI] [PubMed] [Google Scholar]
  41. Ma Y.; Sun L.; Liu C.; Yang X.; Zhou W.; Yang B.; Schwenke G.; Liu D. L. A Comparison of Methane and Nitrous Oxide Emissions from Inland Mixed-Fish and Crab Aquaculture Ponds. Sci. Total Environ. 2018, 637-638, 517–523. 10.1016/j.scitotenv.2018.05.040. [DOI] [PubMed] [Google Scholar]
  42. Tyson R. V.; Simonne E. H.; Treadwell D. D.; White J. M.; Simonne A. Reconciling PH for Ammonia Biofiltration and Cucumber Yield in a Recirculating Aquaponic System with Perlite Biofilters. Hort Science 2008, 43, 719–724. 10.21273/HORTSCI.43.3.719. [DOI] [Google Scholar]
  43. Seawright D. E.; Stickney R. R.; Walker R. B. Nutrient Dynamics in Integrated Aquaculture-Hydroponics Systems. Aquaculture 1998, 160, 215–237. 10.1016/S0044-8486(97)00168-3. [DOI] [Google Scholar]
  44. Yogev U.; Barnes A.; Gross A. Nutrients and Energy Balance Analysis for a Conceptual Model of a Three Loops off Grid, Aquaponics. Water 2016, 8, 589. 10.3390/w8120589. [DOI] [Google Scholar]
  45. Gelfand I.; Cui M.; Tang J.; Robertson G. P. Short-Term Drought Response of N2O and CO2 Emissions from Mesic Agricultural Soils in the US Midwest. Agric. Ecosyst. Environ. 2015, 212, 127–133. 10.1016/j.agee.2015.07.005. [DOI] [Google Scholar]
  46. Gelfand I.; Shcherbak I.; Millar N.; Kravchenko A. N.; Robertson G. P. Long-Term Nitrous Oxide Fluxes in Annual and Perennial Agricultural and Unmanaged Ecosystems in the Upper Midwest USA. Glob. Chang. Biol. 2016, 22, 3594–3607. 10.1111/gcb.13426. [DOI] [PubMed] [Google Scholar]
  47. Weiss R. F.; Price B. A. Nitrous Oxide Solubility in Water and Seawater. Mar. Chem. 1980, 8, 347–359. 10.1016/0304-4203(80)90024-9. [DOI] [Google Scholar]
  48. Hoben J. P.; Gehl R. J.; Millar N.; Grace P. R.; Robertson G. P. Nonlinear Nitrous Oxide (N2O) Response to Nitrogen Fertilizer in on-Farm Corn Crops of the US Midwest. Glob. Chang. Biol. 2011, 17, 1140–1152. 10.1111/j.1365-2486.2010.02349.x. [DOI] [Google Scholar]
  49. Gelfand I.; Robertson G. P.. Mitigation of Greenhouse Gases in Agricultural Ecosystems. In The Ecology of Agricultural Landscapes: Long-Term Research on the Path to Sustainability; Hamilton S. K., Doll J. E., Robertson P. G., Eds.; Oxford University Press: New York, NY, 2015; p 310. [Google Scholar]
  50. Levasseur A.; Brandão M.; Lesage P.; Margni M.; Pennington D.; Clift R.; Samson R. Valuing Temporary Carbon Storage. Nat. Clim. Chang. 2012, 2, 6–8. 10.1038/nclimate1335. [DOI] [Google Scholar]
  51. Kool A.; Marinussen M.; Blonk H.. GHG Emissions of N, P and K Fertilizer Production. In LCI Data for the Calculation Tool Feedprint for Greenhouse Gas Emissions of Feed Production and Util; Van Zeist W. J., Marinussen M., Broekema R., Groen E., Kool a., Dolman M.; Blonk H.. (2012). Gouda, 2012. [Google Scholar]
  52. van Groenigen J. W.; Velthof G. L.; Oenema O.; Van Groenigen K. J.; Van Kessel C. Towards an Agronomic Assessment of N2O Emissions: A Case Study for Arable Crops. Eur. J. Soil Sci. 2010, 61, 903–913. 10.1111/j.1365-2389.2009.01217.x. [DOI] [Google Scholar]
  53. Martins C. I. M.; Eding E. H.; Verdegem M. C. J.; Heinsbroek L. T. N.; Schneider O.; Blancheton J. P.; Orbcastel E. R.; Verreth J. a. J. New Developments in Recirculating Aquaculture Systems in Europe: A Perspective on Environmental Sustainability. Aquac. Eng. 2010, 43, 83–93. 10.1016/j.aquaeng.2010.09.002. [DOI] [Google Scholar]
  54. d’Orbcastel E. R.; Blancheton J.-P.; Aubin J. Towards Environmentally Sustainable Aquaculture: Comparison between Two Trout Farming Systems Using Life Cycle Assessment. Aquac. Eng. 2009, 40, 113–119. 10.1016/j.aquaeng.2008.12.002. [DOI] [Google Scholar]
  55. Lane D.Online Statistics Education: A Multimedia Course of Study http://onlinestatbook.com/.
  56. Pezullo J. StatPages https://statpages.info/.
  57. Sui X.; Shan N.; Hu L.; Zhang C.; Yu C.; Ren H.; Turgeon R.; Zhang Z. The Complex Character of Photosynthesis in Cucumber Fruit. J. Exp. Bot. 2017, 68, 1625–1637. 10.1093/jxb/erx034. [DOI] [PMC free article] [PubMed] [Google Scholar]
  58. Grewal H. S.; Maheshwari B.; Parks S. E. Water and Nutrient Use Efficiency of a Low-Cost Hydroponic Greenhouse for a Cucumber Crop: An Australian Case Study. Agric. Water Manag. 2011, 98, 841–846. 10.1016/j.agwat.2010.12.010. [DOI] [Google Scholar]
  59. Yao Z.; Yan G.; Wang R.; Zheng X.; Liu C.; Butterbach-Bahl K. Drip Irrigation or Reduced N-Fertilizer Rate Can Mitigate the High Annual N2O+NO Fluxes from Chinese Intensive Greenhouse Vegetable Systems. Atmos. Environ 2019, 212, 183–193. 10.1016/j.atmosenv.2019.05.056. [DOI] [Google Scholar]
  60. Liang L. L.; Eberwein J. R.; Allsman L. a.; Grantz D. a.; Jenerette G. D. Regulation of CO2 and N2O Fluxes by Coupled Carbon and Nitrogen Availability. Environ. Res. Lett. 2015, 10, 034008 10.1088/1748-9326/10/3/034008. [DOI] [Google Scholar]
  61. Zou Y.; Hu Z.; Zhang J.; Fang Y.; Li M.; Zhang J. Mitigation of N2O Emission from Aquaponics by Optimizing the Nitrogen Transformation Process: Aeration Management and Exogenous Carbon (PLA) Addition. J. Agric. Food Chem. 2017, 65, 8806–8812. 10.1021/acs.jafc.7b03211. [DOI] [PubMed] [Google Scholar]
  62. Zou Y.; Hu Z.; Zhang J.; Xie H.; Liang S.; Wang J.; Yan R. Attempts to Improve Nitrogen Utilization Efficiency of Aquaponics through Nitrifies Addition and Filler Gradation. Environ. Sci. Pollut. Res. 2016, 23, 6671–6679. 10.1007/s11356-015-5898-0. [DOI] [PubMed] [Google Scholar]
  63. Hashida S. N.; Johkan M.; Kitazaki K.; Shoji K.; Goto F.; Yoshihara T. Management of Nitrogen Fertilizer Application, Rather than Functional Gene Abundance, Governs Nitrous Oxide Fluxes in Hydroponics with Rockwool. Plant Soil 2014, 374, 715–725. 10.1007/s11104-013-1917-4. [DOI] [Google Scholar]
  64. Gross A.; Guy O.; Posmanik R.; Fine P.; Nejidat A. A Novel Method for Combined Biowaste Stabilization and Production of Nitrate-Rich Liquid Fertilizer for Use in Organic Horticulture. Water, Air, Soil Pollut. 2012, 223, 1205–1214. 10.1007/s11270-011-0938-y. [DOI] [Google Scholar]
  65. Posmanik R.; Gross A.; Nejidat A. Effect of High Ammonia Loads Emitted from Poultry-Manure Digestion on Nitrification Activity and Nitrifier-Community Structure in a Compost Biofilter. Ecol. Eng. 2014, 62, 140–147. 10.1016/j.ecoleng.2013.10.033. [DOI] [Google Scholar]
  66. Posmanik R.; Nejidat A.; Bar-Sinay B.; Gross A. Integrated Biological Treatment of Fowl Manure for Nitrogen Recovery and Reuse. J. Environ. Manage. 2013, 117, 172–179. 10.1016/j.jenvman.2012.12.049. [DOI] [PubMed] [Google Scholar]
  67. Garrido-Baserba M.; Hospido A.; Reif R.; Molinos-Senante M.; Comas J.; Poch M. Including the Environmental Criteria When Selecting a Wastewater Treatment Plant. Environ. Model. Softw. 2014, 56, 74–82. 10.1016/j.envsoft.2013.11.008. [DOI] [Google Scholar]
  68. Bohnes F. A.; Hauschild M. Z.; Schlundt J.; Laurent A. Life Cycle Assessments of Aquaculture Systems: A Critical Review of Reported Findings with Recommendations for Policy and System Development. Rev. Aquac. 2019, 11, 1061–1079. 10.1111/raq.12280. [DOI] [Google Scholar]

Associated Data

This section collects any data citations, data availability statements, or supplementary materials included in this article.

Supplementary Materials

es0c00869_si_001.pdf (303.9KB, pdf)

Articles from Environmental Science & Technology are provided here courtesy of American Chemical Society

RESOURCES