Nitrogen combined with biochar changed the feedback mechanism between soil nitrification and Cd availability in an acidic soil

Journal Pre-proof Nitrogen combined with biochar changed the feedback mechanism between soil nitrification and Cd availability in an acidic soil Haochun Zhao, Lu Yu, Mengjie Yu, Muhammad Afzal, Zhongming Dai, Philip Brookes, Jianming Xu

PII:

S0304-3894(19)31585-7

DOI:

https://doi.org/10.1016/j.jhazmat.2019.121631

Reference:

HAZMAT 121631

To appear in:

Journal of Hazardous Materials

Received Date:

3 May 2019

Revised Date:

21 October 2019

Accepted Date:

5 November 2019

Please cite this article as: Zhao H, Yu L, Yu M, Afzal M, Dai Z, Brookes P, Xu J, Nitrogen combined with biochar changed the feedback mechanism between soil nitrification and Cd availability in an acidic soil, Journal of Hazardous Materials (2019), doi: https://doi.org/10.1016/j.jhazmat.2019.121631

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Nitrogen combined with biochar changed the feedback mechanism between soil nitrification and Cd availability in an acidic soil

Haochun Zhaoa,b, Lu Yua,b, Mengjie Yua,b, Muhammad Afzala,b, Zhongming Daia,b, Philip Brookesa,b, Jianming Xua,b* a

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Institute of Soil and Water Resources and Environmental Science, College of Environmental

and Resource Sciences, Zhejiang University, 866 Yuhangtang Road, Hangzhou 310058, China

Zhejiang Provincial Key Laboratory of Agricultural Resources and Environment, Zhejiang

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University, 866 Yuhangtang Road, Hangzhou 310058, China

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* Corresponding author

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Tel & Fax: +86 571 8898 2069

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E-mail address: [email protected]

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Graphical abstract

Highlights

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Mutual responses between soil nitrification and Cd are firstly investigated N inputs did not cause soil acidification and Cd mobilization in Cd-contaminated soil Cd inhibited soil nitrification through eliminating AOB Biochar promoted AOB recovery and restored nitrification of Cd-contaminated soil Biochar could resist the re-acidification and Cd mobilization.

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 

Abstract

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Inorganic nitrogen (N) inputs increase soil nitrification, acidification and trace metal toxicity e.g. cadmium (Cd). Biochar (B) has been widely used for metal immobilization.

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However, little is known about how the combination of N fertilizers with biochar (N-

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B) changes soil Cd availability through altering nitrification process. Here, (NH4)2SO4 or CO(NH2)2 was applied in combination with biochar to an acidic, artificially enriched Cd contaminated soil. Not as we expected, available Cd did not increase following (NH4)2SO4 or CO(NH2)2 addition. Nitrification and acidification of Cd contaminated soils were greatly inhibited, accompanied by elimination of ammonia-oxidizing 2 / 45

bacteria (AOB). Exchangeable H+ of Cd contaminated soils was significantly lower than that of uncontaminated soils, thus inhibiting Cd itself from mobilization. N-B addition nearly halved soil available Cd and significantly increased nitrification by promoting AOB recovery. However, the restored nitrification did not cause soil acidification, due to the high buffering and slow liming effects of biochar. Available Cd continuously decreased with decreasing soil acidity and exchangeable Al. This

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study firstly demonstrated a feedback between soil nitrification and Cd after N

application, and how biochar modified the feedback. Biochar, therefore, provides a

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feasible strategy for eliminating potential Cd toxicity on both soil biological and chemical processes.

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Key words: Nitrification; Cd; Negative feedback; Functional gene; Biochar

1. Introduction

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remediation.

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Since the industrial revolution, human activities have greatly increased soil heavy metal contamination [1]. Unlike organic pollutants, heavy metals are not degraded but

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accumulate in soils and sediments, threatening crop safety and human health through

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bioaccumulation from lower trophic levels into higher ones [2]. Among the heavy metals, pollution associated with cadmium (Cd) continues to be a public health issue because of its high toxicity, wide distribution and persistence [1]. Cd can cause kidney damage, chondral symptoms and cardiovascular disease in humans [3]. Outbreaks of Itai-itai disease caused hundreds of deaths in Japan in the last century [4]. 3 / 45

Soil Cd bioavailability is affected by soil properties including soil pH, organic matter, exchangeable cations, texture and nutrients [5-7], of which soil pH is the most important factor [5, 6]. Many studies have demonstrated that the solubility of Cd greatly increases as soil pH decreases, and the hazard potential is significantly higher in acidic soils [5, 6]. Acidic soils (defined as pH < 5.5) occupy 30% of the world's ice-free land [8]. Most are highly weathered and poorly buffered [9], and thus are more susceptible

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to acidification [10] and metal mobilization. In China, this area of cropland is expanding

rapidly, with a corresponding excessive use of inorganic nitrogen (N) fertilization [11].

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Nitrogen fertilizers, especially ammonium fertilizers, significantly accelerate soil acidification through (i) providing NH4+ directly to promote soil nitrification [12] (ii)

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and indirectly stimulating plant roots to assimilate NH4+, which then exude protons (H+)

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to maintain an internal charge balance [13]. As a consequence, increasing soil acidity accompanied with Cd toxicity may impact soil microbial community structure,

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transcription, functions such as C and N cycling [14, 15], threatening soil flora and fauna, animals and also human beings [16]. Nitrifiers such as ammonia oxidizing

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bacteria (AOB) and archaea (AOA) drive nitrification, playing an important role in N

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cycling and are sensitive to heavy metals [15, 17]. Therefore, it is possible that the niche separation and functions of nitrifiers may be altered by both indirect acidity stress and Cd toxicity.

Previous studies determining how N fertilization affects Cd bioavailability have mostly focused on plant-soil systems and demonstrated that ammonium-based 4 / 45

fertilizers accelerate soil acidification and Cd mobilization [18-20]. However, these studies have not clarified how nitrification affects Cd bioavailability, because of the intricate plant-soil-metal interactions in the rhizosphere [21]. Firstly, it is unclear how much the nitrification process responsible for rhizosphere acidification due to its complexity [22]. For instance, plants excrete not only organic acids, but also other chemicals such as biological nitrification inhibitors (BNI), which can significantly

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depress soil nitrification [23], and in turn, promote N use efficiency [24, 25]. It is also uncertain how soil acidification contributes to Cd transformations in plant-soil systems.

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In addition to acidification, plants and microbes also influence Cd phytoextraction and mobility through absorption, chelation and precipitation [21], etc. Therefore, the precise

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role of nitrification is uncertain in the process of Cd mobilization. Even in single soil

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system, this issue has hardly been clarified, since these work mostly concerned only the chemical process of how soil nitrification effects metal availability, or conversely, the

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biological process of how trace metals influence soil nitrification. To date, no studies have been simultaneously investigated the interactions between the two processes that

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are involved. Thus, the mechanisms whereby diverse N fertilization affects Cd

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availability need to be further discussed, particularly in acidic soils. Biochar has recently received increasing interest in the remediation of heavy metal

contaminated soils due to its unique physical, chemical and biological properties [26]. With a porous structure and high absorption capacity, biochar is considered especially effective in metal absorption [27]. Biochar can also ameliorate soil acidity and increase 5 / 45

soil organic matter content and cation exchange capacity (CEC), thus decreasing Cd availability [28]. Biochar also changes other soil characteristics including increased nutrient availability, water retention, aerobic conditions and habitats

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microorganisms [26]. These changes may influence microbial diversity, activity and their functioning [26, 29]. However, when biochar is combined with different types of N fertilizer (B-N), the remediation effects in Cd contaminated soils are poorly

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understood.

Our aims were to evaluate the risks of Cd pollution in acidic soils amended with

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different inorganic N fertilizers, reveal the interactions between soil nitrification and

Cd, and provide a feasible strategy for eliminating the potential risks of Cd. An

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incubation experiment was conducted to determine the individual and combined effects

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of two types of N fertilizer and biochar on Cd bioavailability in an acidic soil. Specifically, we focused on three main aspects: (i) compare the changes in soil Cd

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availability and soil properties in different treatments, (ii) evaluate the interrelationships between soil nitrification, acidification and Cd mobilization, (iii) determine how Cd

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altered the niches and functions of soil nitrifiers, and how they recovered with and

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without biochar addition.

2. Materials and methods 2.1. Soil sampling Soil samples were collected from the surface layer (0-15 cm) of a bareland area, which is located at Quzhou (E 119.17°, N 29.03°), Zhejiang, China. Quzhou has a 6 / 45

subtropical climate, with an average annual temperature of 17.6 ℃ and annual average precipitation of 1476.5 mm. The soil is classified as a Ferrosol (WRB, 1998). After removing the plants, stones and earthworms, the soil samples were dried and sieved < 2 mm for incubation experiments. The soil properties were: pH (1:2.5 H2O): 4.4; organic C: 3.2 g kg-1; total N: 0.3 g kg-1; available P: 0.3 g kg-1; effective cation exchange capacity (ECEC): 6.27 cmol kg-1; exchangeable base cations (EB): 0.60 cmol

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kg-1; exchangeable acidity (EA): 5.67 cmol kg-1; exchangeable Ca2+, Mg2+, K+ and Na+:

0.34, 0.12, 0.12 and 0.02 cmol kg-1, respectively. Soil exchangeable Al3+ and H+: 5.10

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and 0.57 cmol kg-1. Base saturation: 9.6%. 2.2. Experimental design

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An aqueous solution of CdCl2 2.5H2O was added to the soils for the preparation

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of Cd contaminated soil samples (5mg Cd kg-1dry soil). After mixing evenly, the contaminated soils were incubated at 25℃ in a dark room for 30 days to permit

comparison.

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chemical and biological equilibrium. Soil samples without Cd contamination served for

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Two treatments were set up in the incubation experiment, one set was a single N

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fertilizer group (N group), to explore the interaction between soil nitrification and Cd availability; another set was the combination group of N fertilizer and biochar (N-B group), to explore how the combination affected soil Cd availability. Three N treatments were applied in each group: Control, ammonium sulfate ((NH4)2SO4), urea (CO(NH2)2). The N fertilizers were applied in solution (100 mg N kg-1 dry soil, equivalent to 240 kg 7 / 45

N ha-1), and biochar was added directly and mixed thoroughly with soil at 3% by weight. Biochar used in this study was generated from wheat straw pyrolyzed at 500℃ (Sanli Company, Shandong Province). The characteristics of the biochar are described previously [30]. Each treatment had three replicates, and each replicate contained 500g dry soil. These fertilizer-biochar-soil mixture samples were transferred to plastic pots and incubated at 25℃ in darkness in a randomized design.

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During the incubation, moisture loss was compensated by adding MiliQ water to maintain a constant soil water content (45% of Water Holding Capacity) every day. Soil

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samples were taken at 0, 3, 7, 14, 28, 56, 77 days after the incubation. Fresh soil samples were analyzed immediately for mineral N after sampling. Some soil samples were air-

2.3. Soil chemical analysis

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extraction and functional gene analysis.

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dried and sieved < 2mm for soil chemical analysis, some were stored at -80℃ for DNA

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Soil available Cd was extracted by 0.01M CaCl2 (1 soil:10 CaCl2 ) then measured with a Graphite Furnace Atomic Absorbance Spectrometer (AAnalyst 800,

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PerkinElmer, USA). Briefly, according to the protocols described in the Agricultural

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Chemistry Committee of China [31], soil pH was measured with a pH electrode (1: 2.5H2O). Soil total organic carbon (TOC) and total nitrogen (TN) concentrations (airdried, milled < 200 μm) were determined by dry combustion using an organic carbon auto-analyzer (Analytic Jena multi N/C 3100, Jena, Germany). Soil dissolved organic carbon (DOC) and nitrogen (DON) were extracted with water then analyzed by a Multi 8 / 45

N/C TOC analyzer (Analytic Jena AG, Jena, Germany). Soil exchangeable base cations (EB) including K, Ca, Na and Mg in soil were extracted by NH4OAC, then analyzed using atomic absorption spectrophotometry (Analytik Jena novAA 300, Jena, Germany). Soil exchangeable acidity (EA), comprising Al and H, were extracted by KCl and measured by titration. The soil cation exchange capacity (ECEC) was the amount of EB and EA, and base saturation was calculated from percentage EB divided

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by ECEC. 2.4. Soil mineral N analysis

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Five gram portions of fresh soil were mixed with 25 mL 1 M KCl and shaken for

1 h. After centrifugation, the supernatants were filtered through 0.45μm pore-size nylon

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membrane filters. The solutions were analyzed for NH4+-N and NO3--N concentrations

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with a continuous colorimeric flow analyzer (San++, Skalar, Netherlands). The net nitrification rate, which reflects soil nitrification activity during a specific period of

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time, was calculated from: [(the difference between final and initial NO3--N concentrations)/ (incubation time)] [32]. Because ammonium supply is the dominant

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factor that influences soil nitrification, the ammonium consumption rate can also reflect

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nitrification activity [33]. It was calculated from [(final NH4+-N) minus (initial NH4+N concentration)]/(incubation time) [32]. 2.5. amoA gene quantitative PCR Firstly, the total soil genomic DNA was extracted from 0.5 g soil samples using FastDNATM Spin Kit (MP Biomedicals, LLC) according to the instruction manual. 9 / 45

Concentrations and quality of extracted DNA were measured using a NanoDrop® ND1000 spectrophotometer (NanoDrop Technologies, Montchanin, USA). The isolated DNA samples were stored at -20℃ for subsequent qPCR analysis. We amplified regions of amoA gene from nitrifiers, the bacterial and archaeal amoA

genes

were

amplified

(GGGGTTTCTACTGGTGGT)/amoA2R

primer

pairs

amoA1F

(CCCCTCKGSAAAGCCTTCTTC)

and

(STAATGGTCTGGCTTAGACG)/Arch-

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Arch-amoAF

by

amoAR(GCGGCCATCCATCTGTATGT), respectively. A typical 20 μL reaction

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consisted of 10 μL of SYBR®Premix Ex TaqTM (TaKaRa, Dalian, China), 0.2 μL of each primer, 1 μL of template DNA and nuclease-free DI water to the final volume. All

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reactions were carried out using the Light Cycler 480 (Roche Applied Science), Melting

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curve analysis was performed at the end of each real-time PCR run to confirm PCR product specificity by measuring fluorescence continuously with the temperature

by Di et al [34].

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increasing from 50 to 99 °C. A real-time PCR standard curve was prepared as described

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2.6. Statistical analysis

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Statistical analysis was performed using SPSS 18.0 for Windows (SPSS Inc., IL, USA). Analysis of variance (ANOVA) was performed to determine the effects of nitrogen and biochar on Cd availability, soil mineral N and soil properties, followed by Tukey HSD to check for significant differences between different treatments. The Pvalue of < 0.05 was used to denote statistical significance. Simple correlation analysis 10 / 45

was conducted by R software. Figures and Tables were prepared by Origin 9.1 and R software. 3. Results 3.1 Soil Cd availability After rapid fluctuations in the early stages, the CaCl2-extractable Cd (termed soil available Cd) of contaminated soils gradually decreased and finally reached a dynamic

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equilibrium (Fig. 1). At day 77, soil available Cd concentrations in the individual N

treatments were in the order: (NH4)2SO4 > Control > CO(NH2)2. Compared with the

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Control, available Cd in the (NH4)2SO4 treatment was not obviously increased, while the CO(NH2)2 treatment had a significantly lower value (Fig. 1). Soil pH in the

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individual N treatments had the opposite order of available Cd (Fig. 4a). When

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combined with biochar, the available Cd decreased by 40% - 47% compared with the single N treatments. In the N-B treatments, the CO(NH2)2 + B treatment had the lowest

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available Cd and highest soil pH, while there were no significant differences between the other two treatments (Fig. 1). No Cd was detected in the uncontaminated soils

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during the incubation. Simple correlation analysis showed soil pH strongly regulated

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soil available Cd in both the N group (r = -0.58, p < 0.05) and the N-B group (r = -0.83, p < 0.001) (Fig. 6). 3.2 Soil nitrification and nitrifiers The maximum NH4+-N concentrations occurred in all treatments within two weeks after fertilization (Fig. 2a), then declined gradually (the ammonia consumption rate was 11 / 45

calculated from day 14) (Fig. 2a). In contrast, NO3−-N concentrations increased continuously over time (Fig. 2b), indicating that nitrification was occurring. In the single N group, Cd contaminated soils had significantly higher concentrations of NH4+-N and lower concentration of NO3−-N than uncontaminated ones. In uncontaminated soils, the (NH4)2SO4 treatment had the highest net nitrification rate of 0.48 mg NO3--N g-1 soil per day, slightly higher than the CO(NH2)2 treatment,

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whereas the Control had the lowest net nitrification rate of 0.11 mg NO3--N g-1 soil per day. However, the addition of Cd significantly inhibited nitrification in all treatments.

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In comparison with uncontaminated soils, the net nitrification rates of Cd contaminated soils were inhibited by 18.9%, 71.4% and 68.4% in treatments Control, (NH4)2SO4 and

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CO(NH2)2, respectively, lower than the inhibition of ammonia consumption rates

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(89.96%, 89.66% and 92.75%, respectively). With biochar addition, the inhibitory effects of Cd toxicity on nitrification were alleviated. The ammonia consumption rates

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of the Cd contaminated soils increased 1.6 and 13.5 times, following biochar addition in treatments (NH4)2SO4 and CO(NH2)2, and simultaneously, the net nitrification rates

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improved by 13.6 % and 150%, respectively.

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The AOB amoA gene abundance in uncontaminated soils significantly increased with (NH4)2SO4 and CO(NH2)2 fertilizer inputs, from 6.54 × 103 copies g-1 soil initially to 1.52 × 109 and 1.45 × 1010 copies g-1 soil at day 28, then declined to 7.83 × 104 and 5.69 × 105 copies g-1 at day 77, respectively (Fig. 3a). In contrast, the AOB amoA gene copies in Cd contaminated soils were initially, below the limit of detection, then 12 / 45

recovered gradually. Biochar addition promoted the recovery of AOB, except in the (NH4)2SO4 treatment, while the abundance of AOB in the Control and CO(NH2)2 treatments finally increased by 3 and 27 times, respectively. The growth of AOA was stimulated by Cd, with a double amoA gene copy number in Cd treated soils (Fig. 3b). Nitrogen inputs increased the growth of AOA in Cd treated soils, the archaeal amoA gene copy numbers increased from 4.05 × 106 to 8.20 × 106

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and 5.86 × 106 copies g-1 soil in the (NH4)2SO4 and CO(NH2)2 treatments, respectively. However, when combined with biochar, the AOA abundance was significantly reduced

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by 2.4 and 4.1 times in the above treatments, respectively.

Regression analysis indicated a significant positive relationship between nitrate

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concentration and AOA in Cd contaminated soils in the N group. However, in the N-B

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group, nitrate concentrations were positively correlated with AOB (Table 1). In all Cd contaminated soils, soil nitrification was dominated by AOA in the early stage (from

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day 0 to 28), and by AOB in the later stage (from day 28 to 77) (Table 1). The amoA copies of AOA in Cd treated soils were significantly and positively correlated with

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NH4+ and negatively correlated with soil pH, in both the N and N-B groups (Table 2).

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Similarly, NH4+ was also an important factor in stimulating the abundance of AOB in both Cd contaminated (Table 2) and uncontaminated soils. 3.3 Soil acidification and Cd mobilization In the N individual treatments, soil pH was in the order of CO(NH2)2 > Control > (NH4)2SO4 (Fig. 4a). In Cd contaminated soils, the pH in the (NH4)2SO4 treatment 13 / 45

remained stable and increased in the CO(NH2)2 treatment, while in uncontaminated soils, the pH gradually decreased as nitrification proceeded. The exchangeable H+ in the Cd contaminated soils was significantly lower than in soils without Cd (Fig. 4b), while the exchangeable Al remained unchanged (Fig. 4c). Linear regression analysis showed that soil exchangeable H was significantly correlated with NO3−-N in the N group (r = 0.70, p < 0.01) (Fig. 5a). However, this correlation did not occur in the N-B

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group (Fig. 5b). In all N-B treatments, soil pH increased immediately by 0.1 to 0.3 units

(Fig. 4a), then steadily increased further with time. In contrast to pH, soil exchangeable

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Al decreased from 60% initially to 70% finally in the N-B treatments (Fig. 4c). There

was also no significant difference in soil pH, exchangeable Al and H between Cd

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contaminated soils and uncontaminated soils (Fig. 4).

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Pearson correlation analysis indicated that the exchangeable H also had a profound influence on soil pH (r = -0.47, p < 0.05) in the N group (Fig. 6a), whereas, the

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exchangeable Al played a dominant role in the N-B group (r = -0.62, p < 0.01) (Fig. 6b). As mentioned above, soil pH strongly affected available Cd in both the N group

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and the N-B group. Specifically, the exchangeable H and Al may be the respective main

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driver (Fig 5c, 5d and Fig 6). 4. Discussion 4.1 How nitrification and Cd interact in soil with N fertilizers application? Our study explored how N fertilizers affect soil nitrification and soil Cd availability. Many studies have compared different N forms and found that ammonium 14 / 45

based fertilizers such as CO(NH2)2 and (NH4)2SO4 could decrease soil pH, enhance soil Cd bioavailability and phytoextraction [18, 19]. Recent research used nitrification inhibitors to demonstrate the important role of soil nitrification in acidification and Cd transformations in three soil types [35]. In contrast, our results indicated these two ammonium based fertilizers did not increase Cd availability, and CO(NH2)2 even significantly decreased it. Nitrification and its acidification in Cd contaminated soils

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were greatly inhibited, preventing Cd availability. Thus, a negative feedback between Cd and soil nitrification occurred (Fig. 7).

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Soil nitrification was greatly inhibited by Cd, which is consistent with other studies. For example, there was a strong depression of soil nitrification and N metabolism genes

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caused by short-term heavy metal inputs [36]. Long-term exposure to metal stress also

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negatively affected soil nitrification [14]. Heavy metals could inhibit nitrification activity by disrupting proteins once they were transported across bacteria cell

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membranes, where they interacted with protein functional groups [37]. Active multivalent metal cations such as Cd might exchange with essential metals in their

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metabolic sites and inhibited enzyme functions, especially ammonia monoxygenase

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(AMO) [38, 39]. As the key nitrifiers, ammonia oxidizing bacteria (AOB) were more sensitive to heavy metal stress while ammonia oxidizing archaea (AOA) were more resistant [17, 40]. Similarly, Cd strongly inhibited the growth of AOB but stimulated that of AOA in this acidic soil. Archaea have superior physiological and genetic adaptations which permit their survival under metal stress. Their poorly permeable 15 / 45

membranes and effective transport systems act as walls and pumps, eliminating toxins into the cells [41, 42]. Some metals are also required as cofactors for enzymes linked to metabolic processes, such as the transport of electrons for energy production [43], which may explain why AOA is stimulated by Cd. Our data demonstrated that AOB, rather than AOA, dominates soil nitrification. Firstly, we found the inhibition of soil nitrification accompanying with AOB

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elimination in Cd contaminated soils (Fig. 2 and 3). Secondly, the recovery of AOB in

the later stage contributed to regaining its dominance in nitrification (Table 1), while

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nitrification by AOA was always slight in Cd soils (Table 1) even though it was

maintained at a high level (Fig. 3b). Similarly, AOB instead of AOA, restored

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nitrification in a Zn-contaminated soil [44]. The high biomass of AOA was not

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proportional to the rate of nitrification, probably because of its lower specific cell activity (approximately 10-fold lower than AOB) and smaller cell volume [45]. Another

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reason may be that the AOA populations that survive in such extremely acidic and metal toxic conditions often minimize its cellular and metabolic activity by shutting down the

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transcription and translation [46], or are transformed into dormant states or spores [47].

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In addition, these highly tolerant AOA may preferentially obtain energy through other metabolism processes rather than ammonia oxidation [48]. The recovery of AOB occurred from an extremely low base, possibly due to the stimulation of genotypes of Cd-resistant AOB species which have already existed in soil and the selection of the phenotypes of Cd-tolerant species after short-term exposer to Cd [49]. The newly 16 / 45

selected varieties would be more resistant and adaptive than the original ones [43]. The recovery of AOB is linked to the community energy status obtained from ammonia oxidation, and ammonia provision would be beneficial to its recovery (Fig. 3a) [50]. The greater specific cell activity and substrate affinity [45] can also provide AOB with a dominant ability for the recovery of soil nitrification. When soil nitrification was inhibited, the process of acidification ceased (Fig. 4).

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The pH in Cd contaminated soil was significantly higher than in uncontaminated soils. Application of CO(NH2)2 had the highest pH value due to urea hydrolysis [51]. Soil

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exchangeable Al concentrations were unchanged between Cd contaminated and noncontaminated soils, while exchangeable H decreased greatly in Cd contaminated soils,

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where nitrification was greatly inhibited. Soil exchangeable H was strongly correlated

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with nitrate concentrations (Fig. 5a, 6a) and soil pH (Fig. 6a), indicating it could be a sensitive indicator for soil nitrification and acidification, which is consistent with

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previous study [52]. Lu et al [10] also reported that exchangeable H dominated soil acidification in tropical ecosystems after long-term N deposition.

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Although soil nitrification and its acidification were inhibited by Cd, there was a

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significant positive correlation between exchangeable H and available Cd (Fig. 5c), therefore Cd availability may be controlled by H+ after N inputs. Because of its large hydrated ionic radius, H+ provides a strong electrostatic force to desorb Cd2+ from soil particles [53]. It also contributes to the increased solubility of Cd from minerals or precipitates such as Cd(OH)2 [54]. The tested soil was strongly weathered, acidic, and 17 / 45

deficient in organic matter, may resulting in a low pH buffer capacity [10, 55]. Thus soil pH and available Cd showed high sensitivity to exchangeable H in our study (Fig.6a, 5c). Our results therefore demonstrated that soil nitrification rate affects Cd availability though acidification. Cd functioned as a nitrification inhibitor mainly through eliminating AOB, preventing proton production, thus inhibiting Cd itself from

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mobilization. Here, we only set one high Cd concentration and found the negative

feedback. However, it was also reported that nitrification was unaffected or stimulated

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by trace metal at low concentrations [56]. Thus, the feedback may be determined by bioavailable Cd concentration and needs further investigation.

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4.2 How biochar changed the feedback between soil nitrification and Cd availability?

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Following biochar addition, the relationships between soil nitrification, acidification and Cd availability were totally changed. In the presence of biochar, soil

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Cd toxicity greatly declined, while the inhibition of nitrification was alleviated and the nitrification rate was even stimulated. However, the enhanced nitrification did not

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induce acidification after biochar remediation, rather, soil acidity decreased

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continuously over time, accompanying with a continuous decline of Cd availability (Fig. 7).

Biochar addition mitigated the inhibition of Cd on soil nitrification, which is

consistent with increased soil nitrification after short-term biochar manipulation in a metal contaminated soil [57]. We attribute the restoration of nitrification to the recovery 18 / 45

of AOB. In our study, biochar addition decreased AOA but promoted the restoration of AOB (Fig. 3). Nitrate concentration increased with the increasing AOB in Cd soils following N-B application (Table 1). Biochar inputs changed the niches of AOA and AOB of Cd contaminated soils. Initially, biochar significantly and quickly immobilized Cd when applied to soil (Fig. 1), thus reducing the survival stress and enlarging the reselection niches for AOB that

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are more susceptible. Biochar also improves soil characteristics such as pH and nutrient

availability, and the improved soil conditions may preferentially select AOB rather than

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AOA [58]. The AOA that adapt and flourish under extreme conditions such as acid pH,

heavy metals or extreme temperature, however, lose their competitive advantages when

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the environment becomes more benign [59]. Unlike AOA, which utilizes ammonium

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(NH4+) as the N substrate for oxidation, AOB uses ammonia (NH3) as the major energy source [45]. As for carbon metabolism, AOB fix CO2 [60] whereas AOA fix HCO3-

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[61]. Higher pH contributes to NH4+ being converted into NH3 and HCO3- being transformed into CO32-, decreasing accessible C and N sources for AOA, while

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providing a more favorable habitat for AOB. Biochar could also absorb ammonia and

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subsequently provides N continuously for AOB, like a controlled release fertilizer [62]. This is also supported by our data showing that AOB in uncontaminated soils applied with (NH4)2SO4 and CO(NH)2 declined considerably following ammonia consumption, but this decline could be alleviated by biochar (Fig. 3a). Thus, biochar promotes AOB

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growth by decreasing soluble metals and acidity, as well as providing more available and sustainable substrates. Enhanced nitrification did not cause soil acidification after biochar remediation. The exchangeable H showed neither regular change (Fig. 4b), nor a relationship with soil nitrification (Fig. 5b). While the exchangeable Al declined considerably with the increasing soil pH (Fig. 4c). One possible reason is that the liming effects and buffering

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brought by biochar not only neutralizes protons produced by nitrification, but alleviate

Al toxicity and produce a sustainable increase in soil pH. The liming effect is due to the

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alkalinity of biochar [63]. These major alkaline components such as basic compounds

(e.g. carbonates or oxides) and organic functional groups (e.g. —O, —O-, —COO-),

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could react with acidic ions such as H+ and Al3+ once the biochar is added to soil [63,

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64], rapidly increasing soil pH and decreasing exchangeable Al. Except for these direct mechanisms for short-term acidity alleviation, the absorption and buffering of biochar

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played more important roles in long-term resistance to acidification [65, 66]. High surface area and porosity of biochar provide more adsorption sites for sustainable

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alleviation of Al toxicity [64]. Application of biochar increased pH buffering capacity

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of acidic soil mainly through the protonation of its carboxyl functional groups, thus increasing the proton absorption volume of soil [67]. Additionally, with the alkaline materials released during biochar aging process, the enhanced soil buffering may also cause the process sustainable and slow. The slow liming effects and great buffer capacity of biochar not only effectively reduced soil acidity, but prevented soil re20 / 45

acidification caused by nitrification. Soil available Cd varied consistently with the exchangeable Al (Fig. 5d), and in contrast to soil pH (Fig. 6). Similarly to the fate of Al, Cd availability decreased by nearly half immediately after biochar addition (Fig. 1). The carbonates, oxides and functional groups of biochar could directly immobilize Cd through ion exchange, complexation and precipitation, which are responsible for short-term (2 days) metal

ro of

stabilization [68]. Long-term immobilization in acidic soil was attributed to the

consistently improved soil pH after biochar addition [63, 69]. Available Cd in our soils

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also gradually decreased with the slow liming effects of biochar.

Thus, the feedback of soil nitrification and Cd availability was totally changed

re

following biochar addition. Biochar promoted the recovery of AOB and nitrification by

lP

decreasing Cd toxicity and improving soil habitats. Meanwhile, its specific buffering

nitrification.

na

and liming traits also resisted the re-acidification and Cd mobilization caused by

5. Conclusions

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In this study, we have discussed the interaction between soil nitrification and metal

Jo

through both chemical and biological processes simultaneously, and determined a negative feedback for the first time. Biochar combined with nitrogen provides a feasible strategy for effectively removing Cd and Al, alleviating their environmental biotoxicity and resisting the re-acidification caused by N application. This study is only limited to one Cd concentration and different concentrations of Cd should be considered in further 21 / 45

study. Besides, the eco-toxicity of metal on soil nitrifiers was limited in quantitative copies of functional gene, maybe the community and transcription level need to be further investigated. Consequently, more attention should be applied to the remediation effects of biochar on microbial taxonomy diversity, metabolism and function in metal contaminated sites. Acknowledgements

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This work was financially supported by the National Natural Science Foundation of China (41721001), the Science and Technology Program of Zhejiang Province

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(2018C03028), the 111 Project (B17039), and Agriculture Research System of China (CARS-01-30).

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Fig. 1. Dynamic variations of available Cd (CaCl2 extractable) in contaminated soils during the incubation. Error bars indicate standard deviation; n=3; p< 0.05.

Fig. 2. Dynamics of (a) NH4+-N and (b) NO3--N concentration during the incubation (Cd 0: uncontaminated soils; Cd 5: Cd contaminated soils; B0: without biochar; B1:

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with biochar). Error bars indicate standard deviation; n=3; p< 0.05.

Fig. 3. Abundance of (a) AOB and (b) AOA at day 0, 28 and 77 (Cd 0: uncontaminated

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soils; Cd 5: Cd contaminated soils; B0: without biochar; B1: with biochar). NA signifies lower than the detection limit. Different letters indicate significant difference between

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sampling soils. Error bars indicate standard deviation; n=3; p< 0.05.

Fig. 4. Dynamics of (a) soil pH during the incubation and (b) soil exchangeable H+

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significant difference between sampling soils. Error bars indicate standard deviation;

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n=3; p< 0.05.

Fig. 5. Correlation analysis between exchangeable H and NO3--N in soils in (a) N group and (b) N-B group; Correlation between soil available Cd and (c) exchangeable H in N group and (d) exchangeable Al in N-B group respectively (Cd contaminated soils only). 34 / 45

Fig. 6. Simple Pearson correlation of soil properties and available Cd in (a) N group and (b) N-B group. The upper part represents the coefficient, color red and blue indicates positive (+) and negative (-) correlation respectively, deeper colors indicates stronger correlations, asterisks indicates significant effects (*P < 0.05, **P < 0.01 and ***P < 0.001). The lower part contains the scatterplots of each two parameters, middle

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part represents the frequency distribution.

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Fig. 7. The feedback mechanism between soil nitrification and Cd availability. In N treatments, soil nitrification affects Cd availability through acidification. However, Cd

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functioned as a nitrification inhibitor, preventing AOB growth and protons (H+)

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production, thus inhibiting itself form mobilization. When combined with biochar (NB treatments), the decreased Cd toxicity and the improved soil conditions accelerate the

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recovery of AOB and nitrification. However, the enhanced nitrification cannot induce acidification due to the slow liming effects of biochar, which not only buffers the

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protons produced by nitrification, but also absorbs Cd continuously. (Red dashed line

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indicates the process being inhibited, while blue line indicates a smooth process).

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ro of

-p

re

lP

na

ur

Jo Fig 5.

40 / 45

ro of

-p

re

lP

na

ur

Jo Fig 6.

41 / 45

42 / 45

ro of

-p

re

lP

na

ur

Jo

ro of

-p

re

lP

na

ur

Jo Fig 7.

43 / 45

Table 1. Multiple linear regression analysis of NO3--N and soil nitrifiers in Cd contaminated soil.

Function

R2

Regression equations

F value

P value

15.25

1.33×10-4

Pr(>|t|)

Treatments NO3--N = -2.516(±2.86) + 1.91(±0.62) ×106 ×AOA +

N-B group

--N

NO3

2.78(±0.14)

×10-7

× AOB

0.587

= 0.675(±2.24) + 8.41(±9.13)

× AOA + 1.17(±0.15)

×10-5

×10-7

× AOB

0.751

31.09

Time NO3--N = -0.15(±0.38) + 8.51(±1.00) ×10-7 × AOA + 1.81(±0.91)

28-77 d

--N

NO3

×10-5

× AOB

= 4.70(±1.49) + 1.43(±3.85)

× AOA + 6.28(±1.21)

×10-6

0.781

36.58

×10-7

× AOB

0.461

Jo

ur

na

lP

re

For T values, * p < 0.05, **p < 0.01, ***p < 0.001, **** p < 0.0001.

15.94

44 / 45

4.56×10-7

1.32×10-5

-p

0-28 d

1.45×10-6

Intercept

0.391

AOA

0.006***

AOB

0.840

Intercept

0.766

AOA

0.368

AOB

3.21×10-7****

ro of

N group

Intercept

0.696

AOA

1.04×10-7****

AOB

0.060*

Intercept

0.003**

AOA

0.713

AOB

1.1×10-5***

Table 2. Multiple linear regression analysis of the soil nitrifiers and soil properties in Cd contaminated soil.

Niche

N group

R2

Regression equations

F value

P value

AOA = 2.15 (±0.38) × 107 + 4.40 (±0.47) × 105 × NH4+-N – 4.02 (±0.84) × 106 × pH

0.813

44.58

1.06×10-7

AOB = 6.96 (±3.02) × 106 + 1.54 (±0.37) × NH4

– 1.60 (±0.68) ×

106 ×

pH

0.428

0.455

AOB = -1.87 (±3.09) × 106 + 9.56 (±3.97) × +-N

NH4

+ 3.90 (±6.56) ×

105 ×

pH

lP

AOA = 3.30 (±0.40) × 107 + 3.76 (±0.59) ×

0.298

na

105 × NH4+-N – 6.57 (±0.87) × 106 × pH

9.34

5.25

re

103 ×

All

8.47

AOA = 2.00 (±0.44) × 107 + 3.07 (±5.70) × 103 × NH4+-N – 3.70 (±0.94) × 106 × pH

0.627

2.5

1.65×10-3

35.43

0.016

1.70 ×10-9

AOB = 9.80 (±17.8) × 105 + 1.11 (±0.26) × +-N

NH4

– 2.31 (±3.86) ×

105 ×

pH

NH4+-N

2.19×10-8 ***

pH

1.53×10-4 ***

+-N

0.299

Jo

For T values, * p < 0.05, **p < 0.01, ***p < 0.001, **** p < 0.0001.

45 / 45

9.75

0.033 *

NH4

6.52×10-4 ***

pH

0.030 *

Intercept

2.76×10-4 ***

NH4+-N

0.597

pH

9.65×10-4 ***

Intercept +-N

3.68

×10-4

0.554

NH4

0.027 *

pH

0.560

Intercept

4.33×10-10 ***

NH4+-N

1.35×10-7 ***

pH

3.54×10-9 ***

Intercept

ur

105 ×

1.96×10-5 ***

Intercept ×10-3

-p

N-B group

+-N

Intercept

ro of

104 ×

Pr(>|t|)

+-N

0.586

NH4

1.32×10-4 ***

pH

0.553