## 1 Introduction  
The acceleration of modernization will trigger serious soil pollution problems [1, 2, 3, 4]. Taking the Xiangtan manganese mine as an example, the manganese (Mn) content in the neighboring farmland has seriously exceeded the standard due to the mining activities and improper treatment of tailings and is accompanied by the pollution of other heavy metals [5, 6, 7]. A study showed that the Mn content in the farmland around the Mn mine exceeded the standard by 16.3 times, and, at the same time, the content of lead (Pb) and copper (Cu) also exceeded the standard by 15.4 times and 15.0 times, respectively [8]. In addition, the heavy metal (HM) content at the soil-water interface in the Mn mining area was more affected by Cu pollution than Mn pollution [9]. This complex and multifaceted pollution seriously threatens soil fertility and food security production. Therefore, there is an urgent need to remediate Mn-Cu contamination in agricultural lands around Mn mining areas and restore soil health to ensure food security. Biochar is an organic material produced by the pyrolysis process under high temperature and oxygen-limited conditions [10]. Its surface is rich in a variety of functional groups, which can effectively immobilize HM in the soil and thus inhibit the migration of HM to crops to a certain extent [11; 12; 13]. A study reported that the application of tobacco biochar in a lettuce field (5 t hm\({}^{-1}\)) resulted in a 23% reduction in the bioactive state of Cu in the soil and a 26.4% reduction in Cu concentration in the aboveground portion of lettuce [14]. Similarly, Mn concentration in spinach was significantly reduced by 78.0% after the application of broadleaf wood biochar to Mn-complex contaminated soil [15]. Furthermore, biochar can release dissolved organic matter (DOM) due to its porous structure and large specific surface area [16]. DOM, as the most active form of organic carbon in the soil, has a complex composition containing carboxyl, hydroxyl, aromatic ring, double bond, and other rich functional groups [17; 18]. These functional groups can mediate the transport of HMs in soil. It was shown that there are differences in the functional groups of DOMs from different sources in their response to HMs. For instance, DOM derived from corn stalk undergoes static quenching in Pb solution, where the amide group binds to Pb first, followed by phenolic hydroxyl groups and C-O-O telescopic deformation to produce a quenching response [19]. [PERSON] et al. found that under the background of Cd and Cu contamination, the binding responses of phenolic hydroxyl and carboxyl groups in compost-DOM were the strongest, while the amide and phenolic hydroxyl groups in rice straw-DOM responded the fastest [16]. This indicates that the DOM and HM responses are different mainly related to functional groups. Moreover, the diversity of functional groups depends on the DOM chemical composition. Therefore, DOM substance composition directly affects its binding effect with HM. The results of a large number of studies have shown that DOM is mainly composed of humic-like substances (e.g., humic acid or xanthic acid) and protein-like substances (e.g., tyrosine and tryptophan) [20; 21]. Among them, humic-like substances, especially humic acid (HA), have large molecular weights and are rich in functional groups that can form stable complexes with HM ions [22; 23]. Therefore, their binding capacity with HM far exceeds that of protein-like substances. In summary, the higher the proportion of HA in DOM, the more efficiently it binds to HM, and the complex formed is more stable, with a corresponding decrease in mobility. It is worth noting that HA, as a kind of macromolecular organic matter, not only enhances soil fertility but also regulates the adsorption of biochar on soil DOM [24; 25]. Therefore, the following questions are worth exploring in depth: how will the strategy of combining HA with biochar affect the compositional composition of DOM?